The World Health Organization is a specialized agency of the United Nations with primary responsibility for international health matters and public health. Through this organization, which was created in 1948, the health professions of some 160 countries exchange their knowledge and experience with the aim of making possible the attainment by all citizens of the world by the year 2000 of a level of health that will permit them to lead a socially and economically productive life. By means of direct technical cooperation with its Member States, and by stimulating such cooperation among them, WHO promotes the development of comprehensive health services, the prevention and control of diseases, the improvement of environmental conditions, the development of health manpower, the coordination and development of biomedical and health services research, and the planning and implementation of health programmes. These broad fields of endeavour encompass a wide variety of activities, such as developing systems of primary health care that reach the whole population of Member countries; promoting the health of mothers and children; combating malnutrition; controlling malaria and other communicable diseases including tuberculosis and leprosy; having achieved the eradication of smallpox, promoting mass immunization against a number of other preventable diseases; improving mental health; providing safe water supplies; and training health personnel of all categories. Progress towards better health throughout the world also demands international cooperation in such matters as establishing in(emational standards for biological substances, pesticides and pharmaceuticals; formulating environmental health criteria; recommending international nonproprietary names for drugs; administering the International Health Regulations; revising the International Classification of Diseases, Injuries, and Causes of Death; and collecting and disseminating health statistical information. Further information on many aspects of WHO's work is presented in the Organization's publications.
GUIDELINES FOR DRINKING-WATER QUALITY Volume 2 Health Criteria and Other Supporting Information
World Health Organization Geneva 1984
ISBN 92 4 154169 5 (( World Health Organization 1984 Publications of the World Health Organization enjoy copyright protection in accordance with the provisions of Protocol 2 of the Universal Copyright Convention. For rights of reproduction or translation of WHO publications, in part or in toto, application should be made to the Office of Publications, World Health Organization, Geneva, Switzerland. The World Health Organization welcomes such applications. The designations employed and the presentation of the material in this publication do not imply the expression of any opinion whatsoever on the part of the Secretariat of the World Health Organization concerning the legal status of any country, territory, city or area or of its authorities, or concerning the delimitation of its frontiers or boundaries. The mention of specific companies or of certain manufacturers' products does not imply that they are endorsed or recommended by the World Health Organization in preference to others of a similar nature that are not mentioned. Errors and omissions excepted, the names of proprietary products are distinguished by initial capital letters.
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CONTENTS Page Preface. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PART I. 1.
vn
MICROBIOLOGICAL ASPECTS 3 3 4 9 10 12 15 29 32 32 32 33 33 34 35 35
The bacteriological quality of drinking-water . . . . . . . . . .
1.1 1.2 1.3 1.4 1.5
Waterborne bacterial pathogens. . . . . . . . . . . . . . . . . Rationale for the use of indicator organisms. . . . . . . Nuisance organisms. . . . . . . . . . . . . . . . . . . . . . . . . . Disinfection. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Collection, storage and transport of water samples for bacteriological examination . . . . . . . . . . . . . . . . . . . . 1.6 Recommended methods for the detection and enumeration of organisms indicative of pollution. . . . . . . . . virological quality of drinking-water . . . . . . . . . . . . .
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2. The 2.1 2.2 2.3 2.4 2.5 2.6
General description . . . . . . . . . . . . . . . . . . . . . . . . . Routes of exposure. . . . . . . . . . . . . . . . . . . . . . . . . Health effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Rationale for recommendation ................. Methods of examination for viruses . . . . . . . . . . . . Interpretation and. evaluation of positive findings . .
. . . , . .
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
PART II. 1. 2. Protozoa
BIOLOGICAL ASPECTS 39
Helminths . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1 Group I (Dracunculus, Spirometra) ............. . 2.2 Group II (Schistosoma, Ancylostoma, Necator) . .... . 2.3 Group III (Ascaris, Trichuris, Strongyloides, Enterobius, Fasciolids, Hymenolepis, Echinococcus) . .......... . Free-living organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . .
46 46
49 51 53 56
3.
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ill
IV
CONTENTS
PART III.
HEALTH-RELATED INORGANIC CONSTITUENTS 63 68 76 80 84 91 97 100 106 Ill 120 124 128 135 141 145
1. 2. 3. 4.
Arsenic. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Asbestos. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Barium. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Beryllium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chromium. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Fluoride . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Hardness . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mercury. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nickel . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nitrate and nitrite . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Selenium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Silver. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sodium. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
5. 6. 7. 8. 9.
10. 11. 12. 13. 14. 15. 16.
PART IV.
HEALTH-RELATED ORGANIC CONSTITUENTS 155 155 159 162 162 166 170 176 182 190 191 194 197 203 208
1.
Chlorinated alkanes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.1 Carbon tetrachloride . . . . . . . . . . . . . . . . . . . . . . . . . 1.2 1,2-Dichloroethane . . . . . . . . . . . . . . . . . . . . . . . . . . Chlorinated ethenes . . . 2.1 Vinyl chloride . . . . 2.2 1,1-Dichloroethene. 2.3 Trichloroethene. . . 2.4 Tetrachloroethene . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2.
3. 4.
Polynuclear aromatic hydrocarbons. . . . . . . . . . . . . . . . . . Pesticides. . . . . . . . . . . . . . . . . 4.1 DDT (total isomers) . . . . 4.2 Aldrin and dieldrin . . . . . . 4. 3 Chlordane. . . . . . . . . . . . . 4.4 Hexachlorobenzene . . . . . . 4.5 Heptachlor and heptachlor ...... ...... ...... ...... ...... epoxide
CONTENTS
V
4.6 Lindane . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4. 7 Methoxychlor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.8 2,4-Dichlorophenoxyacetic acid . . . . . . . . . . . . . . . . 5. Chlorobenzenes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1 Chlorobenzene (monochlorobenzene) . . . . . . . . . . . . . 5.2 Dichlorobenzenes. . . . . . . . . . . . . . . . . . . . . . . . . . . . Benzene and lower alkyl benzenes . . . . . . . . . . . . . . . . . . Phenol and chlorophenols. . 7.1 Chlorinated phenols of 7.2 2,4,6-Trichlorophenol. . 7.3 Pentachlorophenol. . . . .......... toxicological .......... .......... ......... significance ......... ......... . . . . . . . . . . . . . . . .
212 217 218 221 222 224 229 233 233 236 237 240
6. 7.
8.
Trihalomethanes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
PART V. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. 13.
AESTHETIC CONSTITUENTS AND CHARACTERISTICS 249 253 256 262 264 268 272 275 279 281 286 290 293 293 297 301 304 307 313
Aluminium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chloride. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Colour . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Copper . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Hardness' . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Hydrogen sulfide. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Iron . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Manganese. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Oxygen, dissolved. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . pH level. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sodium. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sulfate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Taste and odour. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.1 Taste. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.2 Odour. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Temperature. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Total dissolved solids . . . . . . . . . . . . . . . . . . . . . . . . . . . Turbidity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Zinc . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
14. 15. 16. 17.
VI
CONTENTS
PART VI. 1. 2.
RADIOACTIVE MATERIALS 319 320 320 321 322 322 323 325 325 327
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Basic considerations . . . . . . . . . . . . 2.1 Dose-response relationship. . . . 2.2 Dose-equivalent limits . . . . . . . 2.3 Implications for drinking-water ...... ...... ...... quality . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3. 4. 5.
Sources of radiation exposure . . . . . . . . . . . . . . . . . . . . . Guideline value for gross alpha and gross beta activity . . Radon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Index. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
PREFACE The last edition of the International standards for drinking-watera was issued in 1971 and that of the European standards for drinking-waterb in 1970. These standards have now been reviewed, revised, and combined, the resulting publication being issued in three separate volumes under the title Guidelines for drinking-water quality. In brief, the three volumes contain the following: Volume 1: Recommendations. This volume presents recommended guideline values per se, together with the information essential to understanding the basis for the recommended guideline values as well as information on monitoring requirements. Where possible, suggestions are included regarding remedial measures to ensure compliance with the guideline values. The guidelines cover the microbiological, biological, chemical, organoleptic, and radiological quality of drinking-water. Volume 2: Health criteria and other supporting information. The second volume sets out the health criteria for those drinking-water pollutants and other constituents that were examined with a view to recommending guideline values. In addition, it provides information regarding the detection of contaminants in water and measures for their control. It contains a review of the toxicological, epidemiological, and clinical evidence that was available and used in deriving the recommended guideline values. Volume 3: Guidelines for drinking-water quality control for smallcommunity supplies. The final volume deals specifically with the problem of small communities mainly located in rural areas. It contains information on techniques for the assessment and control of contamination of water supplies to these communities, including simple methods for sampling and analysis, sanitary surveys, and other means of investigating and controlling drinking-water quality in these areas. This volume is concerned essentially with the bacteriological safety of water.
Volume 2 thus reviews the evidence for recommending guideline values, summarizes and evaluates the available information on the health and sensory effects of drinking-water constituents, and provides a convenient reference source for those involved in developing and implementing national standards, as well as for those engaged in research work. This volume elaborates greatly on the health risk information presented in volume 1 and should be considered as a vital companion document to it. b
• International standards for drinking-water, 3rd ed. Geneva, World Health Organization, 1971. European standards for drinkmg-water, 2nd ed. Geneva, World Health Orgamzation, 1970.
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GUIDELINES FOR DRINKING-WATER QUALITY
The microbiological and biological aspects of drinking-water quality are covered in Parts I and II of this volume. In addition to a description of the waterborne bacterial pathogens, including the rationale for using indicator organisms, detailed information is provided on the surveillance requirements and possibilities for safeguarding the bacteriological quality of drinking-water supplies. Subsections deal with the collection, storage, and transport of water samples, recommended methods for the detection of various microorganisms, and disinfection practice. A brief account of viruses in drinking-water is included as well. While no guideline values as such are put forward for biological contaminants, the presence of pathogenic protozoans and helminths is of concern. Free-living organisms that may occur in water supplies are also discussed. No guideline values are recommended since standard methodology for surveillance is lacking in this field. Therefore, emphasis is placed on general source protection to minimize health problems from biological agents in drinking-water supplies. A wide range of health-related inorganic and organic chemicals were considered by the task groups that met in the process of preparing the guidelines. Of the 37 inorganic and the 46 organic chemicals they reviewed in detail, guideline values were set for 9 inorganic and 15 organic chemicals, while tentative guideline values were set for 3 organic chemicals. The information on health effects considered in the development of the guideline values, together with other supporting data, is summarized in Parts III and IV of this volume. In addition, summaries are also included for some of the other chemicals and constituents that were reviewed but for which, for various reasons, it was not considered appropriate to recommend health-based guideline values. These summaries relate to asbestos, barium, beryllium, hardness, nickel, nitrite, silver, sodium, vinyl chloride, and certain chlorophenols and chlorobenzenes. The information contained in these summaries usually includes the following: (a) a general description of the chemical, its major sources and concentrations in water; (b) information on routes of human exposure (drinking-water, food, air, etc.) including their relative significance; (c) metabolism (absorption, distribution, retention, elimination, and biotransformation); and (d) evidence of health effects, which includes a description of the adverse biological effects observed and an assessment of their health significance, identification of the most sensitive population groups at risk, relationships between dose and effects, and the prevalence of the effects. Each summary contains a list of the relevant references. A total of 22 constituents and characteristics of drinking-water were carefully examined to ascertain their influence on its aesthetic and organoleptic qualities. Guideline values were recommended for 15 of them. The basic information used in arriving at the guideline values is summarized in this volume with reference to the sources, occurrence, routes of exposure, and the health and other effects. These summaries
PREFACE
ix
will be found in Part V, which also provides information on the influence of temperature, dissolved oxygen, and pH on drinking-water quality. Part VI deals with radioactive materials in drinking-water and was prepared in close cooperation with the International Commission on Radiological Protection (ICRP). It explains the basis for the guideline values for gross alpha and gross beta activity and gives guidance with respect to the application of these guideline values in practice. The preparation of volumes I and 2 of the new guidelines took over three years and involved the active participation of nearly 30 WHO Member States, scores of scientists, and meetings of IO task groups. The work of these institutions and scientists, whose names appear in Annex I to volume I, was central to the successful completion of the guidelines and is much appreciated. The collaboration of the national focal points for the WHO Environmental Health Criteria Programme, various international organizations, and individual experts was most helpful and their continuing participation contributed effectively to the work. The coordinators for this work were Dr H. Gala] Gorchev of WIIO headquarters and Mr W. Lewis of the WHO Regional Office for Europe. The preparation of these guidelines was made possible by the financial support afforded to WHO by the Danish International Development Agency (DANIDA) and by the United Nations Environment Programme (UNEP), which provided funds to offset the costs of publishing this volume. Their contribution is most gratefully acknowledged. Appreciation is also expressed to the Environmental Protection Agency (EPA) in the USA which supported this effort by secondment of Dr Gala] Gorchev for a period of two years. It is recognized that as new information becomes available, the basis of the recommended guideline values will need to be reviewed and revised and new or changed guideline values may have to be recommended. As regards chemicals in drinking-water supplies, such reviews will be facilitated in the future through the International Programme on Chemical Safety (IPCS), which is a cooperative venture of the United Nations Environment Programme, the International Labour Organisation (ILO), and WHO. Two of its main objectives are (a) the evaluation of the effects of chemicals on human health and the environment, and (b) the development of guidelines for exposure limits (such as acceptable daily intakes and maximum permissible or desirable levels in air, water, food, and the working environment) for various classes of chemicals, including food additives, industrial chemicals, toxic substances of natural origin, plastics, packing materials, and pesticides. In particular, it was already recognized by the final task group meeting that new evidence regarding the potential health hazards from asbestos, sodium, nitrate/nitrite, nickel, chloroform, other trihalomethanes, trichioroethene, tetrachloroethene, and carbon tetrachloride was likely to become available over the coming years and would necessitate a reconsideration of these substances.
PART I. MICROBIOLOGICAL ASPECTS
1. THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER The most common and widespread danger associated with drinkingwater is contamination, either directly or indirectly, by sewage, by other wastes, or by human or animal excrement. If such contamination is recent, and if among the contributors there are carriers of communicable enteric diseases, some of the living causal agents may be present. The drinking of water so contaminated or its use in the preparation of certain foods may result in further cases of infection.
1.1 Waterborne bacterial pathogens
Faecal pollution of drinking-water may introduce a variety of intestinal pathogens-bacterial, viral, and parasitic-their presence being related to microbial diseases and carriers, present at that moment in the community. Intestinal bacterial pathogens are widely distributed throughout the world. Those known to have occurred in contaminated drinking-water include strains of Salmonella, Shigella, enterotoxigenic Escherichia coli, Vibrio cholerae, Yersinia enterocolitica, and Campylobacter fetus. These organisms may cause diseases that vary in severity from mild gastroenteritis to severe and sometimes fatal dysentery, cholera, or typhoid. Other organisms, naturally present in the environment and not regarded as pathogens, may also cause occasional opportunist disease. Such organisms in drinking-water may cause infection predominantly among people whose local or general natural defence mechanisms are impaired; this is most likely to be the case in the very old, the very young, and patients in hospitals, for example, with burns or on immunosuppressive therapy. Potable water used by patients for drinking and bathing, if it contains excessive numbers of organisms such as Pseudomonas, Flavobacterium, Acinetobacter, Klebsiella, and Serratia, may produce a variety of infections involving the skin and mucous membranes of the eye, ear, nose, and throat. The modes of transmission of bacterial pathogens include ingestion of contaminated water and food, contact with infected persons or animals, and exposure to aerosols. The significance of the water route in the spread of intestinal bacterial infections varies considerably, both with the disease and with local circumstances. Although Shigella may be waterborne, water is not usually the main route for the spread of 3
4
I.
MICROBIOLOGICAL ASPECTS
shigellosis, but rather person-to-person contact in crowded living conditions; in contrast, cholera is usually waterborne and salmonellosis food-borne. Among the various waterborne pathogens, there exists a wide range of minimum infectious dose levels necessary to cause a human infection. With Salmonella typhi, ingestion of relatively few organisms can cause disease; with Shigella jlexneri, several hundred cells may be needed, whereas many millions of cells of Salmonella serotypes are usually required to cause gastroenteritis. Similarly, with toxigenic organisms such as enteropathogenic E. coli and V. cholerae, as many as 10 8 organisms may be necessary to cause illness. The size of the infective dose also varies in different persons with age, nutritional status, and general health at the time of exposure. The significance of routes of transmission other than drinking-water should not be underestimated as the provision of a safe potable supply by itself will not necessarily prevent infection without accompanying improvements in sanitation and personal habits. Education in simple applied hygiene is essential.
1.2 Rationale for the use of indicator organisms The recognition that microbial infections can be waterborne has led to the development of methods for routine examination to ensure that water intended for human consumption is free from excremental pollution. Although it is now possible to detect the presence of many pathogens in water, the methods of isolation and enumeration are often complex and time-consuming. It is therefore impracticable to monitor drinking-water for every possible microbial pathogen that might occur with contamination. A more logical approach is the detection of organisms normally present in the faeces of man and other warmblooded animals as indicators of excremental pollution, as well as of the efficacy of water treatment and disinfection. The presence of such organisms indicates the presence of faecal material, and thus that intestinal pathogens could be present. Conversely, the absence of faecal commensal organisms indicates that pathogens are probably also absent. Search for such indicators of faecal pollution thus provides a means of quality control. Surveillance of the bacterial quality of raw water is also important, not only in the assessment of the degree of pollution, but also in the choice of the best source and the treatment needed. Bacteriological examination offers the most sensitive test for the detection of recent and therefore potentially dangerous faecal pollution, thereby providing a hygienic assessment of water quality with a sensitivity and specificity that is absent from routine chemical analyses. It is essential that water is examined regularly and frequently as contamination may be intermittent and may not be detected by the examination of a single sample. For this reason, it is important that drinking-water is examined frequently by a simple test rather than infrequently by a more complicated test or series of tests. Priority must always be given to
1.
THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
5
ensuring that routine bacterial examination is maintained whenever manpower and facilities are limited. It must be appreciated that all a bacteriological analysis can prove is that, at the time of examination, contamination, or bacteria indicative of faecal pollution, could or could not be demonstrated in a given sample of water using specified culture methods. In addition, the results of routine bacteriological examination must always be interpreted in the light of a thorough knowledge of the water supplies, including their source, treatment, and distribution. Whenever changes in conditions lead to deterioration in the quality of the water supplied, or even if they should suggest an increased possibility of contamination, the frequency of bacteriological examination should be increased, so that a series of samples from well chosen locations may identify the hazard and allow remedial action to be taken. Whenever a sanitary survey, including visual inspection, indicates that a water supply is obviously subject to pollution, remedial action must be taken, irrespective of the results of bacteriological examination. For unpiped rural supplies, sanitary surveys may often be the only form of examination that can be undertaken regularly. 1.2.1 Organisms indicative of faecal pollution
The use of normal intestinal organisms as indicators of faecal pollution rather than the pathogens themselves is a universally accepted principle for monitoring and assessing the microbial safety of water supplies (1). Ideally, the finding of such indicator bacteria should denote the possible presence of all relevant pathogens. Indicator organisms should be abundant in excrement but absent, or present only in small numbers, in other sources; they should be easily isolated, identified and enumerated and should be unable to grow in water. They should also survive longer than pathogens in water and be more resistant to disinfectants, such as chlorine. In practice, these criteria cannot all be met by any one organism, although many of them are fulfilled by coliform organisms, especially Escherichia coli as the essential indicator of pollution by faecal material of human or animal origin. Other microorganisms that satisfy some of these criteria, though not to the same extent as coliform organisms, can also be used as supplementary indicators of faecal pollution in certain circumstances. The significance that can be attached to the presence or absence of particular faecal indicators varies with each organism and especially with the degree to which that organism can be specifically associated with faeces. Organisms used as bacterial indicators of faecal pollution include the coliform group of organisms as a whole, E. coli and coliform organisms that have been described as "faecal coliforms", faecal streptococci, and sulfite-reducing clostridia, especially Clostridium perfringens. Anaerobic bacteria, such as bifidobacteria and Bacteroides, are
6
I.
MICROBIOLOGICAL ASPECTS
more abundant than coliform organisms in faeces, but routine methods for their detection and enumeration are not yet available. The other groups of organisms also have non-faecal sources in the environment, and may even grow in the aquatic environment, thus decreasing the confidence with which their presence may be associated with excremental pollution. Full identification of these indicator organisms would require an extensive series of tests, which would be impracticable in routine monitoring. Water bacteriologists have therefore evolved definitions of indicator species and groups that are practical rather than taxonomic and are based largely on detection and enumeration in water, usually by either multiple-tube methods or membrane-filtration techniques. 1.2.1.1 Coliform organisms (total co/iforms) Coliform organisms have long been recognized as a suitable microbial indicator of drinking-water quality, largely because these organisms are easy to detect and enumerate in water. They are characterized broadly by their ability to ferment lactose in culture at 35oC or 37°C, and include E. coli, Citrobacter, Enterobacter, and Klebsiella species. Coliform bacteria should not be detectable in treated water supplies, and if found, suggest inadequate treatment or post-treatment contamination (2). In this sense, the coliform test is used as an indicator of treatment efficiency. Although coliform organisms may not be directly related to the presence of viruses in drinking-water, the use of the coliform test is still essential for monitoring the microbial quality of public water supplies (3). The cysts of some parasites are known to be more resistant than coliform organisms to disinfection. The absence of coliform organisms in surface water that has only been disinfected will not necessarily indicate freedom from the cysts of Giardia, amoebae, and other parasites. Furthermore, coliform bacteria are derived not only from the faeces of warm-blooded animals but also from vegetation and soil (4-6). Under certain conditions, coliform organisms may also persist on nutrients derived from non-metallic construction materials. For these reasons, the presence of small numbers of coliform organisms (1-10 organisms per 100 ml), particularly in untreated groundwater, may be of limited sanitary significance provided faecal coliform organisms are absent. 1.2.1.2 Faecal ( thermotolerant) coliform organisms These are coliform organisms that are able to ferment lactose at 44.0 oC or 44.5 oC; they comprise the genus Escherichia and to a lesser extent occasional strains of Enterobacter, Citrobacter, and Klebsiella. Of these organisms, only E. coli is specifically of faecal origin, being always present in the faeces of man, animals, and birds in large numbers, and rarely found in water or soil that has not been subject to faecal pollution. Complete identification of E. coli in terms of modern taxonomy would require an extensive series of tests, which would be
1.
THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
7
impracticable for routine water examination. Therefore, detection and identification of these organisms as faecal organisms or presumptive E. coli is considered to provide sufficient information to assess the faecal nature of pollution. Regrowth of faecal coliform organisms (7) in the distribution system is unlikely unless sufficient bacterial nutrients are present (biochemical oxygen demand (BOD) greater than 14 mg/1), water temperature is above 13 °C, and there is no free chlorine residual.
1.2.2 Other indicators of faecal pollution If there is any doubt, especially when coliform organisms are found in the absence of faecal coliform organisms and E. coli, other indicator organisms may be used to confirm the excremental nature of the contamination. These secondary indicator organisms include the faecal streptococci and sulfite-reducing clostridia, especially C. perfringens. 1.2.2.1 Faecal streptococci Tlu: occurrence of faecal streptococci m water generally indicates faecal pollution (8, 9). This term refers to those streptococci normally present in the faeces of man and animals. It includes S. faecalis, S. faecium, S. durans, S. bovis, and S. avium, as well as strains with properties intermediate between them. These organisms rarely multiply in polluted water and they may be slightly more resistant to disinfection than coliform organisms. However, this indicator group has rarely been recommended for control of drinking-water quality because of their persistence in water with moderate salt concentration (10), such as might occur with blended water supplies. Furthermore, the widespread occurrence of S. faecalis var. liquifaciens may detract from the significance of numbers of faecal streptococci less than 100 per 100 ml in drinking-water, unless strain identification is part of the routine procedure. When used as a supplementary bacterial indicator, the ratio of faecal coliform organisms to faecal streptococci (> 3: 1 for human wastes; < 0.7: 1 for other animal wastes) may be useful in locating the origin of faecal pollution in heavily contaminated sources of raw water, provided sufficient data are collected. In addition, these organisms can be used to assess the significance of doubtful results with the coliform test, particularly if coliform organisms are found in the absence of faecal coliform organisms. They can also be of value in checking water in the distribution system following repairs to mains. 1.2.2.2 Sulfite-reducing clostridia These are anaerobic spore-forming organisms, of which the most characteristic, C. perfringens (C. we/chii), is normally present in faeces though in much smaller numbers than E. coli. Clostridial spores can survive in water longer than organisms of the coliform group and they can resist disinfection if the concentration, contact time, or pH is
8
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MICROBIOLOGICAL ASPECTS
unsatisfactory. Their persistence in disinfected waters may thus indicate deficiencies in treatment (11). However, it would not be desirable to consider these organisms for the routine monitoring of distribution systems since they tend to survive and accumulate; they may thus be detected remote in both time and place from the original source of pollution and consequently give rise to false alarms.
1.2.3 Indicators of water quality
Apart from the colony counts, the use of other microorganisms, including Pseudomonas aeruginosa, has been advocated to assess the hygienic quality of drinking-water (12, 13). However, neither examination for these organisms, nor colony counts are essential for the routine monitoring of hygienic quality. They are of value in certain circumstances in giving an indication of the general cleanliness of the distribution system and in assessing the quality of bottled water. 1.2.3.1 Pseudomonas aeruginosa This organism often occurs in the faeces of man but in much lower numbers than coliform organisms. It is an opportunist pathogen of the very young and old and those already debilitated by disease, being frequently isolated from persons with urinary tract infections and skin burns (14). The organism occurs in raw water, usually in the presence of coliform organisms. However, in drinking-water it may occur in the absence of coliform organisms {15, 16) and the ability of some materials used in the construction of distribution and plumbing systems to support the growth of the organisms may account, in part, for this {17). Although the presence of the organism in potable water should not be ignored, it should not be used for the routine examination of water for the presence of faecal pollution. Examination for P. aeruginosa may be of value in certain circumstances, as for example, in the reconstitution of rehydration mixtures, baby foods, and pharmaceutical preparations, as well as in the surveillance of hospital water supplies and bottled water (18).
1.2.3.2 Colony counts Colony counts may be used to assess the general bacterial content of water. They do not represent the total number of microorganisms present in water, but simply those that are able to form visible colonies in nutrient media under specified culture conditions. They are of little value in detecting the presence of faecal pollution and should not be considered essential for assessing the safety of potable water supplies, although a sudden increase in the colony count from a groundwater
1.
THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
9
source may be an early sign of pollution of the aquifer (19). They are useful in assessing the efficiency of water-treatment processes, specifically coagulation, filtration, and disinfection, the objective being to maintain as low a density as possible in the treated water. They may also be used to assess the cleanliness and integrity of the distribution system and the suitability of water for use in the manufacture of food and drink where, to minimize the risk of spoilage, numbers should be low. The main value of colony counts lies in the comparison of results obtained from regular samples from the same supply so that any significant change from the normal range in a particular location can be detected.
1.3 Nuisance organisms These constitute a morphologically and physiologically diverse group of organisms, which includes planktonic and sessile algae, fungi, crustacea, and protozoa, as well as actinomycetes, and iron and sulfur bacteria. These organisms can cause objectionable tastes, colour, odour, and turbidity and may interfere with treatment processes by blocking strainers and filters. In addition, certain planktonic organisms, although not themselves harmful, may harbour pathogens and protect them against disinfection by chlorine. Most of these nuisance organisms can be controlled relatively easily by the usual water-treatment processes. Furthermore, problems of taste, odour, colour, and turbidity, which may be caused by them, are covered indirectly by the guidelines for the aesthetic characteristics of water given in Part V. The presence of certain organisms in water may be an indication of corrosion of cast iron or of biodeterioration of construction materials that support the growth of microorganisms (17). These include nonmetallic materials, such as plastics, rubbers, jointing compounds, and pipe-lining materials, which provide organic nutrients and thus encourage the growth of microorganisms including sometimes coliform organisms and P. aeruginosa. Although treatment will normally remove such organisms, they may establish themselves in sediment or slime and on materials within the distribution system, where their presence may support the growth of Asellus and other animal populations. Residual disinfection will help control such infestations, but occasional flushing and physical cleaning of pipe sections with polyurethane foam plugs may be needed. Nuisance organisms may also cause problems with groundwater sources by encrusting well-screens, thus causing loss of yield and impairing the aesthetic quality of the supply. Indeed, their presence may indicate organic pollution of the aquifer (20). Routine monitoring of such nuisance organisms cannot be recommended because of their diverse nature and unpredictable occurrence, although bacteriologists should be aware that they can impair water quality. It is not practicable to specify any quantitative limit for nuisance organisms.
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MICROBJOLOOJCAL ASPECTS
1.4 Disinfection The principal reasons for disinfecting drinking-water are to ensure the destruction of pathogens, to maintain a protective barrier against pathogens entering the distribution system, and to suppress bacterial regrowth in the pipe environment. Because of the importance of disinfection in safeguarding the hygienic quality of potable supplies, it is essential that the concentration of disinfectant should be measured frequently and preferably recorded continuously. For small supplies, especially those known to be at risk, simple disinfection and measuring devices should be available.
1.4.1 Disinfection efficiency The comparative efficiency of disinfectants may be expressed in terms of either the relative concentrations needed to attain the same rate of disinfection or the relative rates of disinfection produced by the same concentration of disinfecting agent. However, because of the different nature of the microorganisms and the difficulty of standardizing test conditions, such as pH, temperature, and the chemical characteristics of the water, only generalized statements can be made about the comparative efficiencies of different disinfectants. Within these limitations, disinfecting agents may be grouped according to their efficacy. Thus, the use of chlorine, chlorine dioxide, or ozone is preferable, although for chlorine the pH should be less than 8.0. Because chloramines are only slowly biocidal, their use as primary disinfecting agents for water-treatment purposes is not recommended, although they may be used for the maintenance of residuals in distribution systems where the contact time is longer. Similarly, in decreasing order, the relative resistance of different types of microorganisms and their probable survival may be listed as follows: protozoan cysts, enteroviruses, enterobacteria. Although there are distinct differences in the time required to inactivate enteroviruses as compared with enterobacteria, the minimum conditions of disinfectantresidual and contact-time required to ensure a microbiologically safe water supply can be achieved readily. It is, therefore, recommended that water from potentially polluted sources should always be disinfected. This would ensure inactivation of certain organisms, including some viruses, which may be relatively more resistant than faecal indicator bacteria.
1.4.2 Disinfectant residuals In addition to disinfection efficiency, another important consideration is the ability of these agents to remain as residual disinfectants during the storage and distribution of potable water. Except for ozone, all of the other practicable disinfectants (chlorine, chlorine dioxide, and
1.
THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
11
chloramines) can provide a persistent residual for continued microbial control once the finished water enters the distribution network_ However, chloramines are such slow biocides that any decision as to their use should be evaluated carefully with sufficient bacteriological data collected throughout the distribution system to demonstrate effectiveness in controlling microbial regrowth and providing protection against a moderate degree (1% sewage) of contamination from crossconnections (21). All supplies obtained from surface sources should be provided with disinfection as minimum treatment. Where disinfection is practised, a measurable residual should be maintained throughout the distribution system. Maintenance and monitoring of a chlorine residual offers two benefits. A chlorine residual will suppress the growth of organisms within the system and may afford some protection against contamination entering through cross-connection or leakage. The sudden disappearance of the residual provides an immediate indication of the entry of oxidizable matter into the system or of a malfunction of the treatment process. If chlorine is employed, it is desirable that a free chlorine residual of 0.2 to 0.5 mgjlitre be maintained and monitored daily throughout the entire system. When the residual in the supply is less than that routinely expected at a particular point, then remedial action, including increased chlorination, flushing and a sanitary survey, should be considered as the loss of residual may indicate the entry of pollution into the pipework. Booster or relay chlorination may be needed to ensure that this residual is maintained throughout the system. It is recognized that excessive levels of free chlorine may react with organic matter to produce tastes and odours in some waters. In these instances, the control agency or the medical officer of health should encourage necessary improvements in treatment or distribution and as a temporary measure establish a suitable concentration of chlorine residual to ensure microbiologically safe water. 1.4.3 Effects of turbiditya
Effective disinfection depends upon contact between the disinfecting agent and the microorganisms to be inactivated for an adequate period of time. Various bacteriological and virological studies have demonstrated a marked difference in the extent to which various types of particulate matter in water shield microorganisms from the effects of disinfectants. In general, inorganic particles, such as clay and waterflocculating agents, appear to have little, if any, protective effect. On the other hand, organic particulate matter, whether cell debris, sewage solids, or living or dead organisms such as nematodes or crustaceans, can provide marked protection to microorganisms associated with them. The degree of protection thus afforded is determined to a large extent by a For a fuller d1scuss1on of the effects of turbid1ty on drinkmg-water quality, see Part V, section 16.
12
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MICROBIOLOGICAL ASPECTS
the nature of the particulate matter rather than by the amount present as indicated by turbidity measurements (22). In all processes in which disinfection is practised, the turbidity must always be low, preferably below I nephelometric turbidity unit (NTU) and always less than 5 NTU, otherwise the particulate matter will interfere with the efficiency of disinfection, partly by exerting a disinfectant demand and partly by shielding microorganisms, even in the presence of a residual disinfectant otherwise sufficient to ensure a kill. Excessive water turbidity may also interfere with bacteriological examination, especially by membrane filtration. Low turbidity, particularly in treatment works relying on coagulation, can only be achieved by careful control and operation to ensure that the coagulant dose and pH are optimum, that floc-blankets are stable, and that filter runs are optimized by careful monitoring of headloss and turbidity. Filter backwashing is important to ensure that breakthrough of turbidity does not occur. Where coagulation, sedimentation and filtration are essential to ensure the removal of particulate matter, disinfection must invariably be userl to ensure microbiologically safe water. Any organic particulate matt~r present in potable water during distribution exerts a chlorine demand which reduces the available free chlorine residual, especially in dead-end sections of the system. Regular flushing of mains is desirable to avoid such accumulations. Organic turbidity can also serve as a source of nutrients, which may contribute to bacterial growth within the distribution network, especially in slow-flowing parts. Bacterial growth may enhance the accumulation of iron by bioflocculation. This results in the formation of a matrix of slime, calcium carbonate, and other debris attached to the pipe walls, which may result in deterioration in water quality. By-products of bacterial metabolism or decomposition within the slime may lead to taste and odour problems. Surges in water pressure may also dislodge accumulation of slime and sediments. 1.5. Collection, storage, and transport of water samples for bacteriological examination Care must be taken to ensure that samples are representative of the water to be examined and that no accidental contamination occurs during sampling. Sample collectors should, therefore, be trained and made aware of the responsible nature of their work. Samples should be clearly labelled with the site, date, time, nature of the water and other relevant information and sent to the laboratory for analysis without delay (23-26). 1.5.1 Sample bottles The size of the sample and bottle depends on the analyses to be undertaken, but 200 ml will usually be sufficient for routine examination for coliform organisms by either the multiple-tube or membrane-
1.
THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
13
filtration method. In special surveys, larger volumes may be necessary. Clean, sterile glass bottles should be used. They should be fitted with ground-glass stoppers or screw caps and the neck of the bottle should be protected from contamination by a suitable cover of paper or aluminium foil. Polypropylene bottles that will withstand autoclaving may also be used (23).
1.5.2 Neutralization of disinfectants If the water to be examined is likely to contain chlorine, chloramine, chlorine dioxide, or ozone, then 0.1 ml of an 18 gjlitre solution of sodium thiosulfate per 100 ml of bottle capacity should be added to neutralize any residual disinfectant. This should neutralize at least 5 mg of available chlorine per litre and will be suitable for routine sampling. In special situations where the residual may be greater, additional thiosulfate is required. This concentration of thiosulfate has no significant effect on the coliform organisms, including E. coli, either in chlorinated or in unchlorinated water samples during storage (24). When samples of disinfected water are taken, the concentration of residual disinfectant at the sampling point should be determined at the same time.
1.5.3 Sampling procedures When a number of samples are to be taken for various purposes from the same location, the sample for bacteriological examination should be collected first to avoid the danger of contamination of the sampling point. The sample bottle should be kept unopened until it is required for filling. The stopper should be removed with one hand and, during sampling, neither it nor the neck of the bottle should be allowed to touch anything. The bottle should be held in the other hand near the base. It should not be rinsed out and should be filled so as to leave a small air space when the stopper and cover are replaced. If it is not possible to collect the sample directly in a sample bottle, a sterile stainless steel jug should be used. This can be sterilized by igniting methylated spirits inside it. After filling, the water should be transferred to the sample bottle. 1.5.3.1 Sampling directly from a source When collecting samples directly from a river, stream, lake, reservoir, spring, or shallow well, the aim must be to obtain a sample that is representative of the water to be examined (23). It is therefore, undesirable to take samples too near the bank, or too far from the point of draw-off, or from the sediment near the bottom. Areas of stagnation should be avoided.
14
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MICROBIOLOGICAL ASPECTS
The sample should be taken by plunging the bottle, preferably neck downwards, into the water to a depth of 15-30 em so as to avoid floating debris. The bottle should then be turned neck upwards with the mouth facing the direction of the current. Where no current exists, the bottle should be pushed horizontally through the water. When sampling from a well or from the depths of a lake, reservoir, or cistern, a specially weighted sterile sample bottle or jug should be used. When sampling raw water in areas where schistosomiasis is endemic, waterproof gloves should be worn to avoid direct contact with the water. 1.5.3.2 Sampling from wells When taking a sample from a well fitted with a hand or mechanical pump, the pump should first be operated to flush out stagnant water from the pipework before the sample is taken. If possible, the mouth of the pump should be flamed, preferably by means of a blow lamp or gas (propane) torch, and a further quantity of water pumped to waste before sampling, which should be done by allowing the water from the pump to flow directly into the bottle. 1.5.3.3 Sampling in treatment works As frequent samples must be taken in treatment works, sampling taps should be provided throughout so that all stages of processing can be monitored. The pipework serving these taps should be short. If long lengths cannot be avoided, then the water should preferably be run continuously. Pipes carrying raw or partially treated water should be cleaned regularly to remove slime and deposits. If taps are not available, the samples may be taken by lowering the bottle into a tank or channel. 1.5.3.4 Sampling from storage tanks Whenever possible, sampling taps should be provided in all tanks, cisterns, and service reservoirs used for storing water. When these are not available, samples should be taken by lowering a weighted jug or bottle into the tank. Great care should be taken not to contaminate the water during sampling and to avoid taking the sample from the surface layer or the bottom where scum or sediment may be present. 1.5.3.5 Sampling from taps The majority of samples will be taken from taps-in treatment works, storage tanks, in the premises of consumers, or from public standpipes. Public health and water supply authorities should select sampling sites according to an agreed programme. When sampling from the distribution system, taps must be selected with great care. The taps chosen must be clean and should be supplied with water direct from the
1.
THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
15
public main. Additional samples may be needed from tanks supplying high-rise and multiple-occupancy buildings. Taps that leak between the spindle and gland should be avoided as water will run down the outside of the tap and thus contaminate the sample. External fittings, such as rubber or plastic nozzles and filters, should be removed, and the water run to waste for several minutes to ensure that the water in the pipework is flushed out before the sample is taken. Flaming the tap before the sample is taken should be regarded as an optional procedure. To avoid the problems caused by sampling from domestic taps, water supply authorities should consider the installation of protected sampling taps at strategic points in distribution systems. 1.5.3.6 Sampling from hydrants
Samples from public supplies should be taken from suitable taps supplying water direct from the mains. If this is not possible, they can be taken from street hydrants. Special care is necessary, however, including flushing and hydrant disinfection. 1.5.4 Transport and storage of samples The changes that may occur in the bacterial content of water on storage can be reduced to a minimum by ensuring that samples are not exposed to light and are kept cool, preferably between 4 °C and 10 °Cbut not frozen. They must not be contaminated by the refrigerant. Examination should begin as Don as possible after sampling, preferably within 24 hours. Any delays before examination must be taken into account when interpreting the results, and must be stated in the report. If delays are unavoidable, consideration should be given to the possibility of filtering samples on site and transferring them to transport media in airtight containers (23, 24). Membranes may be kept satisfactorily on such media for up to 3 days before transfer to conventional media for final examination in the normal way. 1.6 Recommended methods for tbe detection and enumeration of organisms indicative of pollution Two basic procedures are available for the detection and enumeration of indicator bacteria in water: the multiple-tube method, in which measured volumes of water are added to replicate tubes of a suitable liquid medium; and the membrane-filtration technique, in which measured volumes of water are passed through a membrane filter that retains bacteria on the surface. It must be appreciated, however, that the results obtained with one method or medium may not be strictly comparable with those given by another for the same group of organisms. The particular method chosen should, therefore, always be applied consistently so as to ensure comparability of results.
16
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MICROBIOLOGICAL ASPECTS
1.6.1 Laboratory facilities and safety
Satisfactory facilities and equipment for the examination of drinkingwater are important for ensuring reliable and reproducible results, especially for the detection of small numbers of faecal indicator organisms. Clean, comfortable working conditions are essential, not only to prevent cross-contamination of samples, but also for the health and safety of laboratory staff. Attention should be given to good laboratory practice, including the training of staff, the correct operation of incubators and water-baths, and careful preparation of media, as well as the use of quality-control procedures (26).
1.6.2 Detection of coliform organisms
Since coliform bacteria are present in large numbers in excrement and can be detected in concentrations as low as 1 per 100 ml, they are a sensitive indicator of the presence of faecal pollution. 1.6.2.1 Definition of the coliform organisms
The term coliform organisms (total coliforms) refers to Gramnegative, rod-shaped bacteria capable of growth in the presence of bilesalts or other surface-active agents with similar growth-inhibiting properties and able to ferment lactose at either 35 oc or 37 oc with the production of acid, gas, and aldehyde within 24-48 hours. They are also oxidase-negative and non-spore-forming. Those that have the same properties at a temperature of 44 oc or 44.5 oc are described as faecal (thermotolerant) coliform organisms. Faecal coliform organisms that ferment both lactose and other suitable substrates, such as mannitol, at 44 oc or 44.5 oc with the production of acid and gas and that also form indole from tryptophan, are regarded as presumptive E. coli. Confirmation of E. coli may be made by demonstration of a positive result in the methyl red test, by failure to produce acetyl methyl carbinol, and by failure to utilize citrate as the sole source of carbon. These are not taxonomic distinctions but practical working definitions used in water examination and they thus include members of several genera. Some organisms that would be identified taxonomically as "coliform" will, therefore, not be recognized as such in water examination. Examples include both anaerogenic and non-lactose fermenting strains of coliform organisms, as well as occasional strains of E. coli that are not thermotolerant. However, such strains are atypical and are greatly outnumbered in the environment by those that give the classical reactions, so that in practice the interpretation of results should not be affected. The choice of tests for the detection and confirmation of the coliform group of organisms should be regarded as part of a professional judgement based partly on the type of water, partly on the objective of the examination, and partly on the laboratory capability.
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THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
17
1.6.2.2 Techniques for the detection of coliform organisms Two basic methods are used for the detection and enumeration of coliform organisms in water: the multiple-tube method and the membrane-filtration method (1, 23-28). The two methods do not give strictly comparable results, one reason being that counts on membrane filters give no indication of gas production from lactose, but for practical purposes they do yield comparable information. For raw water, the detection of coliform bacteria and faecal coliform organisms may be an adequate guide to the microbial quality of the water. In the control of water-treatment processes, coliform organisms (total coliforms) should not be detectable in finished water. Should any coliform organisms be present in a water supply, it is then important that confirmation and differentiation be taken as far as possible in order to determine if the contamination is faecal in origin and to aid in tracing the source. (a) Multiple-tube method
The initial test is presumptive because the acid and gas reaction observed may sometimes be caused by some other organism or combination of organisms. The presumption that coliform organisms have given the observed reaction has to be confirmed by additional tests with further confirmatory and differential media. The proportion of false-positive reactions depends on both the bacterial flora of the water under examination and the medium used. An estimate of the number of presumptive coliform organisms present in a given volume of water can be obtained by inoculation of appropriate volumes into a number of tubes of medium. On incubation, it is assumed that each tube that receives one or more viable organisms in the inoculum will show growth and the positive reactions appropriate to the medium used. Provided that some negative results occur, the most probable number (MPN) of organisms in the original sample may be estimated from the number of tubes giving a positive reaction. Statistical tables of probability are used for this purpose and these, together with 95% confidence limits, are given in Annex 2 to volume 1. The MPN procedure is applicable to water of all types and especially to that in which the turbidity is high. The equipment required is relatively cheap and simple, and positive reactions are easy to read. The technique, however, provides only an estimate of the number of bacteria in any sample, and this is subject to considerable inherent error (24). Subcultures should be made to confirmatory media and also to solid media to ensure purity before further differentiation can be undertaken. Volume of water to be examined. The volumes of samples used in tests with liquid media depend on the suspected bacterial content of the water to be examined. With water of good quality, one volume of 50 ml and five of 10 ml should be suitable. If the water is of doubtful or unknown
18
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MICROBIOLOGICAL ASPECTS
quality, one 50-ml, five 10-ml and five 1-ml quantities should be used. The 50-ml and 10-ml volumes are added to the same quantity of doublestrength medium, and the 1-ml volumes to 5 ml of single-strength medium. Heavily polluted raw water should be diluted in order to obtain some negative reactions, and thus a definite result for the MPN. Choice of media. Four different media are available for use in the presumptive coliform test: minerals-modified glutamate medium (MMGM), Jauryl tryptose broth (LTB), MacConkey broth and lactose broth {1, 23-28). MMGM is a chemically defined medium with limited nutrients that can be utilized by coliform organisms. The selectivity of MacConkey broth and LTB depends respectively on the presence of bilesalts and the surface active agent, Iaury) sulfate; lactose broth is a nonselective medium. After inoculation, tubes of the selected medium are incubated at 35-37°C and inspected at 24 ±2 hours and 48 ±3 hours for the appropriate positive reaction. The gas and acid produced by the fermentation of lactose is detected as follows: in all media, an inverted inner (Durham) tube is used to trap gas, and in MMGM, MacConkey broth, and lactose broth a pH indicator in the medium is also used to demonstrate acidity. Tubes should be tapped to release dissolved gas before they are discarded as negative. The presence of coliform organisms should be confirmed in all tubes giving positive reactions within 24 and 48 hours. This arbitrary time limit excludes occasional members of the coliform group that form gas slowly, but such organisms are generally of limited sanitary significance. The presence of sporebearing organisms can cause false positive reactions, but they are usually eliminated during subsequent confirmatory tests. Confirmatory tests. The presence of coliform organisms in positive presumptive reactions should be confirmed, and followed by further differential tests for faecal coliform organisms and E. coli. Tubes exhibiting positive presumptive reactions at 24 hours and 48 hours are subcultured directly to tubes of selective media containing lactose for incubation for 48 ± 3 hours at the same temperatures used in the presumptive test. Since the original reaction may contain a mixture of organisms, a selective medium is necessary and brilliant-green (lactose) bile (BGB) broth may also be used {1, 23, 24). Production of gas confirms the presence of coliform bacteria in the original presumptive test. Tubes with negative reactions should be tapped gently to release dissolved gas. By reference to MPN tables, an estimate of the confirmed coliform content in 100 ml of the original sample may be obtained. To confirm the presence of faecal coliform organisms, subcultures should be made from tubes with positive presumptive reactions to tubes of media, such as BGB or Eca broth (23) which contain bile-salts for incubation at 44 ±0.25 oc for 24 ± 2 hours.b The production of gas a Escherichza col!. b If the procedures given in Standard methods for th~ examinatzon of water and wastewater ( 23) are followed, then 44 5 :tO 2'C should be substztuted throughout for 44 ±O 25'C.
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THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
19
confirms the presence of faecal coliform organisms in the original tubes and a confirmed MPN value can again be obtained. In addition, the numbers of presumptive E. coli may be obtained by inoculating tubes of tryptone water and testing for indole formation after 24 ±2 hours at 44 or 44.5 oc (23, 24). If positive, the presence of E. coli should be confirmed by differential biochemical tests. For convenience, a single tube of medium-in which both gas and indole formation can be demonstrated at 44 °C-may be used for the confirmation of presumptive E. coli (29). To avoid the problems associated with strains of E. coli deficient in galactoside permease, mannitol has been used as the fermentable substrate. If only one positive reaction (either gas or indole) occurs, the tests should be repeated in separate tubes of media.
(b) Membrane-filtration technique
The number of coliform organisms in water may also be determined by filtration of measured volumes of the sample through membrane filters (1, 24-28). These are normally composed of cellulose esters, typically with pores 0.45 Jlm in diameter, which retain coliform and many other bacteria present in the sample. The membranes are then incubated face upwards on a selective medium. Characteristic acid- or aldehyde-producing colonies develop on the membrane and these are counted as either presumptive coliform organisms or faecal coliform organisms, depending on the temperature of incubation. Since gas production is not detected on membranes, it is presumed that all colonies that produce acid or aldehyde also produce gas. However, the techniques used in subsequent confirmation will demonstrate gas formation. The results are expressed as the number of organisms in I 00 ml of the original sample. In practice the membrane-filtration technique gives results comparable with those of the multiple-tube method. However, if the sample is filtered through two membranes and one incubated at 35 or 37 cc and the other at 44 or 44.5°C, the confirmation procedure is somewhat simplified as a direct estimate of the number of faecal coliform organisms present is possible at the higher temperature. An advantage of the membrane-filtration method is the promptness with which results can be obtained, thus permitting rapid corrective action and return to normal operation. The technique may be used for the examination of most water, except water of high turbidity, when the membrane will block before a sufficient amount can be filtered. Membranes are also unsuitable for water containing few coliform organisms in the presence of many other organisms capable of growing on the media used, since the latter are liable to cover the membrane and interfere with the growth of the coliform organisms. If non-gas-producing, but lactosefermenting organisms such as Aeromonas predominate in the water, all presumptive coliform colonies on membranes should be confirmed because of the high proportion of false-positive results. The oxidase test
20
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MICROBIOLOGICAL ASPECTS
will help in the rapid elimination of such false-positive results. Aerobic spore-bearing organisms, which may cause false presumptive reactions in liquid media, will not do so on membranes. The membrane technique may be modified to encourage the recovery of attenuated organisms. Preincubation at lower temperatures, or the use of less selective media, may allow stressed organisms to recover and start growing, after which the test is completed in the normal way (23, 24). In the membrane-filtration technique, a direct count is made of discrete colonies, but this is still subject to statistical error. In addition, colony morphology can be examined and direct subcultures made, thus reducing the possibility of false-positive reactions from mixed cultures. Although only limited quantities of equipment are required, this is initially more expensive than the equipment used in the multiple-tube test. Membrane filters may be reused provided they remain undamaged after adequate washing and sterilization by boiling, and provided they are only used with the same medium. In addition, it should be appreciated that different membranes have different characteristics; it is thus important to ensure that they are suitable, not only for growth of the organisms sought, but also for the water concerned. The results given by membrane filtration are not necessarily the same as those obtained by the multiple-tube method, although they usually give comparable results in practice. It is essential, therefore, that an adequate series of parallel tests should be carried out by both methods in order to establish that the membrane-filtration technique is suitable for the water concerned. Filtration apparatus and technique. The apparatus consists of a sintered disc supported in silicone rubber gaskets fitted in a base to which a graduated funnel can be attached. The sintered or perforated disc supports the membrane filter. For use, the filter-holding assembly is mounted on a flask with a side-arm connected to a vacuum system. A series of filter-holders may be mounted in a manifold, thus permitting several samples to be filtered at the same time. After filtration of the water, the membrane is removed and placed face-upwards either on a suitable agar medium or on a pad soaked in liquid medium in a Petri dish for incubation at the appropriate temperature. Full details of the equipment can be found elsewhere (1, 23-28, 30). After incubation, the membranes should be examined in a good light. The appearance of colonies will depend on the medium used, but all characteristic colonies should be counted irrespective of size. If necessary, individual colonies may be subcultured into liquid media for confirmation, or on to a solid medium to ensure purity before further differential tests are performed. For the examination of water samples for coliform organisms (total coliforms) and faecal coliform organisms (presumptive E. coli), two separate membranes, appropriate media, and different incubation temperatures are required. Volume of water to be examined. Counts of the total numbers of coliform organisms and of faecal coliform organisms are made with
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THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
21
separate volumes of water, normally of 100 ml. Unless the samples are likely to contain more than 100 coliform organisms in 100 ml, the filtration of 100 ml is necessary for each test. For polluted samples, the volumes should be chosen so that the number of colonies on the membranes lies between approximately 10 and 100. If this volume is less than 10 ml, it should be mixed with a sterile diluent, such as quarterstrength Ringer's solution, lgjlitre peptone water, or a buffered dilution water, so that a minimum of 10 ml is filtered through the membrane. Choice of medium. Various media can be used in the examination for coliform organisms by the membrane-filtration method. Of these, lactose tergitol agar (31), lactose TTC tergitol agar (31), 0 and Iaury! sulfate lactose broth (23, 29) may be used for counts of coliform organisms at 35-37 °C, and faecal coliform organisms at 44 oc. Endo-type media should be used only for coliform counts at 35 or 37 °C, and MFCb broth at 44oC for faecal coliform counts. Although all these media rely on the fermentation of lactose for the detection of presumptive coliform organisms, the characteristic reaction varies with each medium. The characteristic metallic sheen of colonies on Endo media depends on the formation of aldehyde. Confirmatory tests. The extent to which individual colonies should be confirmed depends both on the water and on the reasons for the examination. The information required is similar to that for presumptive positive reactions in the multiple-tube method. To confirm the presence of coliform organisms, gas production from lactose must be demonstrated within 48 hours at 35 or 37 °C; and for faecal coliform organisms within 24 hours at 44 or 44.5 °C. In addition, the ability to produce indole from tryptophan at 44 oc confirms the presence of presumptive E. coli. The confirmatory procedures differ slightly from those for the multiple-tube test because colonies from membranes sometimes grow poorly in selective media, such as BGB broth, which should preferably not be used for direct confirmation (1, 23, 24). Also, because membranes are normally incubated at both 35 and 44 oc or 37 and 44.5 °C, it is possible to confirm colonies directly by demonstrating gas production at 44 oc without performing the test at a lower temperature (23, 24). Gas production at 35 or 37 oc may be demonstrated with lactose peptone water (24) or Iaury] tryptose broth (23, 24) and at 44 oc with lactose peptone water or EC broth (23, 24). Tryptone water should be used for indole production (23, 24). Ideally, all presumptive colonies on membranes should be examined, although this is not always practicable. However, as coliform organisms should not be present in treated waters, all colonies from such samples should be investigated further. With raw water before treatment, the samples should be diluted to yield a manageable number of colonies on membranes. In these circumstances, it is desirable to examine some a TIC = 2,3,5-triphenyltetrazohum chlonde b
Medmm for faecal coliform orgamsms.
22
I.
MICROBIOLOGICAL ASPECTS
suspicious colonies, preferably at least 10, to determine the nature and extent of the pollution. 1.6.2.3. Differentiation of coliform organisms Because of the importance of confirming the identity of any presumptive E. coli in drinking-water, further differential tests are required, including if necessary the use of commercially available identification kits. In temperate climates, such tests may be expected to confirm that organisms that produce acid, aldehyde, gas, or indole at 44 oc are in fact E. coli. In hot climates other coliform organisms, such as Enterobacter spp. (32-34) of lesser hygienic significance, may give the presumptive reaction of E. coli, but can be differentiated by these tests. Differential tests must be performed on pure cultures isolated from confirmatory media by subculture on to plates of a non-selective medium. Typical colonies are then used for the indole, methyl red, Voges-Proskauer, citrate, and oxidase tests, and if necessary other biochemical reactions (27, 28). The results, together with those from the confirmatory media, are used to identify E. coli as defined in section 1.6.2.1. The same tests may also be used for the differentiation of other coliform organisms.
1.6.3 Detection of faecal streptococci Both the multiple-tube and the membrane-filtration methods may be used for the presumptive detection of this group of organisms, but the application of these techniques is subject to the same limitations noted earlier for coliform organisms. The results of these tests may need confirmation (1, 23, 24, 26-28). 1.6.3.1 Definition of faecal streptococci Faecal streptococci belong to Lancefield's serological groups D and Q, which include S. faecalis and its varieties, S. faecium, S. durans, S. bovis, and strains with properties intermediate between them. They also include S. equinus and S. avium. They are capable of growth at 45 oc in the presence of 40% bile and in concentrations of sodium azide that are inhibitory to coliform organisms and most other Gram-negative bacteria. They are catalase negative. Some species resist heating at 60 oc for 30 minutes, and will grow at pH 9.6 and in media containing 65 g of sodium chloride per litre. 1.6.3.2 Multiple-tube method Appropriate volumes of water are added to tubes of single- or doublestrength glucose (dextrose)-azide broth (1, 23, 24) and incubated at 35 or 37 oc ± 0.5 o C for 48 ± 3 hours, the incubation being continued if necessary for 72 hours (24). Tubes that show acid production and
1.
THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
23
turbidity, often with a sediment, are regarded as containing presumptive faecal streptococci. The production of gas is not sought. The presence of faecal streptococci in tubes that give positive reactions should be confirmed by inoculation into ethyl violet azide broth for incubation at 35 ±0.5 oc for 48 ±2 hours. Heavy inoculation of confirmatory media is normally needed (23, 24). Tubes with positive confirmed reaction for faecal streptococci show a blue-purple sediment with turbidity. An alternative confirmatory procedure involves subculture from presumptive tubes on to plates of Pfizer selective enterococcus (PSE) agar. The growth of brown/black colonies with brown haloes in 24 hours at 35 ± 0.5 oc confirms the presence of faecal streptococci (23). The number of presumptive and confirmed faecal streptococci in 100 ml of original sample are calculated as described previously for coliform organisms using the probability tables in volume I, Annex 2. 1.6.3.3 Membrane-filtration technique Two media, KF agar and m-Enterococcus agar (Slanetz & Bartley), which both contain azide but different carbohydrates, are in general use for the enumeration of faecal streptococci by the membrane-filtration technique. With m-Enterococcus agar, selectivity is linked to the temperature of incubation and when incubated at 37 ± 0.5 oc for 4 hours, followed by 44 ± 3 hours at 44 ± 0.25 oc (23, 24), the method is especially selective for S. faecalis and S. faecium (35). This may be an operational advantage as it may supplement the coliform test in confirming the presence of faecal pollution, especially human (24). Overall recovery of faecal strep~ococci on this medium is sometimes poor (10, 35). For general use in the detection of all faecal streptococci from water, KF medium, which gives good recovery and selectivity (36, 37), is most often used. Selectivity of KF agar may be enhanced if the medium is sterilized by boiling rather than autoclaving. Membranes are incubated on KF agar at 35 or 37 ± 0.5 oc for 48 ± 3 hours (23). As this medium is both stable and selective for faecal streptococci, membranes can be transported on it for up to 3 days before completion of the test in the laboratory. Media used in the membrane-filtration method for faecal streptococci are generally very selective and the red colonies formed on both KF and m-Enterococcus agar are usually assumed to be faecal streptococci without further confirmation. If considered necessary, confirmatory tests should demonstrate growth in the presence of 40% bile at 44 oc and a negative catalase reaction (23). The density of faecal streptococci should be expressed in numbers of colonies per I 00 ml of the original sample. 1.6.3.4 Differentiation of faecal streptococci When the test for faecal streptococci is used to assess the source of pollution, it may be necessary to identify the species present. Following
24
I.
MICROBIOLOGICAL ASPECTS
initial examination, all isolates that are catalase-negative may be further identified by tests for growth at 45 'C, at pH 9.6, and in 65 gjlitre sodium chloride, for the hydrolysis of aesculin and starch, as well as for lactose fermentation and reduction of 1 g/litre methylene blue milk.
1.6.4 Detection of sulfite-reducing clostridia Sulfite-reducing clostridia, especially C. perfringens (C. welchii), may also be used as indicators of faecal pollution (38). The organisms in this group are characterized by their ability to form spores and to reduce sulfite to sulfide. This feature is utilized in media for the presumptive detection of clostridia by the formation of a black precipitate of iron sulfide. For the detection of spores, all vegetative cells are inactivated by heating the samples of water at 75-80 oc for 10 minutes (24). 1.6.4.1 Definition of sulfite-reducing clostridia This group of organisms consists of anaerobic sulfite-reducing, sporeforming, Gram-positive, catalase-negative, rod-shaped bacteria. In addition, C. perfringens ferments lactose, sucrose, and inositol with the production of gas, produces a typical "stormy-clot" reaction in litmus milk, hydrolyses gelatin, and produces lecithinase and acid phosphatase. This organism is non-motile. Both the multiple-tube and membrane-filtration methods can be used for the detection of sulfite-reducing clostridia. 1.6.4.2 Multiple-tube method The procedure is similar to that described for coliform organisms except that gas production is not sought. In order to maintain anaerobic conditions throughout the period of incubation, screw-capped bottles must be used and filled with broth containing glucose and sulfite, such as differential reinforced clostridia medium (DRCM) (24). Appropriate volumes of sample are inoculated into bottles of single- and doublestrength medium and incubated at 35 or 37 ± 0.5 oc for 48 ± 3 hours. Tubes that show blackening are presumed to contain sulfite-reducing bacteria and, if the sample was heated before testing, the tubes are presumed to contain sulfite-reducing spore-forming clostridia. The presence of C. perfringens can be confirmed by subculture of growth from bottles with presumptive positive reactions to tubes of litmus milk (22) for incubation at 35 or 37 ± 0.5 oc for 48 ± 3 hours. A typical "stormy clot" reaction, together with acidity. confirms the presence of C. perfringens. If further confirmation is required, tests for motility and nitrate-reduction may be done (39). 1.6.4.3 Membrane-filtration technique No single medium has found general acceptance for the enumeration of sulfite-reducing clostridia in water, but iron sulfite agar (ISA) (38),
1.
THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
25
sulfite-polymyxin-sulfadiazine agar (SPS) (39), and medium for Clostridium perfringens (mCP) (40) have been recommended. The incubation time for each medium is 24 hours. Incubation temperatures vary depending on the medium, being 37 oc for SPS, 45 oc for mCP, and 48 oc for ISA. Anaerobic conditions are necessary for growth, and these can be achieved by the use of agar overlays with ISA and SPS, but not with mCP where the need to expose the plates subsequently to ammonia vapour to demonstrate phosphatase reduction precludes this technique. Incubation of mCP must, therefore, be in an anaerobic atmosphere. If samples are heated at 75-80 oc for 10 minutes before filtering, these techniques will be highly selective for sulfite-reducing clostridia. Confirmation of the presence of C. perfringens is necessary only if iron sulfite agar is used, as SPC and mCP, which contain appropriate antibiotics, are very selective for this organism. Nevertheless, media that contain sulfite have stood the test of time and are recommended as standard media; in addition, they have the advantage of making the membrane-filtration technique similar in use and mode of operation to the MPN method. 1.6.5 Pseudomonas aeruginosa P. aeruginosa can be defined and readily distinguished from other fluorescent Pseudomonas species by the following differential features: production of pigment, growth at 42 °C, hydrolysis of casein, and the use of unusual sources of organic carbon.
1.6.5.1 Definition of P. aeruginosa P. aeruginosa is a Gram-negative, non-sporing, rod-shaped bacterium that grows at 42 °C, may produce pyocyanin and fluorescent pigments, is oxidase- and catalase-positive, reduces nitrate beyond nitrite, liquefies gelatin, hydrolyses casein but not starch, oxidizes glucose, and reduces acetamide to ammonia.
1.6.5.2 Multiple-tube method Media containing asparagine enhance pigment formation, and this forms the basis of the modification of Drake's broth in current use for the multiple-tube method (23). The addition of ethanol (20 ml/1) to the medium prevents the growth of other Gram-negative bacteria during incubation at 35 or 37 °C. Appropriate volumes of sample are added to tubes of single- and double-strength broth for incubation at 35 or 37 oc for 4 days. All tubes showing growth should be examined daily for pigment production and fluorescence in ultraviolet light. The presence of P. aeruginosa in these presumptive reactions should be confirmed by subculture to cetrimide milk agar or acetamide broth (23). Casein hydrolysis, together with the
26
I.
MICROBIOLOGICAL ASPECTS
production of a blue-green fluorescent pigment after incubation of milk agar at 41.5 oc for 24 ±2 hours, and the production of alkaline conditions, as indicated by a purple coloration, within 36 hours at 35 oc or 37 oc with acetamide broth, confirms the presence of P. aeruginosa. 1.6.5.3 Membrane-filtration technique King's A medium modified by the addition of ethanol (23) is usually used. After filtration, membranes are placed on modified King's A medium and incubated for 48 ±3 hours at 35 or 37°C. Fluorescent, green-pigmented colonies are considered to be P. aeruginosa. Colonies should be confirmed on milk agar at 41.5°C (23) for casein hydrolysis and also the production of green pigment and fluorescence within 24 hours. 1.6.6 Colony counts
The usual procedures in current use for estimating the bacterial content of water are the pour-plate and surface-spread methods (1, 23-28, 31). A membrane-filter technique has also been developed recently (41). For colony counts, the nutrient media used will support the growth of only a proportion of microorganisms present in any water sample, and this will vary with the medium. Anaerobic organisms will not usually grow, even in pour plates. In addition, because microorganisms also occur in clumps and chains, counting those that actually form colonies will underestimate considerably the actual number of viable microorganisms present in the sample. As different microorganisms have different optimum temperatures for growth, it is normal to incubate two plates prepared from the same sample at 35-37 oc for 1-2 days and another set at 20-22°C for 3 days. The period of incubation influences the colony count, and it is, therefore, important to adhere strictly to the same practice so that the results are always comparable. For bottled water, it is usual to incubate plates at 35 or 37 oc for 3 days (23). The count is expressed as the number of colony-forming units in I ml of the original sample, stating the medium, the temperature and duration of incubation, and the method used. Briefly, the technique may be outlined as follows: a series of tenfold dilutions of the sample is made, the number depending on the nature and history of the water. From each dilution, I ml is added to each of two sterile Petri dishes; molten, nutrient agar at 44-46°C is added (15ml) to each plate and the sample and medium mixed by rotation. After the agar has set, the plates are inverted and incubated at the desired temperature. The colonies on each pair of plates are then counted, and the number of microorganisms in the original sample obtained by multiplying the arithmetic mean of the counts by the reciprocal of the dilution. Where there are no plates with counts between 30 and 300 colonies, the colonies should still be counted but the result should be recorded as an estimate.
1.
THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
27
1.6. 7 Examination for pathogenic organisms Although direct search for specific pathogenic bacteria has no place in the routine bacteriological examination of water, there are occasions when examination for intestinal pathogens may be necessary as, for example, during an epidemic or in the evaluation of a new source. The chances of success will then be greater if large samples of water are examined, and if media selective for certain intestinal pathogens are used. Examination will include some, if not all, of the following stages: concentration of the organisms in the sample, inoculation into enrichment broth, subculture on to selective agar media, and biochemical and serological examination of suspect colonies. Rather than rely on a single method, it is better to use as many methods as possible so that no opportunity to detect a pathogen is missed (23, 24, 26-28). This is especially so for the detection of Salmonella since no single method is suitable for all serotypes. 1.6. 7 .I Concentration of samples The technique used will depend largely on the amount of particulate matter in the water. In waters of low turbidity, the sample may be passed through membrane filters. As turbidity increases in raw waters, filtration through diatomaceous earth (23, 24) or cartridge filters (42) may be u~ed to enhance filtration and permit processing of larger sample volumes. Alternatively, use may be made of the gauze-pad technique (23), especially when the numbers of pathogens are small or their presence is not continuous. 1.6.7.2 Salmonella Sample concentrates may require pre-enrichment in buffered peptone water, followed by enrichment in broth containing either tetrathionate, selenite, magnesium chloride, or malachite green. These may be subcultured on to media such as brilliant green, bismuth sulfite, xyloselysine desoxycholate (XLD) agar, desoxycholate citrate, or MacConkey agar, and suspect colonies examined biochemically and serologically. Biochemical screening tests should include triple sugar iron agar, indole production, decarboxylase, and /1-galactosidase activity. Serological testing should include agglutination with polyvalent anti-0, anti-H, and anti-Vi sera. Prior elimination of auto-agglutinable strains is essential. When S. typhi is sought, selenite F medium is preferred. The multipletube procedure is used for estimation of the number of Salmonella present. 1.6.7.3 Shigella Since coliform bacteria and most strains of Proteus vulgaris are antagonistic to Shigella, it is advisable to choose selective enrichment media that minimize accumulations of volatile compounds and by-
28
I.
MICROBIOLOGICAL ASPECTS
products derived from these antagonists (23). Nutrient broth adjusted to pH 8.0 (less favourable pH for coliform growth) may be used. Successful enrichment of Shigella may also be achieved with an autocytotoxic medium based on trypticase soy broth containing I mmol/litre 4-chloro2-cyclopentylphenyl P-o-galactopyranoside, 2.5 gjlitre lactose, and citrate buffer at pH 6.2 (43). Incubate 6-18 hours at 35 oc. Streak cultures at 6 and 18 hours to XLD agar. Submit suspect colonies to biochemical screening tests and confirm suspect colonies with Shigella antisera (polyvalent and type-specific sera). 1.6. 7.4 Cholera and non-cholera vibrios Alkaline peptone water or taurocholate tellurite peptone water are used for primary enrichment, with subculture to thiosulfate citrate bile salt sucrose agar or taurocholate tellurite gelatin agar as selective media (44). Suspect cultures are inoculated into Kligler iron agar. After 18 hours' incubation, V. cholerae produces a distinctive yellow colour, without any gas production. These cultures are then further screened for urease and oxidase activity; those strains that are urea-negative and oxidase-positive should be submitted to a reference laboratory for further biochemical tests and serological grouping. 1.6. 7.5 Enteropathogenic E. coli The techniques for the detection of faecal coliform organisms in water are used. The colonies are confirmed as E. coli and if the epidemiological evidence warrants it, subcultures may be submitted to a reference laboratory for serological grouping and, if necessary, tests for en terotoxigenicity. 1.6.7.6 Yersinia enterocolitica M-Endo agar has been found to be the most suitable all-purpose medium owing to the distinct morphological characteristics of colonies at both 25 and 35 oc incubations. Colonies are dark red and well defined after 72 hours' incubation. Growth on MacConkey's agar is equally good provided 25 oc incubation is used (45). All suspect isolates should be screened biochemically at 25 oc and 35 oc on rhamnose, raffinose, and melibiose. If epidemiological evidence warrants it, subcultures should be submitted to a reference laboratory for serological grouping and antibiotic susceptibility tests. 1.6.7.7 Campylobacter fetus A membrane-filter technique used successfully for isolation of this pathogen involves blood agar containing vancomycin, polymyxin, and trimethoprim (46). Cultures are incubated at 42-43 oc under reduced oxygen tension in an anaerobic jar for 3 days and inspected daily for nonhaemolytic grey mucoid colonies (1-2-mm diameter). Suspect
1.
THE BACTERIOLOGICAL QUALITY OF DRINKING-WATER
29
colonies are Gram-stained for typical curved, S-shaped forms and tested for a positive oxidase and catalase reaction, motility, and inability to grow aerobically at 36 oc. Subcultures should be submitted to a reference laboratory for further biochemical tests. Serotyping is not practical on isolates from sporadic outbreaks because of the heterogeneity of the organism and the use of a common antigen or pool of differential serotypes is not yet practical.
REFERENCES I. International standards for drinking-water, 3rd ed. Geneva, World Health Organization, 1971. 2. KABLER, P. W. & CLARK, H. F. Coliform group and fecal coliform organisms as indicators of pollution in drinking water. Journal of the American Water Works Association, 52: 1577 (1960). 3. AKIN, E. W. ET AL. A virus-in-water study of finished water from six communities. Cmcinnati, US Environmental Protection Agency, 1975. (Environmental protection technology series, EPA-600/1-75-003). 4. GELDREICH, E. E. ET AL. The occurrence of coliforms, fecal coliforms and streptococci on vegetation and insects. Applied microbiology, 12: 63 (1964). 5. PAPAVASSILIOU, J. ET AL. Coli-aerogenes bacteria on plants. Journal of applied bacteriology, 30: 219 (1967). 6. GELDREICH, E. E. ET AL. The faecal coli-aerogenes flora of soils from various geographical areas. Journal of applied bacteriology, 25: 87 (1962). 7. DEANER, D. G. & KERR!, K. D. Regrowth of fecal coliforms. Journal of the American Water Works Association, 61: 465 (1969). 8. GELDREICH, E. E. & KENNER, B. A. Concepts of fecal streptococci in stream pollution. Journal of the Water Pollution Control Federation, 41: R336 (1969). 9. KENNER, B. A. Fecal streptococcal indicators. In: Berg, G., ed., Indicators of viruses in water and food, Ann Arbor, Ann Arbor Science, 1978. 10. GELDREICH, E. E. Fecal coliform and fecal streptococcus density relationships in waste discharges and receiving waters. CRC critical reviews in environmental control, 6: 349-369 (1976). II. KooL, H. J. Treatment processes applied in public water supply for the removal of micro-organisms. In: Proceedings of a Symposium on Biological Indicators of Water Quality, Newcastle, 1-15 October 1978, vol. 2. Umversity of Newcastle, 1978, pp. 17-1-17-31. 12. FEACHEM, R. ET AL. Sanitation and d1sease: Health aspec1s of excreta and wastewater management. Baltimore, Johns Hopkins University Press, 1981 (World Bank Studies in Water Supply and Sanitation, No. 3). 13. GELDREICH, E. E. Current status of microbiological water quality criteria. American Society for M1crobw/ogy news, 47: 23-27 (1981). 14. HoADLEY, A. W. The significance of fluorescent pseudomonads in water. In: Hoadley, A. W. & Dutka, B. J., ed., Bacterial indicators of potential health hazards associated wah water. Philadelphia, American Society for Testing and Materials, 1977; pp. 80114. 15. REITLER, R. & SELIGMAN, R. Pseudomonas aeruginosa m drinking water. Journal of applied bacteriology, 20: 145-150 (1957). 16. NEMEDI, L. & LANYI, B. Incidence and hygienic importance of Pseudomonas aeruginosa m water. Acta m1crobiologica Academiae Scientiarium Hungaricae, 18: 319326 (1971). 17. BuRMAN, N. P. & CoLBOURNE, J. S. Effects of non-metallic materials on water quality. Journal of the Institute of Water Engineers and Scientists, 33: 11-18 (1979).
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MICROBIOLOGICAL ASPECTS
18. MossEL, D. A. A. ET AL. Microbiological quality assurance for weaning formulae. In: The mlcrobio/oglca/ safety of food, London, Academic Press, 1973, pp. 77-78. 19. MuLLER, G. Bactenal md1cators and standards for water quality m the Federal Republic of Germany In: Hoadley, A W. & Dutka, B. J., ed. Bactena/ md1cators. of potentw/ health hazards assocwted wl/h water, Philadelphia, Amencan Society for Testing and Matenals, 1977, pp 159-167 20. TAYLOR, E. W The pollutiOn of surface and underground waters Bnllsh Water Works AssoC/atwn JOurnal, 42: 582-603 ( 1960) 21. SNEAD, M. C ET AL. Biological evaluation of benefits of mamtammg a chlonne residual m water supply systems. Warer research, 14 403-408 (1980). 22. HoFF, J. C. & GELDREICH, E. E Effects of turbidity and other factors on the inactivatiOn of ~1ruses by chlonne. ln. Proceedmgs of the 1978 Annual Amencan Water Works AssoCiation Conference and Exposllwn. AtlantiC C11y, NJ, Denver. CO, A WWA, 1978 (Paper No. 35-1 C). 23. AMERICAN PUBLIC HEALTH ASSOCIATION. Standard methods for the exammation of water and wastewater, 15th ed., Washington, DC. APHA. 1980, 1134 pp. 24. DEPARTMENT OF HEALTH AND SOCIAL SECURITY. The bactenologu·al nammatwn of water supplies. London. HM Stationery Office. 1969 (Reports on Public Health and Medical Subjects. No. 71 ). 25. UNION OF SOVIET SociALIST REPUBLICS. [All Unwn State Standard Dnnkmg-warer methods of sanitary bactenologu·al analys1s]. Moscow, GOST 18963-73, 1973 26. GELDREICH, E. E. Handbook for emluatmg water bacteriological laboratones. Cincmnatl, US Envuonmental Protection Agency, 1975 (EPA-670/9-75-006) 27 [Methods for the unification of the sanitary microbwlog1cal exammatwn of water ] BadElster, Council for Mutual Economic Assistance, 1979. 28. COUNCIL FOR MUTUAL ECONOMIC ASSISTANCE. [Standard methods for water quality exammation. Part I Methods of chem1cal exammatwn of water], 3rd ed. Moscow, CMEA, 1977. (A summary translatiOn of the 1st edition has been published under the title Standard methods for the water qua lay exammation for the member countries of the Council for Mutual EconomiC Ass1stance. Prague, The Ministry of Forestry and Water Management in cooperatwn with the Hydraulic Research Institute, 1968.) 29. PUBLIC HEALTH LABORATORY SERVICE AND STANDING COMMITTEE 0~ ANALYSlS Smgle-tube confirmatory tests for Eschenchw coil. Journal of hyg1ene ( Camhndge), 85· 51-57 (1980). 30 C AIRNCROSS, S. & FEACHEM, R Small water supplies. London, The Ross Institute, 1978. 31. VIAL, 1 Bactenolog1cal analys1s of dnnkmg water. Luxembourg, CommiSSIOn of the European Communities, 1977. 32. RAGHAVACHARI, T. N. S. & IYER, P. V. S. The occurrence of aerogenes group of coliform orgamsms in faeces and its Significance m water analysis Jndwn Journal of med1cal research, 28: 55-60 (1940) 33. BoiZOT, G. E. An examinatiOn of the modified EIJkman method applied to pure coliform cultures obtamed from waters in Singapore. Journal of hyg1ene (Cambndge), 41: 566-569 ( 1941 ). 34. EviSON, L. M. & lAMES, A. A comparison of the distribution of mtestinal bactena in Bntish and East Afncan water sources. Journal of applied bactenology, 36: 109-118 (1973). 35. STANFIELD, G. ET AL. Isolatwn of faecal streptococCI from sewage. Stevenage, Water Research Centre, 1978. 36. CLAUSEN, E. M. ET AL. Fecal streptococci: indicators of pollution. In: Hoadley, A. W. & Dutka. B. J., ed., Bacterial indicators of potentwl health ha::.ards assocwted wah water, Philadelphia, American Society for Testing and Materials, 1977, pp. 247-264. 37 Summary report of a Working Group on Bacteriological Exammatwn of Water Copenhagen, WHO Regional Office for Europe, 1975. 38. BoNDE, G. J. Bactenal mdication on water pollution. In: Droop, M. R. & Jannasch, H. W., ed., Advances in aquatic m1crobwlogy, vol. I, London, Academic Press, 1977, 381 pp.
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39. ANGELOTII, R. ET AL. Quantitation of Clostridium perfringens in foods. Applied microbiology, 10: 193-199 (1962). 40. CABELL!, V. J. Clostridium perfringens as a water quality indicator. In: Hoadley, A. W. & Dutka, B. J., ed., Bacterial indicators of potential health hazards associated with water, Philadelphia, American Society for Testing and Materials, 1977, pp. 65-79. 41. TAYLOR, R. H. & GELD REICH, E. E. A new membrane filter procedure for bacterial counts in potable water and swimming pool samples. Journal of the American Water Works Association, 71: 402-405 (1979). 42. LEVIN, M. A. ET AL. Quantitative large-volume sampling technique. Applied microbiology, 28: 515-521 (I 974). 43. PARK, C. E. ET AL. Improved procedure of selective enrichment of Shigella in the presence of Escherichia coli by use of 4-chloro-2-cyclopentylphenyl beta- Dgalactopyranoside, Canadian journal of microbiology, 23: 563-566 (1977). 44. Guidelines for the laboratory diagnosis of cholera. Geneva, World Health Organization, 1974. 45. HIGHSMITH, A. K. ET AL. Isolation of Yersinia enterocolitica from well water and growth in distilled water. Applied and environmental microbiology, 34: 745-750 (1977). 46. SKIRROW, M. B. Campylobacter enteritis: A "new" disease. British medical journal, 2: 9-11 (1977).
2. THE VIROLOGICAL QUALITY OF DRINKING-WATER 2.1 General description Viruses of major concern in relation to waterborne transmiSSion of infectious disease are essentially those that multiply in the intestine and are excreted in large numbers in the faeces of infected individuals {1). Concentrations as high as 10 8 viral units per g of faeces have been reported. Even though replication does not occur outside living hosts, enteric viruses have considerable ability to survive in the aquatic environment and may remain viable for days or months (2). Viruses enter the water environment primarily by way of sewage discharges. With the methods at present available, wide fluctuations in the number of viruses in sewage, up to a maximum of 10 6 units/litre, have been found (3). On any given day, many of the 100 or so known enteric viruses can be isolated from municipal sewage, the specific types being those prevalent in the community at that time. Procedures for the isolation of every virus type that may be present in sewage are not yet available. Sewage treatment may reduce the concentration of viruses 10- to 1000-fold, the actual extent depending mainly on the nature and degree of treatment given. Even tertiary treatment does not yield an effluent consistently free from viruses (4). As sewage mixes with receiving water, viruses are carried downstream, remaining viable for varying periods of time depending upon temperature and a number of other less well-defined factors. Consequently, viruses are likely to be present in sewage-polluted water. At the intake to water-treatment plants, counts of up to 49 viral units/litre have been recorded (5).
2.2 Routes of exposure It is generally believed that the primary route of exposure to enteric viruses is by direct contact with infected persons or by contact with faecally contaminated objects. However, because of the ability of viruses to survive and because of the low infective dose, exposure and consequent infections may occur by less obvious means, including ingestion of contaminated drinking-water. Explosive outbreaks of viral hepatitis and gastroenteritis resulting from sewage contamination of water supplies have been well documented epidemiologically (6). In contrast, the transmission of low levels of virus through drinking-water of potable quality, although suspected 32
2.
THE VIROLOGICAL QUALITY OF DRINKING-WATER
33
of contributing to the maintenance of endemic enteric viral disease within communities (7), has not yet been demonstrated. In some developing areas, water sources may be heavily polluted and the water-treatment processes may be less sophisticated and reliable. Because of these factors, as well as the large number of persons at risk, drinking-water must be regarded as having a very significant potential as a vehicle for the environmental transmission of enteric viruses. As with other microbial infections, enteric viruses may also be transmitted by contaminated food and aerosols in addition to the usual mode of direct contact. 2.3 Health effects
Enteric viruses are capable of producing a wide variety of syndromes, including rashes, fever, gastroenteritis, myocarditis, meningitis, respiratory disease, and hepatitis. In general, asymptomatic infections are common and the more serious manifestations rare. However, when drinking-water is contaminated with sewage, two diseases may occur in epidemic proportions- gastroenteritis and infectious hepatitis. Apart from these infections, there is little, if any, epidemiological evidence to show that adequately treated drinking-water is concerned in the transmission of virus infections. Gastroenteritis of viral origin may be associated with a variety of agents. Many of these have been identified only recently (8), occurring as small particles with a diameter of 27-35 nm in stools of infected individuals with diarrhoea. Most of them have not been characterized chemically or cultured in the laboratory. Viral gastroenteritis, usually of 24-72 hours' duration with nausea, vomiting and diarrhoea, occurs in suspectible individuals of all ages. It is most serious in the very young or very old where dehydration and electrolyte imbalance can occur rapidly and threaten life if not corrected without delay. Hepatitis, if mild, may require only rest and restricted activities for a week or two, but when severe it may cause death from liver failure, or may result in chronic disease of the liver. Severe hepatitis is tolerated less well with increasing age and the fatality rate increases sharply beyond middle age. The mortality rate is higher among those with pre-existing malignancy and cirrhosis (9). 2.4 Rationale for recommendation
Theoretically, one viral particle is capable of initiating the infectious process; indeed studies with human volunteers have shown that, under somewhat artificial conditions, one unit of vaccine poliovirus detectable by tissue culture could initiate infection (10). However, the ingestion of viral particles-as in drinking-water-will not necessarily ensure that the essential initial contact with susceptible cells will take place. In practice, for most people, the minimum infective dose is likely to be greater than
34
I.
MICROBIOLOGICAL ASPECTS
one viral unit and it can best be expressed by the statistical probability of infecting a certain proportion of an exposed population with a given viral dose. In respect of good public health practice, however, it is difficult to disagree with the conclusion of the Safe Drinking Water Committee of the National Academy of Sciences, USA, that "The presence of infective virus in drinking-water is a potential hazard to the public health, and there is no valid basis on which a no-effect concentration of viral contamination in finished water might be established." (1 1). None the less, realistically and economically, it is questionable how far this conclusion can be applied and much further work is clearly required. In the absence of information relating the degree of viral contamination of water to disease, there is no scientific basis on which to set the level of a practical virological standard (12). Therefore, the most reasonable approach for controlling the transmission of viruses through drinking-water is to recommend consistently meeting the treatment criteria that have been found through years of experience to be effective in preventing obvious cases of waterborne viral disease. These treatment practices should accord with the criteria that have been shown experimentally to be effective in removing or inactivating seeded viruses. At the present time, the frequent examination of potable water for the occurrence of faecal indicator bacteria remains the only practical and economical approach for routinely monitoring the microbiological safety of drinking-water. 2.5 Methods of examination for viruses Methods for concentrating viruses from water samples have evolved rapidly during the last few years but their reliability, limits of detection, and precision are not yet well established. For raw water sources, a number of methods, as described in the manual Examination of water for pollution control (1 3), may be used. For finished water, the virus concentration method as described in Standard methods for the examination of water and wastewater (14) is recommended for large volumes. The assessment of water samples for viruses also has the limitation that a minimum of two weeks is required before a result can be given, by which time the water tested would have been distributed and consumed. It must be emphasized that examination of drinkingwater for viruses should not supplant bacteriological monitoring or other quality control measures, such as sanitary surveys and physicochemical measurement of the finished product; it should be regarded as augmenting them. It is obviously desirable that examination for coliform organisms should be done at the same time as viral examination, if necessary on larger volumes of water; if coliform organisms are detected, immediate remedial measures should be instituted without awaiting the virological results. In general, there are many more faecal bacteria in wastewater than there are viruses. This has led to the hope that suitable bacteria
2.
THE VIROLOGICAL QUALITY OF DRINKING-WATER
35
indicative of faecal pollution would also serve as indicators for the presence of viruses in all waters, thereby making direct examination for viral pollution unnecessary. However, bacteria used as conventional indicators to evaluate the safety of potable water supplies have been shown to be less resistant than viruses to environmental factors, and to water and wastewater treatment processes. As a result, enteric viruses could be present in water with little or no signs of bacterial pollution. Schemes for the recycling of wastewater for domestic use are being considered in some cities, while in many others, water for potable supplies is obtained from contaminated surface sources with a significant proportion of wastewater. In both situations, the risk of viruses penetrating the water-treatment processes-including pretreatment storage and disinfection-must be evaluated carefully. The possible use of bacteriophages of enteric bacteria as indicators for the potential presence of enteroviruses should be considered further. The speed and economy of bacteriophage tests compared with those for the detection of enteroviruses make such a proposition attractive (15). 2.6 Interpretation and evaluation of positive findings
In virological work with samples free from viruses or containing only a few viruses, the risk of accidental contamination during sampling and examination in the laboratory is a very real one. The utmost care should therefore be taken and examination of samples of drinking-water preferably separated from work with other viral material. Laboratory quality-control systems should be developed and participating laboratories encouraged to standardize their procedures. As a matter of course, it is suggested that two separate samples for virological examination should always be collected from the same site. One of these should be examined as soon as possible and the second one placed in cold storage. In the event of a positive finding in the first sample, the likelihood of accidental contamination should be assessed and the second sample should be assayed for virus only if considered necessary in the light of this assessment. Confirmed viral isolates should be type-identified as far as possible. The possible significance to health of positive findings should be evaluated in association with the public health authority.
REFERENCES I. WHO Technical Report Series, No. 639, 1979 (Human viruses in water, wastewater and soil: report of a WHO Scientific Group). 2. AKIN, E. W. ET AL. Enteric v1ruses in ground and surface waters: a review of their occurrence and survival. In: Proceedings of the 13th Water Quality Conference, Umversity of Illmois, 1971. 3. BuRAS, N. Concentration of enteric viruses in wastewater and effluent: a two-year survey. Water research, 10: 295-298 (1976).
36
I.
MICROBIOLOGICAL ASPECTS
4. MIELE, R. P. Pomona VIrus study- final report. Los Angeles, Samtation Districts of Los Angeles County, 1977. 5. SLADE, J. S. Enterov1ruses in partially punfied water. Journal of the lnsll/ute of Water Engineers and Scientists, 31: 219-225 (1977). 6. CRAUN, G. F. & McCABE, L. J. Review of the causes of waterborne disease outbreaks. Journal of the American Water Works Association, 65: 74-84 (1973). 7. BERG, G. Virus transmtssion by the water veh1cle. I. Viruses. Health laboratory science, 3: 86-89 (1966). 8. MADELEY, C. R. Viruses in the stools. Journal of clmical pathology, 32: 1-10 (1979). 9. NATIONAL RESEARCH COUNCIL. Proceedings of a Symposium on V1ra/ Hepatitis, Washington, DC, National Academy Press, 1975. 10. PLOTKIN, S. A. & KATZ, M. Mimmal infect1ve doses of viruses for man by the oral route. In: Berg, G., ed. Transmission of viruses by the water route, New York, Interscience Publishers, 1967. II. NATIONAL RESEARCH COUNCIL. Drinking water and health, vol. I, Washington, DC, National Academy Press, 1977. 12. GAMBLE, D. R. Viruses in drinkmg-water: Recons1deratwn of evidence for postulated health hazard and proposals for v1rolog1cal standards of purity. Lancet, 1: 425-8 (1979). 13. SuESS, M. ]., ed. Exammation of water for pollution control, vol. 3, Oxford, Pergamon Press, 1982. 14. Standard methods for the exammation of water and wastewater, 15th ed., Washington, DC, AMERICAN PUBLIC HEALTH ASSOCIATION, 1980. 15. KoTT, Y. E1 AL. Bacteriophages as viral pollution indicators. Water research, 8: 165171 (1974).
PART II. BIOLOGICAL ASPECTS
1. PROTOZOA 1.1 General description Of the intestinal protozoa pathogenic for man, three may be transmitted by drinking-water: Entamoeba histolytica, Giardia spp., and Balantidium coli. These organisms are the etiological agents of amoebiasis (amoebic dysentery), giardiasis, and balantidiasis, respectively, and they have all been associated with drinking-water outbreaks (1-4). Various, usually free-living, amoeba (e.g., Naegleria, Hartmannella, and Acanthamoeba spp.) can be waterborne agents of frequently fatal disease. However, waterborne infection with these organisms is almost always associated with recreational contact rather than with the drinking of water. E. histolytica is distributed worldwide and exists in trophozoite and cyst stages. Infection occurs by ingestion of cysts ranging in size from 10 to 20 J.Lm (average, 12 J.Lm). Since E. histolytica is primarily a parasite of primates, man is the reservoir of infection. Dysenteric individuals pass only trophozoites, which are environmentally susceptible to drying and to changes in temperature and salt concentration, and most or all of the parasites in this active amoeboid stage are destroyed by gastric juice (5). Consequently, chronic cases and carriers who excrete cysts are more important sources of infection. Various surveys throughout the world have indicated a prevalence of 0.8-50% for E. histolytica infections (6). Carrier rates during epidemics have been estimated at up to 63 '/~ (3). The average number of cysts passed per carrier per day has been estimated at 1.5 x 10 7 and the density of cysts in sewage, assuming a carrier rate of 50 'lo during an epidemic, has been calculated to be 5000 cysts/litre (3). Giardia spp. have worldwide distribution and are flagellates existing in trophozoite and cyst forms. In addition to man, Giardia has been found in numerous mammal species (7, 8) and in psittacine birds (9). The organism that infects man has been designated Giardia Iamblia, Lamblia intestinalis, or Giardia intestinalis. With the exception of G. muris, which occurs in mice, there are no features, morphological or otherwise, that allow differentiation of the flagellate occurring in different animal species. It had been considered that the organism was highly hostspecific but this has been questioned ( 10) and recent studies (8) indicate that other animals may act as reservoirs of infection for man. As with E. histolytica, infection occurs by ingestion of cysts. Trophozoites are found in the faecal material only when there is acute watery diarrhoea and it 39
40
II.
BIOLOGICAL ASPECTS
is believed these would not survive long in the environment outside an animal host. The cysts are ovoid and are 8-12 11m long by 7-10 11m wide. In various surveys that have been conducted throughout the world, the prevalence of Giardia infection in man has ranged from 2.4 to 67.5% (6). The infection is more common in children than in adults (7). An asymptomatic carrier state is common but the ratio of cases to carriers has not been determined. Mean daily production of cysts by an infected adult has been calculated to range from 2.1 x 10 8 to 7.1 x 10 8 cysts. The density of cysts in domestic raw sewage has been estimated at 10 000-24 000 cysts/litre at a prevalence rate in the human population of 10-25% (11). Beaver (Castor canadensis) have been implicated as the source of contamination in at least one community drinking-water outbreak in the USA (12). Balantidium coli is a ciliated organism with worldwide distribution and both the trophozoite and cyst stages can be infective for man (7). The spherical to ovoid cysts are 40-60 11m in diameter, are yellowish to greenish, and have a two-membrane cyst wall. Human infections usually occur as the result of ingestion of food or water contaminated with faecal material from infected swine (5). Other hosts include lesser primates and, rarely, the dog and rat. B. coli is very common in swine, with surveys indicating a prevalence of 21-100 ';:0 ( 7). It is considerably less prevalent in man, with twelve surveys throughout the world indicating a prevalence of 0. 77 %. Asymptomatic carrier infections can occur in man. Pathogenic Naegleria is the most frequently recognized etiologic agent of primary amoebic meningoencephalitis. Amoebae of the genus Naegleria exist in trophozoite, flagellate, and cyst forms (13). The trophozoites are usually slug-like or pear-shaped and range in size from 8 to 14 Jlm. Most have a single central nucleus but occasional bi- or multinucleated forms occur. Reproduction is by simple binary fission. The trophozoite will transform into a flagellate stage upon dilution of culture media with distilled water. The flagellate organism is actively motile, pear-shaped, and has 2-4 anterior flagellae. The cysts are circular and about 8-12 11m in diameter. Pathogenic species can be differentiated from non-pathogenic Naegleria species by intranasal instillation of the cultured amoebae into mice. 1.2 Routes of exposure The transmission of pathogenic intestinal protozoa to man occurs through any mechanism by which material contaminated with faeces containing viable organisms from infected individuals can reach the mouth. 1.2.1 Drinking-water Since man is the primary reservoir for infection with E. histolytica, the contamination of water supplies with domestic sewage can lead to the
1.
PROTOZOA
41
transmission of this organism through drinking-water. Outbreaks traced to sewage contamination of drinking-water have been reported (3). The potential for waterborne transmission may be somewhat greater in the tropics where the carrier rate often exceeds 50 ~~o as compared to more temperate regions where prevalence in the general population is generally less than 10\. The cysts can survive several months in water at 0 °C, 3 days at 30 oc, 30 minutes at 45 °C, and 5 minutes at 50 oc (5) and they are one of the most chlorine-resistant pathogens known (14). Waterborne outbreaks of giardiasis have been reported primarily from the USA. In only about 50% of the waterborne outbreaks in the USA has an etiological agent been identified. In those outbreaks where the nature of the etiological agent was established, Giardia Iamblia was the most commonly identified pathogen during the period 1972-77; 23 waterborne outbreaks of giardiasis have been reported in the USA since 1965 (1). Drinking-water has also been implicated as the vehicle of transmission in outbreaks occurring in travellers to the USSR (15, 16). Most of the outbreaks have been associated with untreated drinkingwater or water receiving disinfection only (2). The infective dose has been shown to be small, with 10 cysts administered in gelatin capsules causing infection in man (17, 18). Information on survival of the organisms in the environment and on resistance to disinfection is incomplete at the present time. However, there are indications that a small percentage of the cysts can survive freezing and that cysts can remain viable in drinking-water up to 77 days at 80 oc (19). Preliminary studies indicate that Giardia spp. fall somewhere between E. histolytica and enteric viruses in their resistance to inactivation by chlorine (14, 20). Considering the distribution of Giardia spp. in man and in a wide variety of domestic and wild animals, and the suggestion that the organisms may not be as host-specific as previously believed, Giardia spp. probably have the greatest potential for transmission through drinking-water of the three intestinal protozoa under discussion. The only reported waterborne outbreak of balantidiasis occurred in the Truk District of Micronesia in 1971 (4). It was concluded that the epidemic probably resulted from contamination of water supplies by pig faeces when a devastating typhoon destroyed pig pens and precarious water-catchment facilities. Transmission of amoebic meningoencephalitis has been recorded rarely in those bathing in domestic water supplies and a case occurred from nasal ablutions with water.
1.2.2 Food E. histolytica cysts may be transmitted by food, and raw vegetables may be a source of infection (7). Although the cysts rarely survive on the hands for more than 10 minutes, except under the fingernails, foodhandlers who are carriers may be an important source of transmission of the organism. Indirect evidence for this route of
42
II.
BIOLOGICAL ASPECTS
exposure is provided by one study where inspection and treatment of foodhandlers in Venezuela for E. histolytica infection decreased the amoebic dysentery rate from 36.84 to 0.61 per 1000 per year over a three-year period (7). Another study in China correlated transmission with the eating of cold bread served with the hands and concluded that transmission by foodhandlers is probably more important than other routes of exposure in that country. Contamination of food may also occur via houseflies, which have been shown to be capable of transmitting the organism in their droppings (5). Only one food-borne outbreak of giardiasis has been reported (21), and it is believed that an infected individual who prepared the food was the source of the contamination. Although Giardia cysts have been detected on strawberries (22) and on vegetables (23), the significance of food-borne transmission in the epidemiology of giardiasis is unknown. Levine (7) indicates that Balantidium cysts remain alive for weeks in pig faeces if they do not dry out and, again, it seems logical that if food becomes contaminated with faecal material from infected pigs or man, food-borne transmission could occur. 1.2.3 Air Airborne transmission of intestinal protozoan pathogens directly to man does not seem a likely route of exposure because of the susceptibility of the organisms to inactivation by drying. Although the spread of disease to the lungs and pleura is a complication in untreated amoebiasis patients with liver abscess, primary respiratory infection has not been reported and, if infection does occur as a result of inhalation of contaminated air with subsequent ingestion of mucus-trapped organisms, this route would seem to be several orders of magnitude less important than person-to-person, food, or water transmission. 1.2.4 Other routes of exposure (24, 25),
Sexual transmission of Giardia and E. histolytica has been reported especially among homosexuals. Freeman (5) states that although swimming-pools have not been definitely incriminated in amoebiasis, they are a potential source of E. histolytica infection. The significance of the recreational water route for transmission of Giardia infections has not been determined, but it is probably low compared with other routes of exposure. Apart from the one drinking-water outbreak of balantidiasis, the only other outbreaks have been in mental institutions and they were attributed to person-to-person spread due to poor personal hygiene and the habit of coprophagy among the patients 1.2.5 Relative significance of routes of exposure Of the three intestinal protozoan pathogens, E. histolytica is the most prevalent worldwide. In the USA, person-to-person spread appears to be
(4).
1.
PROTOZOA
43
the most common mode of transmission (26). In other parts of the world, contamination of food by infected foodhandlers appears to be the most significant means of transmission (7). A potential for waterborne outbreaks does exist and the maintenance of a high endemic level of infection in developing countries through drinking-water transmission appears possible, though it is by no means clear that this is so. The transmission of Giardia infection between children and adults seems to be rare where good personal hygiene is practised. However, the transmission of infections among preschool-age children in day-care centres (27) and similar institutions is probably common. Burke (28) has concluded that the prevalence of endemic infection in many countries is inversely related to socioeconomic level and prevalence is highest where substandard sanitary practices are found. In the USA, waterborne transmission is apparently a significant route of exposure, but evidence on the relative significance of different routes is lacking. The incidence of balantidiasis in humans is low and direct contact with pigs appears to be the main route of transmission of the causative organism. The potential exists for transmission of the organism in food and water contaminated with pig faeces. Almost all recorded cases of amoebic meningoencephalitis result from recreational rather than domestic use of water, but epidemiological evidence from developing countries is very scarce.
1.3 Health effects Though most infections with E. histolytica are asymptomatic or cause only minor symptoms, fatalities can occur (26). The usual clinical manifestations are gastroenteritis with symptoms ranging from mild diarrhoea to fulminating bloody dysentery. Liver abscess is the most common metastatic complication. Pathogenicity appears to depend on strain virulence and on host factors, including nutritional status of the individual and associated bacterial flora (7). Giardiasis symptoms range from mild self-limiting enteritis to chronic debilitating diarrhoea and asymptomatic infections occur (28). Acute, subacute, and chronic stages of infection have been described (29-31). Mortality from Giardia infection has not been reported in man or any other animal, with the possible exception of psittacine birds (9). Balantidiasis can present as an acute bloody dysentery (5) but an asymptomatic carrier state also occurs in man (7). Man is highly resistant to infection and when the disease does occur it is usually mild and self-limiting (4). The amoeba Naegleria fow!eri is a widely distributed amoebo-ftagellate found in soil and fresh water (32). In the last decade, it has been shown in both tropical and temperate countries to give rise to a lethal meningoencephalitis and about 100 cases have been described (33), usually in those who have been swimming in natural waters. The portal
44
II.
BIOLOGICAL ASPECTS
of entry is the nasopharynx from which the amoebae enter the brain by penetration of the olfactory mucosa and cribriform plate. Knowledge about these organisms is still incomplete, but it is clear that the majority are free-living, that multiplication can take place during storage of warm water, and that although infection in man is very rare in relation to the people exposed to possible infection, the prognosis for clinical cases is very bad and treatment difficult (34). 1.4 Monitoring A monitoring programme is not recommended because no standard quantitative or qualitative methods are available. Experimental methods for the concentration and detection of E. histolytica and Giardia cysts are available ( 11, 35) but they are recommended for use only in conjunction with concurrent epidemiological studies on endemic or epidemic occurrences. The methods available at present are inefficient; the concentration techniques are not reproducible; the identification of organisms in concentrated samples is difficult, and at least in the case of Giardia spp., the viability and origin of detected cysts (11) cannot be determined. In addition, no recommendations can be made regarding the frequency of sampling. In situations where disease outbreaks occur from drinking water contaminated with pathogenic intestinal protozoa, boiling of water may provide effective control for inactivation of Giardia (19), E. histolytica, and B. coli. Attempts should be made to identify and remove sources of contamination. A sanitary survey should be conducted to identify and correct treatment and distribution system deficiencies. Epidemiological data should be gathered for correlation with physical and chemical data on the treatment plant and distribution system. Guidelines for investigating waterborne illness are available (36). The collection of this information could be important in preventing or controlling further outbreaks. 1.5 Rationale for recommendation As a group, the intestinal pathogenic protozoa occur in large numbers in the faeces of infected individuals in man and a wide variety of domestic and wild animals. The infective dose in man for Giardia Iamblia is as low as 10 cysts ingested orally, and for both Giardia and E. histolytica, it has been presumed that one viable cyst may establish infection in a susceptible host (17, 37). Simple methods for readily detecting, identifying, and counting intestinal protozoa in drinking-water are not available (11). In view of these considerations, protection of sources, adequate treatment, and ensuring integrity of water in distribution are the most effective ways of preventing infection. Coliform organisms do not appear to be a good indicator for Giardia or E. histolytica in treated water because of the increased resistance of
1.
PROTOZOA
45
these protozoans to inactivation by disinfection. In non-disinfected water, the presence of indicator bacteria could suggest the presence of pathogenic protozoa. Previous studies have shown that diatomaceous earth filtration is effective for removal of cysts of E. histolytica and Giardia and that E. histolytica can also be effectively removed by granular media filtration (38). Diatomaceous earth achieved up to 99.998% removal of Giardia cysts, and a recent study (39) indicated that removal of over 99.99% of Giardia cysts is possible using granular media filtration with attention given to coagulant dose, filtration rate, turbidity, and backwash procedures. The basis for recommending coagulation, sedimentation, and filtration if protozoa are to be removed from surface waters and unprotected groundwater takes into consideration the inadequacy of the coliform indicator group for this group of organisms, the lack of standard methods for detection, the resistance of the organisms to disinfection, and the demonstrated effectiveness of properly applied filtration for removal of pathogenic intestinal protozoa.
2. HELMINTHS A great variety of helminth eggs and larvae have been detected in drinking-water and it is clear that all those infective to man should be absent if the drinking-water is to be safe. However, the vast majority of such helminths are not primarily waterborne and it is neither feasible nor necessary to monitor water for them on a routine basis. Two groups of helminths are more directly related to water supply: those transmitted wholly by the ingestion of infected copepod intermediate hosts (Group I) and those whose cercariae are directly infective to man (Group II). Most of the remaining species are grouped into a third category (Group III). This cuts across the formal taxonomy of the organisms involved, which belong to two animal phyla, the Nemathelminthes or roundworms and the Platyhelminthes or flatworms. Within the latter phylum are the Trematoda (flukes) and Cestoda (tapeworms). Helminths that may potentially be transmitted by drinking-water are listed in the table on p. 47. 2.1 Group I (Dracunculus, Spirometra) 2.1.1 General description Group I comprises helminths developing in aquatic copepods and acquired by man drinking water containing the intermediate host crustacea. The most important member of the group is Dracunculus medinensis, the guinea-worm, a filarial parasite of man. Female worms mature in the deep tissues and then migrate to lie subcutaneously in a limb. Numerous larvae develop within the body of the female, inducing a blister in the skin of the host, which breaks down. The female exposes her prolapsed uterus there and the larvae are discharged whenever water is sensed by the worm. The life cycle continues if these rhabditiform larvae reach water containing copepods of the genera Cyclops, Eucyclops, Mesocyclops, and Macrocyclops, which ingest the larvae. Development to the third-stage larvae takes place in the body of the copepod and these are infective to man if ingested (40). It follows that Dracunculus is particularly transmitted at unprotected wells and pools rather than through piped water supplies. Infection occurs in a patchy distribution in Africa, especially the West African Sahel, west Asia, and as far east as Afghanistan, India, and south USSR, as well as Indonesia. Man is the predominant definitive host. Tapeworms of the genus Spirometra, though much rarer in man, also 46
2.
HELMINTHS
47
Helminths potentially transmitted by drinking-water Main category Tremoda (flukes) Name Group• Ill Ill Ill II II II II II Ill Ill Ill Ill I I I I Ill II II Ill I Ill II Ill Ill Ill Ill
Fasciola gtganttca Fasciola hepatica Fasciolopsis buski Schistosoma haematobium Schistosoma intercalatum Schistosoma Japonicum Schistosoma mansoni Schistosoma mekongi Echinococcus granulosus Echinococcus multilocularis Hymenolepsts nana Multiceps sp. Spirometra mansoni Spirometra mansonoides Spirometra proliferum Spirometra theileri Taenia solium Ancylostoma brasiliense Ancylostoma duodenale Ascaris lumbflcotdes Dracunculus medmensis Enterobtus vermiculafls Necator americanus Strongylotdes stercora/is Toxocara spp. Trichuris trichiura Uncinaria stenocephala
Crestoda (tapeworms)
Nematoda (roundworms)
• Note that except for Dracunculus, very few mdeed of these helmmths are ch1efly transmitted through dr1nkmg-water.
have a stage in aquatic copepods. Adult worms are found in the small intestine of cats. Eggs of Spirometra spp. pass out in the faeces and hatch in water to produce coracidium larvae which are ingested by copepods, in which they develop to the procercoid stage (41, 42). If man ingests the copepod, the larvae undergo further development in the tissues to the plerocercoid or sparganum stage. This stage may also infect man from the flesh of an alternative intermediate host used as a poultice on lesions, as is the practice in parts of Asia. Species of Spirometra are found patchily distributed in the Americas between the USA and Uruguay, and also in east Asia (44, 45), Kenya, and the United Republic of Tanzania (43).
2.1.2 Routes of exposure Drinking-water containing infected copepods is the only source of infection with Dracunculus and this is the only animal parasite that can be eradicated solely by provision of safe drinking-water. Spirometra is
48
II.
BIOLOGICAL ASPECTS
an occasional parasite of man and may reach him by other routes than drinking-water.
2.1.3 Health effects Guinea-worm infection, or dracontiasis (dracunculosis), is a major disabling disease especially in agricultural communities in the Sahel and the Indian subcontinent. The pain of the infection and arthritis of the nearest joint to the worm immobilizes the infected person for several weeks. Since maturation of the worm closely coincides with the rainy planting season, the disability has a disproportionate effect on productivity. Profound allergic reactions may accompany migration of the female, including erythema, urticaria, and intense pruritus, while systemic symptoms may include vomiting and diarrhoea. Sepsis of the blister may occur and an abscess may develop following rupture of the worm during attempted extraction. Fibrosis is sometimes a sequel and ectopic worms may also lead to abscess formation in other viscera. Dracontiasis as a public health threat is a disease of major importance locally, and the local prevalence may exceed 30%. Sparganosis is a much less common disease in which the lesions result from subcutaneous larvae which can cause oedema and inflammation, especially upon death of the parasite. Serious consequences follow if the eye is involved, a complication that occurs particularly in South-East Asia.
2.1.4 Monitoring Dracontiasis is a problem of small unpiped water supplies (e.g., step wells or reservoirs) where regular monitoring is often impracticable. Investigation of such a situation involves the study of the prevalence of infection in man and the collection of copepods using a plankton net or other container, followed by microscopic examination to detect the parasitic filarial larvae. Since prevention is achieved by protecting the water source (see below), the preferred method of monitoring is to determine that sources have been protected. Monitoring for Spirometra is not feasible nor indicated.
2.1.5 Rationale for recommendation A single infected copepod containing a single larva is capable of infecting man with sparganum or with Dracunculus, though the worm burden will depend on the number of infective larvae ingested and their sex. Since one fertilized adult female guinea-worm can cause severe disease, infective stages should be absent from drinking-water. As this is the sole route of transmission to man for Dracunculus, it is a matter of importance. In view of the way that rhabditiform larvae reach the copepods by being washed into wells from the limbs of those drawing
2.
HELMINTHS
49
water, it is clear that source protection is the best approach to prevention. The use of well surrounds that rise above ground level and drain away from the well usually suffices, though capping the well and fixing a pump is preferable. In emergency situations the infected copepods may be killed by addition of temephos (Abate) granules to wells in the doses required for insect larva control (46).
2.2 Group II (Schistosoma, Ancylostoma, Necator) 2.2.1 General description Group II comprises a miscellaneous group of flukes and roundworms whose infective larvae are able to penetrate the human skin and mucous membranes. They may therefore be transmitted through drinking-water, but are more of a hazard when water is used for washing or bathing. The main genus concerned is Schistosoma. The schistosomes infecting man belong to three main species: S. haematobium, which infects the veins of the vesical plexus and occurs chiefly in Africa and west Asia; S. mansoni, which is found in Africa, parts of South and Central America, and certain Caribbean islands; and S. japonicum, which is found in China, Indonesia, the Philippines, and other parts of east Asia. Both S. mansoni and S. japonicum are found in the portal venous system. S. intercalatum occurs in west central Africa and S. mekongi, a schistosome resembling S. japonicum, is found in the Mekong River Basin in South-East Asia. Indeed, it is now being recognized that within each of the major schistosome species there are strain differences related to geographic distribution and host variations. The adult worms are long-lived and the sexes are separate. The numerous eggs pass through the blood vessel walls and tissues to escape in the urine (S. haematobium) or faeces (other species). The eggs hatch on reaching freshwater liberating miracidia which penetrate an appropriate aquatic snail host and undergo development and multiplication for a month or more, following which the cercariae, or larvae infective to man, are shed into the water, and swim about using their bifurcated tails. They are just visible to the naked eye. They rapidly penetrate human skin in contact with the infected water and migrate and mature in the body to complete the cycle. Cercariae of non-human schistosomes and other related flukes may attempt to penetrate the human skin and die there, producing a rash and intense irritation known as schistosome dermatitis. S. japonicum infects a variety of domestic and wild animals as well as man. The human hookworms Ancylostoma duodenale and Necator americanus, both with a wide tropical and subtropical distribution, have eggs that hatch and develop in the soil to the third-stage larvae, which reinfect man by penetrating the skin. Hookworms of domestic animals may also invade man and their larvae wander in the skin producing local symptoms.
50
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BIOLOGICAL ASPECTS
2.2.2 Routes of exposure Schistosome infections are acquired when infected water is used for domestic activities, bathing, or washing. Ingested cercariae can penetrate the buccal mucous membranes, but this is a minor route of entry (47). The relevance of drinking-water is that it is used for washing if readily available and the benefits of safe water are gained only if there is reduced contact with the infected sources previously used. While the possibility of piped untreated surface water transmitting schistosomiasis is real, most transmission is from unpiped sources such as pools, wells, and also cisterns used for religious ablutions. Schistosome dermatitis is a hazard of recreational and occupational water use rather than drinking-water. Ancylostoma larvae have been shown to be infective in drinking-water and this may be an appreciable but not major route of transmission (48, 49).
2.2.3 Health effects The human schistosomes are a cause of severe morbidity and sometimes death in a number of the 200 million people infected worldwide. Pathology is due mainly to the host's reaction to eggs that failed to escape. Primary lesions are mainly in the liver, intestine, and around the bladder, but the most severe consequences are due to secondary damage of the upper urinary tract, to bladder cancer, and to liver fibrosis and its haemodynamic consequences. Hookworms chiefly lead to iron-deficiency anaemia, while the other helminths in this group produce skin lesions.
2.2.4 Monitoring Detection of schistosome cercariae in water is a research procedure unsuited to routine monitoring. The cercariae may be concentrated from water by filtration or detected by immersing small suitable rodents, allowing time for development, and dissecting them. Cercariae on paper, glass fibre, or acetate filters tend to lose their shape and species identification is not reliable, though various differential staining methods have been devised. It is more feasible to look for appropriate vertebrate hosts and determine infection rates in them.
2.2.5 Rationale for recommendation Since a single cercaria is infective, cercariae should be absent from drinking-water and there is no safe level. In the absence of routine monitoring assays, reliance has to be placed on preventive measures if a significant risk from drinking-water is suspected in an area. The cercariae have a free-living life of under 48 hours and storage for this period renders water safe (50). It is likely that storage for 24 hours will greatly reduce infectivity. Slow sand filters, provided they are properly
2.
HELMINTHS
51
operated, will remove the majority of cercariae, and disinfection at a residual level of 0.5 mg of free chlorine per litre for 1 hour will kill cercariae of the human schistosomes (51). A sounder approach is to use a source that does not contain the host snails and that is not subject to excretal contamination.
2.3 Group III (Ascaris, Trichuris, Strongyloides, Enterobius, Fasciolids, Hymenolepis, Echinococcus) 2.3.1 General description A large number of helminths have resistant eggs or cysts infective to man. If these gain access to drinking-water and are ingested, man becomes infected. In the case of all the species listed in this group in the table on p. 47, other methods of transmission, such as food and direct faecal-oral routes are far more important than the drinking-water route and for practical purposes safe disposal of excreta is of major importance, while attention to drinking-water alone will not significantly reduce transmission in most parts of the world. The most widespread intestinal helminths, Ascaris lumbricoides and Trichuris trichiura, produce resistant eggs of characteristic appearance within which the embryo must undergo development in the external environment before it becomes infective. When eggs are ingested by man, the larvae emerge in the intestine and undergo complex migrations before returning there to mature and lay eggs, at the rate of some 200 000 daily. The eggs are excreted in the faeces and, being heavy, settle relatively rapidly in water. Less ubiquitous are Strongyloides stercora/is, whose larvae are the infective form, Enterobius, the pinworm, whose sticky, less resilient eggs are more suited to direct faecal-oral transmission, and some animal nematodes that undergo limited development in man. The liver flukes of the genera Fasciola and Fasciolopsis that can infect man and other mammals develop in snails and the emerging cercariae then encyst on water plants; man is infected by eating the plants or peeling them with his teeth. The cysts may gain access to drinking-water (45). The human tapeworms of the genera Hymenolepis, with a direct life cycle, and Echinococcus, where man is infected by ingesting eggs usually acquired from dogs, have the potential for spread in drinking-water.
2.3.2 Routes of exposure These helminths all have a faecal-oral transmission and the infective stages tend to be of fairly high relative density. In no case is drinkingwater the predominant vehicle of transmission, although the eggs do get into water from time to time, particularly those of the common Ascaris and Trichuris.
52 2.3.3 Health effects
II.
BIOLOGICAL ASPECTS
The intestinal helminths produce a great range of symptoms. Many infections are subclinical, a few are fatal. It is likely that the majority produce low-grade chronic effects that are hard to quantify in the individual but are collectively important. Diversion of nutrients from the host to feed the worms is considerable. The intestinal helminths make up by their very high prevalence (52) for the limited pathology they produce in many individuals infected.
2.3.4 Monitoring While in outbreak or research situations helminth eggs and larvae may be extracted from drinking-water by filtration and most eggs can be identified microscopically, these procedures are unsuitable for routine use and the frequency of water contamination does not justify them.
2.3.5 Rationale for recommendation A single fertilized egg, mature larva, or encysted cercaria can cause infection. Therefore these should be absent from drinking-water and this is best achieved by protecting the source from faecal contamination. Should they gain access to raw water, most will be removed by filtration procedures, especially using slow sand filters, while all are relatively resistant to chlorination, especially Ascaris (53).
3. FREE-LIVING ORGANISMS 3.1 General description The free-living organisms generally considered significant in water supplies include plankton and macroinvertebrates and these are the groups discussed in detail in this section. Plankton consists of microscopic or small organisms that live primarily in suspension in the water column. Phytoplankton consists of the free-living bacteria, fungi, and algae. The algae are chlorophyll-bearing organisms, having different colours. They are autotrophic, unicellular or colonial, and motile or non-motile. The fungi and bacteria are largely heterotrophic. Zooplankton consists of free-living protozoa, rotifers, cladocera, copepods, worms and, in their early stages of development or during brief excursions, the larvae of some aquatic insects and fish. The macroinvertebrates include the larger benthic animals, such as aquatic insect larvae, crustacea, and gastropods. Plankton organisms are important in water supplies because they interfere with water-treatment processes, produce substances toxic to humans, harbour human pathogens, and contribute organic matter, which may be converted to halogenated organic compounds during chlorination. The macroinvertebrates may affect the efficiency of water distribution systems and the acceptance of water by consumers. The algae may present problems because of their occurrence in raw surface water, and uncovered storage reservoirs containing treated (finished) water. In addition to these sites, zooplankton may occur in open wells. The occurrence of macroinvertebrates related to water supply problems is limited to distribution systems. The abundance and species composition of algae in surface water supplies are governed by a combination of natural environmental conditions (54). Maximum algal counts in surface water range from only a few to many millions of organisms per ml. The algae in surface waters are most frequently dominated by the diatoms, followed by the coccoid green and blue-green algae. Winter populations are generally dominated by the pennate diatoms, whereas in summer, the centric diatoms, coccoid greens, and blue-greens are most abundant. In late summer, or where temperatures remain above 25 oc throughout the year, the bluegreen algae often comprise the bulk of the algal biomass (55-58). Extensive studies of algae in water supplies have been published by Palmer (59). 53
54
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BIOLOGICAL ASPECTS
A wide variety of free-living organisms has been found in water supply distribution systems, including fungi, algae, protozoa, rotifers, worms, water fleas, shrimps and other crustacea, spring-tails, midge larvae, mussels, and snails. Initial reports on these occurrences were based on investigations following complaints of the appearance of animals at consumers' taps. More recently, systematic studies have revealed the presence of organisms in numerous pipe networks even where there was no obvious reason to suspect their presence. It is now believed that most systems will contain some animals and a list of organisms found to infect certain water mains has been compiled by Collingwood (60).
3.2 Health effects There is mounting evidence that some toxic substances produced by some algae in water supplies may have a significant adverse effect on public health. Poisonous species occur in two major groups of algae: yellow-greens (Xanthophyta) and blue-greens (Cyanophyta). The toxic blue-green algae are the most important in fresh water supplies (61-64).
Toxic substances released by algae may penetrate the treatment works. In laboratory experiments, aluminium coagulation, filtration, and chlorination were not effective in the removal of algal toxins. Even treatment by activated carbon in quantities similar to those used in water treatment was unsuccessful (65). The occurrence of public health problems related to algae is quite limited. A relationship between high concentrations of blue-green algae and outbreaks of gastroenteritis in humans has been reported in India (69), the Philippines (68), and the USA (66, 67). Substances similar to the endotoxins of Gram-negative bacteria have been isolated from several blue-green algae (70). One of the most severe recent outbreaks of waterborne gastroenteritis, which affected approximately 5000 persons in Sewickley, Pennsylvania (71, 72), was linked to a bloom of the ubiquitous filamentous blue-green alga, Schizothrix calcicola, in open finished-water reservoirs. Algae and theu extracellular products can represent a large part of the organic matter in surface water supplies and may be important sources of precursors in the formation of trihalomethanes during chlorination (73-76).
The free-living animals that have been reported to occur in water mains in temperate countries are not known to cause disease directly, but some have been shown to ingest and protect pathogenic organisms under laboratory conditions (77). Where domestic water is stored on site it is essential to cover it and prevent breeding of mosquito vectors of disease in the storage containers.
3.
FREE-LIVING ORGANISMS
55
3.3 Other effects-taste and odour, colour and turbidity, interference with treatment, and infestation of distribution systems Algae in water supplies may be aesthetically unpleasant, and may interfere with water treatment by increasing the chlorine demand, causing taste and odour problems, and clogging filters (59, 78). The growth of either aquatic plants or animals may constitute a natural source of odour and taste. In surface waters, algae constitute the main problem, whereas animal forms may proliferate in underground waters, reservoirs, or pipelines. Many algae secrete oils that are liberated either during metabolic activity or when dead cells disintegrate. These oils impart typical odours and tastes to the water (79, 80). Coloration and turbidity may be a problem (81) in unpiped and treated waters when the reservoirs are open or when the treatment is insufficient. In addition, the growth of algae and other organisms can interfere with the maintenance and operation of water-treatment systems by clogging filters (82, 83).
3.4 Monitoring Adequate methods are available for the collection and analysis of freeliving organisms in water supplies (35, 84-89).
3.4.1 Sampling Algae and other microorganisms may undergo rapid temporal changes, varying significantly in abundance and species composition from day to day (90). The microbial biomass and species composition should therefore be determined frequently if the data are to be used effectively to modify water-treatment processes, control taste~ ,-and odours, and detect harmful concentrations of microorganisms in finished water supplies.
3.4.2 Biomass The size, shape, and volume of different algae differ greatly, and counts alone do not provide an accurate estimate of the amount of organic matter (biomass) contributed by each taxon. The biomass is commonly estimated by measuring cell surface area, cell volume, or chlorophyll content of the algae (35, 84, 85).
3.4.3 Endotoxins The Limulus amoebocyte lysate test for endotoxins is relatively simple and is being used with increasing frequency (70, 72, 91). However, this test does not differentiate between endotoxins of bacterial and algal origin; a quantitative relationship between endotoxin concentrations and
56
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health effects has not been established; and the significance and interpretation of test results remain to be determined. 3.5. Rationale for recommendation Organisms present in water supplies may cause adverse effects on health, aesthetic problems, objectionable odours and taste, and can also interfere with water treatment. Although the organisms found to infest distribution systems in temperate climates are not associated with known adverse health effects, it is desirable for aesthetic reasons that their appearance at consumer taps be minimized. The present state of the art does not permit the establishment of maximum allowable limits, but it is recommended that, wherever possible, free-living organisms be removed from drinking-water. This may be achieved by protection of the source, implementing good practices of treatment, periodic and systematic swabbing and flushing of pipelines, and monitoring of water quality. REFERENCES I. CRAUN, G. F. Waterborne giardiasis in the United States: a revtew. American journal of public health, 69: 817-819 (1979). 2. CRAUN, G. F. Waterborne outbreaks of giardiasis. In: Jakubowski, W. & HolT, J. C., ed. Waterborne transmission of giardwsis, Cincinnati, US Environmental Protection Agency, 1979 (EPA-600/9-79-001). 3. CHANG, S. L. & KABLER, P. W. Detection of cysts on Entamoeba histolytica in tap water by the use of membrane filter. American journal of hygiene, 64: 170-180 (1956). 4. CENTER FOR DISEASE CONTROL. Balantidiasis - Truk District Micronesia. Morbidity and mortality weekly report, 21: 59 (1972). 5. FREEMAN, B. A. Burrows textbook oj m1crobwlogy, 21st ed. Philadelphia, Saunders, 1979. 6. BELDING, D. L. Textbook of parasitology, 3rd ed. New York, Meredith Publishing Co., 1965. 7. LEVINE, N. D. Protozoan parasites of domestic animals and of man, 2nd ed. Minneapohs, Burgess Publishmg Co., 1973. 8. DAVIES, R. B. & HIBLER, C. P. Ammal reservoirs and cross-species trarismtsswn. In: Jakubowski, W. & HolT, J. C., ed. Waterborne transmisswn of giardiasis, Cmcmnat1, US Environmental Protection Agency, 1979 (EPA-600/9- 79-001 ). 9. PANIGRAHY, B. ET AL. Zoonotic diseases in psittacine birds: Apparent increased occurrence of chlamydiosis (psittacosis), salmonellosis and g1ard1asis. Journal of the American Veterinary Medical Assocwtion, 175: 359-361 (1979). 10. FILICE, F. P. Studies on the cytology and life history of a Giardia from the laboratory rat. University of Californta publications in zoology, 57: 53-145 (1952). II. JAKUBOWSKI, W. & ERICKSEN, T. H. Methods for detecting Giardw cysts in water supplies. In: Jakubowski, W. & HolT, J. C., ed. Waterborne transmission of giardiasis, Cincinnati, US Environmental Protection Agency, 1979 (EPA-600/9-79-001). 12. KIRNER, J. C. ET AL. A waterborne outbreak of giardiasis in Camas, Washington. Journal of the American Water Works Association, 70: 35-40 (1978). 13. LOCKEY, M. W. Primary amebic meningoencephalitis. Laryngoscope, 88: 484-503 (1978). 14. HoFF, J. C. Disinfection resistance of Giardia cysts: Origins of current concepts and research in progress. In: Jakubowski, W. & HolT, J. C., ed. Waterborne transmission of giard1asis, Cincinnati, US Environmental Protection Agency, 1979 (EPA-600(9-79-001).
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15. JoKIPJJ, L. & JoKIPII, A. M. M. Giardiasis in travelers: A prospective study. Journal of infectzous diseases, 130: 295-299 (1974). 16. CENTER FOR DISEASE CONTROL. Giardiasis- in residents of Rome, N.Y., and in travelers to the Soviet Union. Morbidity and mortality weekly report, 24: 371 (1975). 17. RENDTORFF, R. C. The experimental transmission of human intestinal protozoan parasites. II. Giardia Iamblia cysts given in capsules. American journal of hygiene, 59: 209-220 (1954). 18. RENDTORFF, R. C. The experimental transmissiOn of Giardia Iambiia among volunteer subjects. In: Jakubowski, W. & Hoff, J. C., ed. Waterborne transmission of giardiasis, Cincinnati, US Environmental Protection Agency, 1979 (EPA-600/9-79-001). 19. BINGHAM, A. K. ET AL. Giardia spp.: Physical factors of excystation in vitro, and excystation vs. eosin exclusion as determinants of viability. Experimental parasitology, 47: 284-291 (1979). 20. HOFF, J. C. ET AL. Inactivation of Giardia muris cysts by chlorine. In: Abstracts from the 79th Annual Meeting of the Amerzcan SoCiety for Microbiology, Los Angeles, ASM, 1979. 21. OsTERHOLM, M. T. ET AL. An outbreak of foodborne giardiasiS. New England journal of med1cine, 304: 24-28 (1981). 22. JACKSON, G. Comment made in open discussion of paper by G. R. Healy on the presence and absence of Giardia Iamblia in studies on parasite prevalence in the USA. In: Jakubowski, W. & Hoff, J. C., ed. Waterborne transmission of giardiasis, Cincinnati, US Environmental Protection Agency, 1979 (EPA-600/9-79-001). 23. T AY, J. ET AL. Search for cysts and eggs of human intestinal parasites in vegetables and fruits. In: Conference on Cooperative Research Needs for the Renovation and Reuse of Municipal Wastewater in Agriculture, 15-19 December 1980, Cocoyoc, Morales, Mexico (Sponsored by the Institute of Water Research, East Lansing). 24. MILDVAN, D. ET AL. Venereal transmissiOn of enteric pathogens in male homosexuals. Journal of the American Medical Association, 238: 1387-1389 (1977). 25. HURWITZ, A. L. & OwEN, R. L. Venereal transmissiOn of intestinal parasites (medical information). Western journal of medzcine, 128: 89-91 (1978). 26. KROGSTAD, D. J. ET AL. Current concepts in parasitology-amebiasis. New England JOUrnal of medicine, 298: 262-265 (1978). 27. BLACK, R. E. ET AL. Giardiasis in day-care centers: Evidence of person-to-person transmission. Pediatrics, 60: 486-491 (1977). 28. BURKE, J. A. The clinical and laboratory diagnosis of giardiasis. CRC critical reviews in clinical laboratory sciences, 7: 373-391 (1977). 29. WoLFE, M. S. Giardiasis. Journal of the Amerzcan Medical Association, 233: 13621365 (1975). 30. WoLFE, M. S. Current concepts in parasitology-giardiasis. New England JOurnal of medicine, 298: 319-321 (1978). 31. WoLFE, M. S. Managing the patient w1th giardiasis: Clinical, diagnostic and therapeutic aspects. In: Jakubowski, W. & Hoff, J. C., ed. Waterborne transmission of giardiasis, Cincinnati, US Environmental Protection Agency, 1979 (EPA-600/9-79-001). 32. LAWANDE, R. V. The seasonal incidence of primary amoebic meningoencephalitis in northern Nigeria. Transactions of the Royal Society of Tropical Medicine and Hygiene, 74: 141 (1980). 33. THONG, Y. G. Primary amoebic meningoencephalitis: fifteen years later. Medical JOUrnal of Australia, 1: 352-354 (1980). 34. KASPRZAK, W. ET AL. Studies on some pathogenic strains of free-living amoebae isolated from lakes in Poland. Annates de Ia Societe beige de Medecine tropica/e, 54: 351-357 (1974). 35. AMERICAN PUBLIC HEALTH ASSOCIATION. Standard methods for the examination of water and wastewater, 15th ed., Washington, DC, 1980. 36. INTERNATIONAL ASSOCIATION OF MILK, fOOD, A)IID ENVIRONMENTAL SANITARIANS, INC. Procedures to investigate waterborne illness, 1st ed. Ames, lA, IAMFES, 1979. 37. BEAVER, P. C. ET AL. Experimental Entamoeba histolytica infections m man. American JOUrnal of tropical medicine and hygiene, 5: 1000-1009 (1956).
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38. LOGSDON, G. S. ET AL. Water filtration techniques for removal of Gtardta cysts and cyst models. In: Jakubowski, W. & Hoff, J. C., ed. Waterborne transmission of giardiasis, Cincmnati, US Environmental Protection Agency, 1979 (EPA-600/9-79001). 39. LOGSDON, G. S. ET AL. Alternative filtration methods for removal of Gtardia cysts and cyst models. Journal of the American Water Works Association, 73: 111-118 (1981). 40. MULLER, R. Dracunculus and dracunculiasis. Advances in parasitology, 9: 73-151 (1971). 41. MUELLER, J. F. The laboratory propagation of Spirometra mansonoides (Mueller, 1935) as an experimental tool. II. Culture and infection of the copepod host, and harvestmg the procercoid. Transactions of the Amerzcan Microscopical Society, 78: 245-255 (1959). 42. MuELLER, J. F. Host-parasite relationships as illustrated by the cestode Sp1rometra mansono1des. In: McCauly, J. E., ed. Host-parasite relationships, Corvallis, Oregon State University Press, 1966 (Proceedings of the 26th Annual Biology Colloquium, Corvallis, Oregon, 23-24 April 1965). 43. 0PUNI, E. K. & MULLER, R. L. Studies on Spirometra thei/en (Baer, 1925) n. comb. 1. Identification and biology in the laboratory. Journal of helmmthology, 48: 15-23 (1974). 44. FAUST, E. C. ET AL. Animal agents and vectors of human disease, 4th ed. Philadelphia, Lea & Febiger, 1975. 45. FAUST, E. C. ET AL. Craig and Faust's clinical parasllology, 8th ed. Philadelphia, Lea & Febiger, 1970. 46. MULLER, R. Laboratory experiments on the control of cyclops transmntmg gumea worm. Bul/etm of the World Health Organization, 42: 563-576 (1970). 47. MALDONADO, J. F. & PERKINS, K. W. Sch1stosomwsis in America. Barcelona, Editorial Cientifico-medica, 1967. 48. OKAMOTO, K. An experimental study of the migratiOn route and development of Ancylostoma duodenale in pups after oral infection. Journal of the Kyoto Prefectural Medical University, 70: 135-152 (1961). 49. HIGO, A. An experimental study on the migration route and the development of Ancylostoma duodenale in pups after cutaneous infection. Journal of the Kyoto Prefectural Medical Unwersity, 70: 851-874 (1962). 50. JONES, M. F. & BRADY, F. J. Survival of Schistosoma japonicum cercanae at vanous temperatures in several types of water. National Institute of Health bul/etm, 189: 131136 (1974). 51. CoLES, G. C. & MANN, H. Schistosomiasis and water works practice in Uganda. East African medical journal, 48: 40-43 (1971). 52. STOLL, N. R. Thiiil wormy world. Journal of parasllology, 33: 1-18 (1947). 53. KELLER, P. Sterilization of sewage sludges. I. A review of the literature pertaining to the occurrence and viability of parasitic ova m sewage with special reference to Ascans lumbricoides. Journal and proceedings. Institute of Sewage Purification, Part I, 92-99 (1951). 54. LUND, J. W. G. The ecology of the freshwater phytoplankton. Biological reviews of the Cambridge Philosophical Society, 40: 231-293 (1965). 55. US PUBLIC HEALTH SERVICE. Annual Compilation of Data. National Water Quality Network, Basic Data Branch, Division of Water Supply and Pollution Control, Department of Health, Education and Welfare, Cincinnati, OH (1957-1963). 56. WEBER, C. I. ET AL. Phytoplankton. In: Flynn, K. C. & Mason, W. T., Jr., ed. The freshwater Potomac, aquatic commumties and environmental stresses, Interstate Commission on the Potomac River Basin, Rockville, MD, 1978. 57. WILLIAMS, L. G. & ScoTT, C. Principal diatoms of major waterways of the United States. Limnology and oceanography, 7: 365-379 (1962). 58. WILLIAMS, L. B. Possible relationships between plankton-diatom species numbers and water quality estimates. Ecology, 45: 809-823 (1964). 59. PALMER, C. M. Algae and water pollution. Cincinnati, US Environmental Protection Agency, 1977.
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60. COLLINGWOOD, R. W. Occurrence, significance and control of organisms m distribution systems. Barcelona, International Water Supply Association Congress, 1966, 13 pp. 61. GORHAM, P. R. Toxic algae. In: Jackson, D. F., ed. Algae and man, New York, Plenum Press, 1964, pp. 307-336. 62. SCHWIMMER, M. & SCHWIMMER, D. Algae and medicine. In: Jackson, D. F., ed. Algae and man, New York, Plenum Press, 1964, pp. 368-412. 63. SCHWIMMER, M. & SCHWIMMER, D. Medical aspects of phycology. In: Jackson, D. F., ed. Algae, man and the environment, Syracuse, Syracuse University Press, 1968, pp. 279-358. 64. MooRE, R. E. Toxins from blue-green algae. BioSczence, 27: 797-802 (1977). 65. INGRAM, W. M. & PRESCOTT, B. W. Toxic fresh water algae. The American midland naturalist, 52: 75-87 (1954). 66. TISDALE, E. S. Ep1demic of mtestinal disorders in Charleston, West Virginia, occurring simultaneously with unprecedented water supply conditions. Amencan journal of public health, 21: 198-200 (1931). 67. VELDEE, M. V. Epidemiological study of suspected waterborne gastroenteritis. American journal of public health, 21: 1227-1235 (1931). 68. DEAN, A. G. & JONES, T. C. Seasonal gastroenteritis and malabsorption at an American military base in the Philippines. American journal of epzdemw/ogy, 95: 111127 (1972). 69. GuPTA, R. S. & DASHORA, M. S. Algal pollutants and potable water. In: Pajasek, R. B., ed. Drinking water quality enhancement through source protection, Ann Arbor, Ann Arbor Science, 1977, pp. 431-459. 70. KELETI, G. ET AL Composition and biological properties of lipopolysaccharides isolated from Schizothrix calcicola (ag.) Gomont (Cyanobacteria). Appiled and environmental microbiology, 38: 471-477 (1979). 71. LIPPY, E. C. & ERB, J. Gastrointestinal illness at Sewickley, Pennsylvania. Journal of the Amencan Water Works Association, 68: 606-610 (1976). 72. KAY, G. P. ET AL Algal concentration as a quality parameter of finished drinking waters in and around Pittsburg, Pennsylvania. Journal of the American Water Works Assoczatzon, 72: 170-176 ( 1980). 73. MILLER, S. Drinking water and its treatment (An interview with Professor Sontheimer, University of Karlsruhe). Environmental science and technology, 14: 510-514 (1980). 74. HOEHN, R. C. ET AL. Algae as sources of tnhalomethane precursors. Water and sewage works, 126: 66-67 (1979). 75. BRILEY, K. F. ET AL. Trihalomethane production from algal precursors. Paper presented at the Third Conference on Water Chlorination: Environmental Impact and Health Effects, Colorado Springs, 31 October-2 November 1979, Ann Arbor, Ann Arbor Science, 1980, Vol. 3. 76. TARDIFF, R. G. Health effects of organics: R1sk and hazard assessment of ingested chloroform. Journal of the American Water Works Associatwn, 69: 658-661 (1977). 77. CHANG, S. L. ET AL Survival and protection agamst chlorination of human enteric pathogens in free-living nematodes isolated from water supplies. American journal of tropical medicine and hygzene, 9: 136-142 (1960). 78. PALMER, C. M. Algae in water supplies. Cincinnati, US Department of Health, Education and Welfare, 1959. 79. Cox, C. R. Operation and control of water treatment processes. Geneva, World Health Organization, 1964. 80. COLLINGWOOD, R. W. The effect of algal growth on the quality of treated water. In: Proceedings of a Symposium on Biological Indicators of Water Quality, Newcastle, October 1-15, 1978, University of Newcastle, 1978. 81. FOGG, G. E. Extracellular products of algae in freshwater. Archiv fur Hydrobwlogie, Beiheft. Ergebnisse der Limnologie, 5: 1-25 (1971). 82. PALMER, C. M. Algae and other interference organisms in New England water supplies. Journal of the New England Water Works Association, 72: 27-46 (1958). 83. DEGREMONT. Memento technique de /'eau, 8th ed. Rueil-Malmaison, Degremont, 1978. 84. WEBER, C. I. Bio/ogzca/ field and laboratory methods for measuring the quality of
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surface waters and wastes. Cincinnati, US Environmental Protection Agency, 1973. 85. WEBER, C. I. Recent developments in the measurement of the response of plankton and periphyton to changes in their environment. In: Glass, G., ed. Bioassay techniques and envrronmental chemistry, Ann Arbor, Ann Arbor Publishers, 1973, pp. 119-138. 86. WELCH, P. S. Limnological methods. New York, McGraw-Hill, 1948. 87. Determinacoes biolbgicas. sao Paulo, Companhia de Tecnologia de Saneamento Ambiental, 1978. 88. BRANCO, S. M. Hidrobiologia aplicada a engenharia sanitaria. 2nd ed., Sao Paulo, Companhia de Tecnologia de Saneamento Ambiental, 1978. 89. International standards for drinking-water, 3rd ed. Geneva, World Health Organization, 1971. 90. WEBER, C. I. & MooRE, D. R. Phytoplankton, seston and dissolved organic carbon in the Little Miami River at Cincinnati, Ohio. Limnology and oceanography, 12: 311-318 (1967). 91. JoRGENSEN, J. H. ET AL. Comparison of Ltmulus assay, standard plate count, and total coliform count for microbiological assessment of renovated wastewater. Applied and envrronmental microbiology, 37: 928-931 (1979).
PART III. HEALTH-RELATED CONSTITUENTS
INORGANIC
1. ARSENIC 1.1 General description 1.1.1 Sources
Arsenic occurs naturally in all environmental media and is usually present in the form of compounds with sulfur and with many metals (copper, cobalt. lead, zinc. etc.) (J). The average concentration in the earth's crust is about 2mg/kg (2). Although arsenic exists in various valency states and in both organic and inorganic forms, the levels of environmental arsenic are normally reported in terms of total arsenic (3). In some localized geographic areas, commercial use and production of arsenic compounds have resulted in significant elevation in the amounts of environmental arsenic above natural background levels. 1.1.2 Occurrence in water
Many arsenic compounds are water-soluble and, thus. contamination of water can occur. The chemical form of arsenic in water has not been fully elucidated, but both tri- and pentavalent forms have been identified; some forms of organic arsenic have been found in water (1 ). Geothermal discharges in New Zealand have been found to contain significant quantities of arsenic. In spite of its ubiquitousness in nature. most of the arsenic found in water derives from industrial discharges; the highest concentrations other than those occurring naturally m spring-waters are usually in areas of high industrial activity (4). 1.2 Routes of exposure 1.2.1 Drinking-water
A large number of water supplies contain very low levels of arsenic, i.e., well below 10 Jlg/litre (1-3). In special situations, gross contamination of well supplies has occurred and several thousand micrograms of arsenic per litre of water have resulted (5). There is little information concerning the form or species of arsenic in water supplies. 1.2.2 Food
Arsenic is present in most foodstuffs at concentrations generally below I mg/kg of dry weight (1). Marine fish can contain higher levels than 63
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this, and shellfish may contain levels well over 50 mg/kg (2. 5); the arsenic in seafood is mainly in an organic form (1 ). The average dietary sources of arsenic seem to give rise to an intake of about 30 pgjkg of body weight per day (2, 6, 7). Somewhat higher estimates have been quoted for certain countries (3, 8).
1.2.3 Air
The air in non-urban and non-industrial areas contains very low levels of arsenic, i.e., generally less than O.Olpg/m 3 (1). In urban areas and particularly industrial towns, levels of arsenic have exceeded lpg/m 3 ( l). A typical airborne exposure might be about 0.2 pgjm 3 of air, mainly as inorganic arsenic.
1.2.4 Other routes of exposure
1.2.4.1 Industrial exposure Industrial exposure to fumes containing arsenic compounds does ari:;e, especially in smelting operations; levels of arsenic exceeding I mg/m 3 of air have been recorded (1). Levels of arsenic as high as 380mg/kg have been reported in soil near smelters (9). 1.2.4.2 Tobacco Tobacco contains some arsenic, although, as a result of the decreasing use of arsenical sprays, the levels nowadays are low (1. 8 ). 1.2.4.3 Miscellaneous routes Certain pharmaceutical products contain arsenic and, for soi1''e individuals, the exposure to arsenic via these sources can be high. For the general population, the contribution from such routes of exposure ro arsenic is negligible in comparison with those from food, water, and air.
1.2.5 Relative significance of different routes of exposure
Based simply on intake of total arsenic, the contribution from the normally low levels in water in relation to general overall exposure is relatively small; however, at an arsenic level of 0.05 mg/litre in drinkingwater, this may equal or even exceed the total dietary intake in a nonoccupationally exposed individual. It is not possible to make estimates of the amounts of arsenic taken up by the body via various routes because very little is known about the particular forms of arsenic in food and water; this governs the proportions absorbed (2, 6).
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ARSENIC
65
1.3 Metabolism The form of arsenic affects its absorption. Elemental arsenic is very poorly absorbed. Some trivalent and pentavalent inorganic arsenic compounds are readily absorbed (2); organic arsenic, too, is generally well absorbed ( 6 ). Precise figures for the fractions absorbed of the different forms of arsenic do not appear to have been reported. Following exposure to arsenic, it enters the blood (8) and subsequently it is found mainly in the liver, muscles, kidneys, spleen, and skin (1); smaller quantities are also found in the brain, heart, uterus, thyroid, and pancreas, as well as in the hair and nails ( 10). Arsenic is transferred across the placenta (1). The biological half-life of arsenic appears to be between about ~en hours and a few days (1). There are no data indicating that arsenic accumulates with age. Human subjects are able to transform inorganic arsenic into monomethyl and dimethyl compounds (1); the overall mechanism of biotransformation is little understood, however (5). The excretion of arsenic compounds is mainly via urine (1, 2, 5). Trivalent arsenic may inhibit the activity of many enzymes by reacting with sulfydryl groups; reactions involving such groups are considered to be responsible for the toxic action of arsenic compounds (5).
1.4 Health effects There is no firm evidence that arsenic in any form is essential to man, although it is known that some organic arsenic compounds are beneficial as a growth stimulant for animals (2). The toxicity of arsenic compounds depends on the chemical and physical form of the compound, the route by which it enters the body, the dose and duration of exposure, dietary levels of interacting elements, and the age and sex of the exposed individual (2). Inorganic arsenic is more toxic than organic arsenic; trivalent inorganic arsenic is more hazardous than the pentavalent form. It is recommended that, when water is found to contain arsenic at levels of 0.05 mg/litre, an attempt should be made to ascertain the valency and chemical forms of the element. Acute poisoning by arsenic involves the central nervous system, leading to coma and, for doses of 70- 180 mg, to death ( 11). The gastrointestinal tract, nervous system, the respiratory tract, and the skin can be severely affected (2). Chronic poisoning is manifested by general muscular weakness, loss of appetite and nausea, leading to inflammation of the mucous membranes in the eye, nose, and larynx; skin lesions may also occur. Neurological manifestations and even malignant tumours in vital organs may also be observed (2). Poisoning may appear with doses as low as 3-6 mg/day over extended periods ( 12). One fatality was reported where for two and a half years well-water containing up to 7.6 mg/litre was drunk (13). Even as little as 0.6 mg of arsenic per litre of water might have been responsible for some infant deaths in Chile
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(14-16), although there is some uncertainty about this incident (2).
Although there is no evidence of adverse health effects with a concentration of 0.05 mg/litre, absorbed amounts of arsenic can be detected in the hair of people drinking water at this concentration (2). In China (Province of Taiwan) (17, 18), skin cancer in some villages was found to be associated with drinking well-water containing average arsenic concentrations of about 0.5 mg/litre~ there is some doubt about the actual levels of arsenic, however (2). Evaluations of the risk of getting skin cancer have been made by a WHO Task Group and the US Environmental Protection Agency (19). a Applying a linear nonthreshold model, a WHO Task Group (19) estimated that a lifetime exposure to arsenic in drinking-water at a concentration of 0.2 mg/litre gave a 5 ~-;, risk of getting cancer of the skin (this model assumes that the metabolism of arsenic is the same at low as at high exposures). Various manifestations, such as hyperpigmentation, keratoses, and lung cancer, have been observed where high occupational exposure to arsenic has occured (9). REFERENCES I. COMMISSION OF THE EuROPEAN COMMUNITIES. Trace metals: exposure and health effects. Oxford, Pergamon Press, 1979. 2. Guidelmes for Canadian drinking water quality, 1978. Quebec, Mm1stry of Supply and Services, 1979 (Supporting documentation). 3. Quality cnteria for water. Washington, DC, US Environmental Protection Agency, 1976. 4. Some metals and meta/he compounds. Lyon, International Agency for Research on Cancer, 1980 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 23). 5. NATIONAL RESEARCH COUNCIL. Drinkmg water and health. National Academy of Sciences, Washington, DC, 1977. 6. UNDERWOOD, E. J. Trace elements in human and animal nutrition. New York, Academic Press, 1977. 7. WHO Technical Report Senes, No. 532, 1973 (Trace elements m human nutntion), pp. 49-50. 8. Toxicology of metals, 1'01. II. Washington, DC, US Environmental Protection Agency, 1977 (Environmental Health Effects Research Series). 9. NATIONAL RESEARCH COUNCIL. Arsenic. Medical and biological effects of environmental pollutants. Washington, DC, National Academy Press, 1977. 10. MEALEY, J. ET AL. Radioarsenic in plasma, urine, normal tissues, and intracranial neoplasms. Archwes of neurology and psychwtry, 81: 310 (1959). II. VALLEE, B. L. ET AL. Arsenic toxicology and bwchemistry. A.M.A. archives of industrial health, 21: 132 (1960). 12. LISELLA, F. S. ET AL. Health aspects of arsenicals m the environment. Journal of environmental health, 30: 157 (1972). 13. WYLLIE, 1. InvestigatiOn of source of arsenic in well water. Canadian journal of public health, 28: 128 (1937). 14. RoSENBERG, H. G. Systemic arterial disease with myocardial infarction. Report on two infants. Circulation, 47: 270 (1973).
a
GREATHOUSE, D. G. Maximum acceptable limit for arsenic in drinking-water. A criteria document prepared for the World Health Organization, 1980, unpublished.
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67
15. ROSENBERG, H. G. Systemic arterial disease and chronic arsenicism in infants. Archives of pathology, 97: 360 (1974). 16. ZALDIVAR, R. Arsenic contamination of drinking water and foodstuffs causing endemic chronic poisoning. Beitriige zur Pathologie, 151: 384 (1974). 17. TSENG, W. P. ET AL. Prevalence of skin cancer in an endemic area of chronic arsenic1sm in Taiwan. Journal of the Natwnal Cancer Institute, 40: 453 (1968). 18. YEH, S. Skin cancer in chronic arsenicism. Human pathology, 4: 469 (1973). 19. Arsenic. Geneva, World Health Organization, 1981 (Environmental Health Cnteria 18).
2. ASBESTOS 2.1 General description 2.1.1 Sources Asbestos is a general term for fibrous silicate minerals of the serpentine or amphibole mineral groups. Six minerals have been characterized as asbestos: chrysotile, crocidolite, anthophyllite, tremolite, actinolite, and amosite. Chrysotile is a member of the serpentine group; the others belong to the amphibole group. These various forms of asbestos are composed of 40-60% silica, as well as oxides of iron, magnesium, and other metals. Asbestos is introduced into natural waters by the dissolution of asbestos-containing minerals and ores and from industrial effluents. There is some indication that atmospheric pollution may also contribute to the asbestos content of natural waters (1). Sedimentation, resuspension, migration, and chemical reactions affect the movement, abundance, and fate of asbestos fibres in water. The length of time from introduction of asbestos fibres into water until their disappearance is unknown (2). The use of asbestos-cement (which contains 170 g of asbestos per kg80% chrysotile and 20% crocidolite) for pipes in distribution systems could contribute to the asbestos content of drinking-water.
2.1.2 Occurrence in water Asbestos is commonly found in domestic water supplies. Typical background levels in rivers and lakes are considered to be about I million fibres per litre (3), although reported values range from less than I million to 10 million fibres per litre. Levels vary considerably, depending upon proximity to industrial sources. The asbestos content of untreated water from the Ottawa River has been reported to be 9.5 million fibres per litre (1). Generally, ordinary sand filtration removes about 90 ~~ of the individual asbestos fibres from water supplies (2). The most effective method for removal involves chemical coagulation with iron salts and polyelectrolytes followed by filtration. Based on surveys of asbestos concentrations in the drinking-water supplies in Canada, it was found that about 5% of the Canadian public consume water with fibre concentrations exceeding 10 million fibres per litre and about 0.6% 68
2.
ASBESTOS
69
consume water with concentrations exceeding 100 million fibres per litre. Levels ranged up to 2000 million fibres per litre in some asbestos-mining communities (4). 2.2 Routes of exposure 2.2.1 Drinking-water Since reported levels for the asbestos content of drinking-water supplies vary tremendously and since wide margins of error are associated with fibre-to-mass conversions, it is difficult to make an accurate estimate of the average daily intake from this source. Based on the results of a Canadian national survey and the assumption that the daily water consumption is 2 litres, the intake of asbestos from drinkingwater for 95 'lo of the population is probably less than 0.0001 mgjday. 2.2.2 Food The extent of asbestos contamination of solid foodstuffs has not been fully researched, owing to the lack of a practical, reliable analytical method. Foods that contain soil particles, dust, or dirt almost certainly contain asbestos fibres. Foodstuffs can also derive asbestos from water or impure talc used in their preparation (talc may be used as a dusting powder in chewing gum, on coated rice, and as an antisticking agent for moulded foods) (5). Asbestos may also be introduced into foods from impure mineral silicates, such as talc, soapstone, or pyrophyllite, which are used as carriers for spray pesticides (6). Asbestos fibres make excellent filter materials, and at one time asbestos was widely used in the food industry for clarification of beverages and other liquids. In some cases, the use of asbestos filters increases the concentration of asbestos in the finished product. Concentrations of 0.151 million fibres per litre have been found in some English beers (7). For Canadian beers, concentrations of 4.3-6.6 million fibres per litre have been recorded; in soft drinks, levels were between I. 7 million and 12.2 million fibres per litre {1). However, use of asbestos filters in the food industry has declined. 2.2.3 Air Asbestos is present in the air as a result of natural processes, such as weathering of rocks, or from industrial emissions. In the USA, typical levels of asbestos in air range from I x Io-s to 10 x Io-s mgjm 3 (8, 9). It has been estimated that I x I0- 6 mg of chrysotile asbestos could contain up to I million fibrils (8). If the concentration of asbestos in air is assumed to be less than 3 x Io-s mgjm 3 (concentrations recorded in urban areas in the USA are usually below this level (10)) and the daily respiratory volume to be
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20m 3 , then the daily intake of asbestos from air would be less than 0.0006 mg. 2.2.4 Relative significance of different routes of exposure It is apparent then that exposure to asbestos varies considerably, depending on proximity to industrial and natural sources. In general, however, intake from air is greater than intake from water. It should be noted, however, that fibre size distribution in air may be quite different from that in water. There are few quantitative data available on levels of asbestos present in food.
2.3 Absorption and distribution 2.3.1 Ingestion
Asbestos fibres are ingested with food, beverages, and drinking-water, and a significant proportion of inhaled asbestos cleared from the airways by mucociliary action is subsequently swallowed (1 1). The fate of ingested asbestos fibres is the subject of some controversy; a number of authors have concluded that asbestos fibres cross the walls of the gastrointestinal tract and penetrate into other tissues (1 2-15), but others claim that there is no clear evidence that this occurs (11, 16, 17). When a suspension of chrysotile fibres was injected directly into the stomachs of anaesthetized rats, it appeared that asbestos fibres were capable of penetrating the gut wall and migrating to the blood, spleen, omentum, brain, and other tissues (13). There was no evidence of lesions or transmigration of fibres across the gut wall in a well-controlled study in which rats consumed a diet containing 50 g of chrysotile per kg of feed for 21 months (1 1). Further work is required to determine the amount of asbestos fibres that cross the walls of the human gastrointestinal tract under normal conditions and if the number of absorbed fibres is sufficient to cause adverse local or systemic effects (1 2). 2.3.2 Inhalation
Evidence from several sources, including animal experiments, pathological observations, and physical studies, indicates that asbestos can be deposited in airways of the respiratory tract by sedimentation or by interception (18). Deposition by sedimentation is determined mainly by the diameter of the fibre; deposition by interception is determined by fibre length. After inhalation of small amounts of asbestos, short fibres are ingested by macrophages and then removed by mucociliary mechanisms; longer fibres are usually coated and subsequently fragment, disappearing within 18 months (16). Others are coated and remain as asbestos or ferruginous bodies. The clearance of asbestos from the lungs depends
2.
ASBESTOS
71
therefore on the fibres involved (shape, size, distribution, and chemical type), the presence of any factor that might affect the activity of the alveolar macrophages, and the level of exposure (16). Animal experiments have shown that inhaled asbestos fibres can move from the lung or trachea to other tissues (16). Fibres have been found in lymph nodes of guinea pigs exposed to crocidolite. The fibres presumably pass through the lymphatic system. Asbestos bodies have also been observed in the thyroid of guinea-pigs that had inhaled anthophyllite. In tissues of human mesothelioma patients, asbestos has been found in the lung, lymph nodes, and peritoneum; asbestos bodies have also been identified in the spleen and small bowel (19). Few data on levels of asbestos in tissues are available, and there is a need to determine the number of fibres that must accumulate in various tissues before disease develops (20). In one study, there were few asbestos fibres found in the lungs of normal patients; an intermediate number were present in asbestos-exposed patients with mesothelioma, and many thousands of such fibres were identified in the lung tissue of severely asbestotic patients (21). 2.4 Health effects 2.4.1 Ingestion 2.4.1.1. Animal studies The presence of asbestos fibres in many sites in the colonic epithelium and lamina propria has been noted in rats fed chrysotile asbestos (1 2). Groups of 32 Wistar SPF rats were fed Italian talc or Canadian chrysotile at the rate of 100 mgjday in malted milk powder on 5 days a week for 100 days over a six-month period; 16 controls were fed only malted milk. One gastric leiomyosarcoma was observed in each of the two groups fed chrysotile and talc, but none in the controls (22). In a study in which rats consumed filter paper containing chrysotile (approximately 25 mgjkg of body weight) daily for life, malignancies appeared in 8-14 months (23). 2.4.1.2 Epidemiological studies Although there have been a number of epidemiological studies conducted to date, the results of only one have shown a marginally significant association between asbestos levels in drinking-water and cancers of the digestive tract (24). Cancer incidence data for 721 census tracts of 5 Bay Area counties in California were examined in relation to chrysotile fibre concentrations in drinking-water, which were reported to range from not detectable to 36 million fibres per litre. The proportion of stomach cancers in males attributed to asbestos exposure was 10 '/~, based on the comparison of incidence data in low-exposure areas to
72
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HEALTH-RELATED INORGANIC CONSTITUENTS
incidence data in high-exposure areas, with due allowance for socioeconomic factors. These findings need to be confirmed. A study of 22 municipalities in Quebec did not reveal excess cancer mortality that could be related to the presence of asbestos in drinkingwater (25). The areas studied included Thetford Mines and Asbestos, where asbestos mining operations have been established for almost a century. Epidemiologists in the USA are currently investigating gastrointestinal cancer mortality data for Duluth, Minnesota, where asbestos concentrations of up to 100 million fibres per litre have been present in the public water supply since 1955; no evidence to suggest an increase in deaths due to gastrointestinal cancers has yet been found (26-28). The latency period for development of cancer from occupational inhalation of asbestos is 20-40 years, and therefore more complete data are required before any firm conclusions can be drawn from the Duluth studies. No association has been found in a study conducted in Pensacola, Florida, where levels up to 38 million fibres per litre have been recorded (24). Similarly, preliminary results of a study conducted in the Puget Sound region of Washington State have demonstrated no association between asbestos concentrations in drinking-water and cancer of the colon, stomach, kidney, or total alimentary tract. A more precise case-control study in this area is still under way (29). As mentioned previously, the use of asbestos-cement pipes in distribution systems is a potential source of asbestos contamination of drinking-water. Generally, it has been concluded that the concentrations of asbestos in drinking-water resulting from the use of asbestos-cement pipes do not present a hazard to human health (30-32). A study of cancer incidence in Connecticut over a 35-year period failed to show a relationship between the use of asbestos-cement pipes and the incidence of gastrointestinal cancer (33). A further study on asbestos-cement pipes and drinking-water in relation to cancer incidence was published by Meigs et a!. (34), also with apparently negative results. However, the hypothesis that ingested asbestos fibres cause cancer cannot be ruled out at the present time. 2.4.2 Inhalation
Occupational exposure to airborne asbestos has resulted in pulmonary fibrosis (asbestosis), pleural calcification, bronchogenic carcinoma of the lung, malignant mesothelioma of the pleura and peritoneum, and cancer of the gastrointestinal tract. Although data are inconclusive, it has also been suggested that there is a relationship between occupational inhalation of asbestos and malignancies of the ovary and larynx (3538). Asbestosis has been described as a "chronic inflammatory reaction in the terminal bronchioles and alveoli of the lung, with considerable fibrosis, leading to distortion and eventual obliteration of the alveoli" (39). The latency period ranges from 7 to 20 years. Generally, it is
2.
ASBESTOS
73
believed that all types of asbestos can cause asbestosis (40, 41) and that fibre-type is of less importance than in the etiology of asbestos-induced cancer (41, 42). A significant relationship between occupational exposure to airborne asbestos and incidence of bronchial carcinoma has been repeatedly noted (43-47). From 14% to 50% of patients with asbestosis die from bronchial carcinoma (40, 41). Evidence strongly suggests a synergistic relationship between cigarette smoking and asbestos in the causation of bronchial carcinoma (40, 46, 48, 49). An increasing incidence of asbestos-induced malignant mesotheliomas of the pleura and peritoneum has been noted in the Federal Republic of Germany, Great Britain, South Africa, the USA, and elsewhere (40). Few mesotheliomas have been identified in Canada (43). The latency period between first asbestos exposure and appearance of the tumour is from 20 to 40 years (50). There appears to be a relationship between prolonged inhalation of asbestos and the incidence of gastrointestinal cancer (47, 51). It has also been reported that gastrointestinal cancers account for one third of all malignant neoplasms in workers in Quebec chrysotile mines and mills and that the proportion is higher in those exposed to more than 400 million particles per cubic foot (14 000 million particlesjm 3 ) (41, 48). In a review of the subject (52), only one paper did not support the conclusion that occupational exposure to inhaled asbestos leads to an increased incidence of cancer of the digestive system.
REFERENCES I. CUNNINGHAM, H. M. & PONTEFRACT, R. D. Asbestos fibres m beverages and dnnkmg water. Nature, 232: 332 (1971). 2. Asbestos in the Great Lakes with emphasts on Lake Superior. A report to the International Joint CommissiOn, Great Lakes Research Adv1sory Board, 1974. 3. KRAMER, J. R. & MuRDOCH, 0. Asbestos research at McMaster University. Canadwn research and development, Nov.-Dec.: 31 (1974) 4. A natwnal survey for asbestos fibres in Canadian drinkmg water supplies. Ottawa, Department of National Health and Welfare, 1979 (Environmental Health Directorate Publication 79-EMD-34). 5. EISENBERG, W. V Inorganic particle content of food and drugs. Enl'ironmental health perspectit·es, 9: 183 (1974). 6. KAY, K. Inorganic particles of agncultural origm. Ennronmental health perspectfl•es, 9: 193 (1974). 7. BILES, B. & EMERSON, ·T. R. Exammatwn of fibres m beer. Nature, 219: 93 (1968). 8. SELIKOFF. I. J. ET AL. Asbestos air pollutiOn. Archives of em•tronmental health, 25: I (1972). 9. NICHOLSON, W. J. & PuNDSACK, F. L. Asbestos in the environment. In: Biological effects of asbestos, Lyon, International Agency for Research on Cancer, 1972 (IARC Scientific Publicatwns, No. 8). 10. BRUCKMAN, L. & RUBINO, R. A. Asbestos: ratwnale behmd a proposed a1r quality standard. Journal of the Air Pollution Control Assocwtwn, 25: 1207 (1975). II. GRoss. P. ET AL. Ingested mineral fibers: do they penetrate tissue or cause cancer? Archit'es of environmental health, 29: 341 (1974).
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12. LEE, D. H. K. Biological effects of ingested asbestos: report and commentary. Environmental health perspectives, 9: 113 (! 974). 13. CuNNINGHAM, H. M. & PoNTEFRACT, R. D Penetration of asbestos through the digestive tract of rats. Nature, 243: 352 (1973). 14. WESTLAKE, G. E. Asbestos fibers in the colome wall. Em•ironmental health perspectil'es, 9: 227 (1974). 15. AMACHER, D. E. ET AL. Effects of ingested chrysoule on DNA synthesis in the gastrointestinal tract and liver in the rat. Environmental health perspectit·es, 9: 319 (1974). 16. HoLT, P. F. Small animals m the study of the pathological effects of asbestos. Environmental health perspecllves, 9: 205 (1974). 17. BoLTON, R. E. & DAVID, J. M.G. The short-term effects of chronic asbestos ingestion m rats. Annals of occupatwnal hygiene, 19: 121 (1976). 18. T!MBRELL, V. The inhalation of fibrous dusts. Annals of the New York Academy of Sciences, 132: 255 (1965). 19. GODWIN, M. C. & JAGATIC. J. Asbestos and mesotheliomas. Environmental research, 3: 391 (1970). 20. BROWN, A. Lymphohematogenous spread of asbestos. Environmental health perspectives, 9: 203 (1974). 21. FoNDIMARE, A. & DESBORDES, J. Asbestos bodies and fibers in lung tissues. Em•1ronmental health perspectives, 9: 147 (1974) 22. WAGNER, J. C. ET AL. Animal experiments with talc. In: Inhaled particles and vapors, IV, New York, Pergamon Press, 1977. 23. GIBEL, W. ET AL. Animal experimental mvestigatwns of the carcinogenic activity of asbestos filter material followmg oral admmistration. Archit• fur Geschwulstforschung, 46: 437 (1976). 24. McCABE, L. J. & MILLETIE, J. R. Health effects and prevalence of asbestos fibers in drinking water. Proceedings of the American Water Works Association Annual Conference, San Francisco, 24-29 June I979, Denver, CO, AWWA. 25. WIGLE, D. T. Cancer mortality m relatwn to asbestos m munictpal water supplies. Archives of environmental health, 32: 185 (1977). 26. KAY, G. H. Asbestos in drinking water. Journal of the Amencan Water Works Associatwn, 66: 513 (1974). 27. LEVY. B. S. ET AL Investigatmg possible effects of asbestos m city water: surveillance of gastrointestinal cancer in Duluth, Minnesota. American journal of epidemiology, 103: 362 (1976). 28. MASSON, T. J. ET AL. Asbestos-like fibers in Duluth water supply: relatwn to cancer mortality. Journal of the Amencan Med1cal Associatwn, 228: 1019 (1974). 29. SEVERSON, R. K. A study of the effects of asbestos in drinking water on cancer mcidence in the Puget Sound regwn. Seattle, University of Washington, 1979 (M.Sc. thesis). 30. The American Water Works Research Foundation. A study of the problem of asbestos m water. Journal of the American Water Works AssoCiation, 66: I (1974). 31. ELZENGA, C. H. J ET AL. A prehmmary investigation mto the appearance of asbestos in Dutch drinking water. Netherlands, 1972 (report of Drmking Water Group of the TNO Health Organization Support Committee). 32. OLSON, H. L. Asbestos in potable water supplies, Journal of the American Water Works Assocwtion, 66: 515 (1974). 33. HARRINGTON, J. M. ET AL. An investigation of the use of asbestos cement pipe for public water supply and the incidence of gastromtestmal cancer in Connecticut, 19351973. American ;ournal of epidemiology, 107: 96 (1978). 34. MEIGS, J. W. ET AL. Asbestos cement pipe and cancer m Connecticut 1955-1974. Journal of environmental health, 42: 187-191 (1980). 35. STEEL, P. M. & McGILL, T. Asbestos and laryngeal carcinoma. Lancet, 2: 416 (1973). 36. SHETIIGARA, P. T. & MORGAN, R. W. Asbestos, smoking and laryngeal carcinoma. Archives of environmental health, 30: 517 (1977). 37. MoRGAN, R. W. & SHETIIGARA, P. T. Occupational asbestos exposure, smoking and laryngeal carcinoma. Annals of the New York Academy of Sciences, 271: 308 (1976).
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38. GRAHAM, H. & GRAHAM, R. Ovanan cancer and asbestos. Environmental research, 1: 115 (1967). 39. DAVIES, P. ET AL. Asbestos induced selective release of lysosomal enzymes from mononuclear phagocytes. Nature, 251: 423 (1974). 40. PARKES, W. R. Occupational lung disorders. London, Butterworths, 1974. 41. HARINGTON, J. S. The biological effects of mineral fibres, especially asbestos, as seen from in vitro and in vivo experiments. Annales d'anatomie pathologique, 21: 155 (1976). 42. ANDERSON, H. A. ET AL. Asbestos disease resulting from household exposure to occupatwnal dusts. Chest, 66: 318 (1974). 43. McDoNALD, J. C. Cancer in chrysotile mines and mills. In: Biologzcal effects of asbestos, Lyon, International Agency for Research on Cancer, 1972 (IARC Scientific PublicatiOns, No. 8). 44 WEBSTER. I. Malignancy in relation to crocidolite and amosite. In: Bwlogical effects of asbestos, Lyon, International Agency for Research on Cancer, 1972 (IARC Scientific Publications, No. 8). 45. MEURMAN, L. D. ET AL. Mortality and morbidity of employees of anthophyllite asbestos mines of Finland. In: Bwlogical effects of asbestos, Lyon, International Agency for Research on Cancer, 1972 (IARC Scientific Publications, No. 8). 46. NEWHOUSE, M. L. Cancer among workers in the asbestos textile industry. In: Biological effects of asbestos, Lyon, Internatwnal Agency for Research on Cancer, 1972 (IARC Scientific Publications, No. 8). 47. SELIKOFF, I. J. ET AL. Cancer risk of insulatiOn workers in the United States. In: Biological effects of asbestos, Lyon, International Agency for Research on Cancer, 1972 (IARC Scientific Publications, No. 8). 48. SELIKOFF, I. J. ET AL. Asbestos exposure, smoking and neoplasia. Journal of the American Medzcal Association, 204: 104 (1968). 49. SELIKOFF, I. J. Air pollution and asbestos carcinogenesis: investigation of possible synergism. In: Mohr, L. L. et a!., ed., Atr pollution and cancer in man, Lyon, International Agency for Research on Cancer, 1977 (IARC Scientific Publications, No. 16, p. 247). 50. Asbestos. Lyon, International Agency for Research on Cancer, 1977 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 14) pp. 62-65. 51. SELIKOFF, I. J. ET AL. Asbestos exposure and neoplasia. Journal of the American Medical Association, 188: 22 (1964). 52. SCHNEIDERMAN, M. A. Digestive system cancer among persons subjected to occupational inhalation of asbestos particles: A literature review with emphasis on dose response Envtronmental health perspectir•es, 9: 307 (1974).
3. BARIUM 3.1 General description 3.1.1 Sources
Barium is present in the earth's crust in a concentration of 0.5 g/kg and the mineral barytes, barium sulfate, is the commonest source. Barium carbonate (witherite) is another form, although less common. Traces of barium are present in most soils. Barium compounds are used in oil drilling, production of paints, the processing of diesel fuels, the manufacture of paper, rubber, linoleum and similar products, the manufacture of ceramic glazes and enamels, and in medical diagnostics. 3.1.2 Occurrence in water
Most waters contain some barium, but the concentration is generally well below 0.1 mg/litre (1, 2), although some underground sources may contain levels as high as 10 mg/litre in geothermal brines (3). The source is normally natural mineral matter; although barium sulfate, the predominant form, is only slightly soluble in water, in the presence of certain common anions the solubility of barium can be markedly enhanced. The chemical form of barium in tap-water is not definitely known (4, 5). 3.2 Routes of exposure 3.2.1 Drinking-water
Relatively little is known about levels of barium in tap-water throughout the world. However, a survey of water quality in 100 cities in the USA found levels of barium in the range 0.002-0.38 mg/litre, with a median level of 0.043 mg/litre (6) and, in a survey of 2595 water samples, less than 0.1% contained barium at a level exceeding 1 mgjlitre (4). In the USSR, the maximum allowable concentration for barium in drinking-water has been set at 0.1 mgjlitre (7). 3.2.2 Food
Barium is present in traces in many foodstuffs (J); brazil nuts are an especially rich source and contain up to several thousand JJ.g/g (1). There 76
3.
BARIUM
77
is relatively little published information on dietary intake, but some values ranging between less than 0.1 to nearly 2 mg/day have been published (8). Dietary intakes are estimated to be about 0.061.2 mg/day in the USA (9, 10) and 0.4-1.2 mg/day in Canada (5). 3.2.3 Air Little published information is available, but an average concentration of 0.005 ng/m 3 air has been reported for cities in the USA (1). 3.2.4 Other routes of exposure 3.2.4.1 Industrial exposure Barium can be detected in the working atmosphere of industrial environments, and a threshold limit (based on health effects) has been set at 0.5 mg/m 3 of air for soluble barium compounds in industrial situations in the USA (4). 3.2.4.2 Smoking Traces of barium can be detected in tobacco, but because little barium is inhaled during smoking, the exposure from this source is unimportant. 3.2.5 Relative significance of different routes of exposure Two simplified examples of exposure (where the levels in water are elevated) have been calculated to illustrate the relative intake. Assuming a daily water consumption of 2 litres with a barium concentration of 0.05 mg/1, the daily intake of barium from water would be 0.1 mg. With a daily air intake of 20 m 3 and ambient air levels of 0.005 ng of barium per m 3 , the daily exposure to barium via inhalation is 0.1 ng/day. This amount is negligible compared with barium intake via food or water. Assuming a typical daily food intake of 0.5-1.2 mg of barium, the total daily intake from all sources is 0.6-1.3 mg. No reliable information is available for the exposure of children to barium. As the levels of barium in drinking-water are unlikely to exceed 0.05 mg/litre. the intake in relation to normal total intake is unlikely to be more than about 15%. However, where the barium levels in water are high, i.e., over 1 mg/litre, more than three-quarters of the total barium intake could result from drinking-water. The uptake of barium is difficult to quantify accurately. Barium in food is poorly absorbed (8), but since a high proportion could be absorbed if ingested as a soluble salt in water, then the contribution made by drinking-water to the absorption of the element could be especially large.
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3.3 Metabolism There is no evidence that barium is essential for human nutnuon. Insoluble forms of barium, such as barium sulfate (used medicinally), are very poorly absorbed, and have a very low toxicity. Soluble barium salts are readily and rapidly absorbed (4); in this form, 50 °/~ or more of ingested barium is absorbed (1). Only a small fraction of barium in normal foodstuffs appears to be absorbed (8). Most absorbed barium is present in bone (8), but it has also been detected in a number of other tissues, such as kidney, liver, and heart (8). The metabolic pathways for barium are similar to those for calcium (4). Excretion of barium occurs more readily than with calcium; about a quarter of any ingested barium is excreted within 24 hours (4, 5). 3.4 Health effects
Animal experiments using drinking-water suggest that 5 mg of barium per litre, as barium acetate, would cause no toxic effects even after prolonged exposure (1 1). In rats, a total dose of 250 mg of barium over a period of 4-13 weeks also caused no adverse effects (12). Barium is acutely toxic when soluble salts are ingested in excess; if taken as the chloride, the fatal dose of barium for an adult is about 550-600 mg (4). In high doses, it induces a strong prolonged stimulant action on all muscles, including those of the heart and gastrointestinal tract (4). The acute toxic dose has been found by one research group to be 200-500 mg (1). In another study, 125 mg as a single dose has been suggested to be the threshold for acute toxic effects (1 3). In certain industries, it is claimed that a benign form of pneumoconiosis is associated with the inhalation of dusts containing barium (4). A recent epidemiological study indicated that, where levels of barium in water were up to 10 mg/litre, a statistical association was found with cardiovascular death rates, but because there were population changes in the communities studied, the finding must be treated with a certain degree of caution (3). Epidemics of barium poisoning have been reported in China (14). "Pa Ping" disease was attributed to the prolonged ingestion of table salt containing up to 250 g of barium chloride per kg (14).
REFERENCES I. Toxicology of metals, vol. II. Washington, DC, US Environmental Protection Agency, 1977 (Environmental Health Effects Research Series). 2. Quality criteria for water. Washington, DC, US Environmental Protection Agency, 1976. 3. BRENNIMAN, G. R. ET AL Cardiovascular disease death rates in communities with elevated levels of barium m drinking water. Environmental research, 20: 318 (1979). 4. NATIONAL RESEARCH COUNCIL. Drinking water and health. Washington, DC, National Academy of Sciences, 1977.
3.
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79
5. Guidelines for Canadian drinkmg water quality, 1978. Quebec, Ministry of Supply and Services, 1979 (Supporting documentation). 6. DURFOR, C. N. & BECKER, E. Selected data on public supplies of the 100 largest cities in the United States, 1962. Journal of the American Water Works Association, 56: 237 (1964). 7. KRASOVSKY, G. N. ET AL. [A proposed maximum allowable concentration for barium in water.] Gigiena i sanitarija, (6): 86 (1980) (in Russian). 8. UNDERWOOD, E. J. Trace elements in human and animal nutrition. New York, Academic Press, 1977. 9. GORMICA, A. Inorganic elements in food used in hospital menus. Journal of the American Dietetic Association, 56: 397 (1970). 10. SCHROEDER, H. A. ET AL. Trace metals in man: strontium and barium. Journal of chronic diseases, 25: 491 ( 1972). 11. SCHROEDER, H. A. & MITCHENER, M. Lifetime studies in rats: effects of aluminum, barium, beryllium and tungsten. Journal of nutrition, 105: 420 (1975). 12. Guidance for the issuance of variances and exemptions. Washington, DC, US Environmental Protection Agency, Office of Drinking Water, 1979. 13. BROWNING, E. Barium. In: Toxicity of industrial metals, London, Butterworths, 1969. 14. POLSON, C.]. & TATTERSALL, R. N. Barium. In: Clinical toxicology, London, Pitman Medical, 1969.
4. BERYLLIUM 4.1 General description 4.1.1 Sources
Beryllium is commonly found as part of feldspar mineral structures and may exist as the mineral beryl in small localized deposits (1). The amounts mined have been very small. The primary source of beryllium in the environment is the burning of fossil fuels, although contamination is normally slight (2). Because of its light weight and high tensile strength (3), the metal is valuable as a constituent of special alloys for applications such as space vehicles, X-ray windows, and certain electrical components. 4.1.2 Occurrence in water
Beryllium can enter waterways through the weathering of rocks, atmospheric fallout, and industrial and municipal discharges. However, levels in fresh water appear to be very low, generally below 1 .ug/litre (4). 4.2 Routes of exposure 4.2.1 Drinking-water
Few surveys have been carried out to determine levels of beryllium in drinking-water. Concentrations ranging from 0.01 to 1.2 .ug/litre, with a mean concentration of 0.2 ,ug/litre, were found in a survey in the USA (5). 4.2.2 Food
Relatively little information is available concerning the beryllium content of food. Data from the USA indicate that about 100 ,ug/day could be a typical dietary intake (6). A study in the United Kingdom estimated that the average dietary intake could be less than 15 ,ug/day (7). The beryllium content of various foodstuffs collected in New South Wales, Australia, ranged from 0.01-0.12 mg/kg (8). 4.2.3 Air
Data from the USA indicate that atmospheric beryllium is detected infrequently and usually in small amounts (4); levels generally are within the range 0.3-3 ngjm 3 of air (6). 80
4.
BERYLLIUM
81
4.2.4 Other routes of exposure 4.2.4.1 Industrial exposure Industrial processes in which exposure to beryllium can occur include the mining and extraction of beryllium, aerospace equipment, alloy machining, electroplating, and atomic energy industries. Exposure to beryllium in industrial settings is via inhalation and skin contact. Without ventilation, concentrations of beryllium in air as high as 23 J.,tg/m 3 have been reported (3). In the USA, the current occupational standard for exposure to beryllium is 2 J.,tg/m 3 of air (9).
4.2.5 Relative significance of different routes of exposure The contribution of beryllium in air to the total body burden is negligible compared with the contribution from food and water combined. The calculations below, therefore, consider the contributions made by food and water only. (a) Daily diet: 10 Jl9 of beryllium per day (adults) Weekly beryllium intake (119) Beryl hum concentration 1n water Water only Food only Total water contnbut1on (%)
Total
1 119/litre 2 119/litre
14 28
70 70
84 98
17 28
(b) Daily diet: 100 Jl9 of beryllium per day (adults) Weekly beryllium intake ( 119) Beryllium concentration in water Water only Food only Total water contribution (%)
Total
1 119/litre 2 119/litre
14 28
700 700
714 728
2 4
4.3 Metabolism Dermal absorption of beryllium is negligible, since beryllium becomes bound to certain epidermal constituents, but there have been reports of beryllium dermatitis and granulomatous ulcerations of the skin at sites where insoluble beryllium compounds were embedded; conjunctivitis has also been described (10). Compounds of beryllium are not readily absorbed when ingested, since they tend to form insoluble compounds at physiological pH levels (3). Less than 1 % of ingested beryllium is
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absorbed (3, 11). Inhaled soluble compounds of beryllium are absorbed from the lungs into the blood and are transported mainly as orthophosphate colloids, being ultimately deposited in the bone as insoluble hydroxides. Insoluble beryllium compounds tend to remain in the lungs indefinitely after inhalation (4, 6). Absorbed beryllium is either excreted in the urine or deposited in the kidneys or bone (3). Excretion of beryllium is fairly rapid (ll, 12). 4.4 Health effects
The inhalation of beryllium has been demonstrated to be harmful to human beings (1 1). Acute respiratory exposure can result in severe health effects, including rhinitis, pharyngitis, pneumonitis, and pulmonary oedema (1 1). Beryllium is very poorly absorbed from the gastrointestinal tract, and its toxicity via this route of entry is low; no reports of oral toxicity in humans have been found in the available literature. Inhalation, intratracheal instillation, and intravenous injection have been shown to induce tumours in experimental animals. Although firm evidence of human cancer induction by exposure to beryllium is lacking, the evidence is suggestive (4). Some studies of tumour induction in experimental animals as a result of ingestion of beryllium compounds at levels up to 500 mg/kg of diet have been inconclusive. In two studies, rats and mice ingesting beryllium in drinking-water at a concentration of 5 mgflitre for life did not show a statistically significant increase in tumours as compared with controls (13, 14). In another study, rats were fed beryllium in the diet at levels of 5, 50, and 500 mg/kg of feed. The authors concluded that there was no evidence of a carcinogenic response related to beryllium ingestion (1 5). Epidemiological studies conducted in the USA have not shown a significant correlation between ingestion of beryllium and human cancer. However, the International Agency for Research on Cancer (16) concluded: "There is sufficient evidence that beryllium metal and several beryllium compounds are carcinogenic to three experimental animal species ... Taken together, the experimental and human data indicate that beryllium should be considered suspect of being carcinogenic to humans." This conclusion is related essentially to inhalation of beryllium. The only standard limits for beryllium in water are those issued by the USSR, where a maximum allowable concentration of 0.2 Jig/litre has been set (1).
REFERENCES I. SASH!NA, L. A. [Experimental data to substantiate the maximum permissible concentration of beryllium in water bodies.] Gigiena i sanitarija, (2): I0 ( 1965) (in Russian).
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BERYLLIUM
83
2. TEPPER, L. B. Beryllium. In: Lee, D. H. K, ed. Metallic contaminants and human health, New York, Academic Press, 1972. 3. OAK RIDGE NATIONAL LABORATORY. Reviews of the environmental effects of pollutants VI. Beryl/tum. Cmcinnati, Health Effects Research Laboratory, 1978 (EPA-600/1-78028). 4. Ambient water quality criteria for beryllium. Washington, DC, US Environmental Protection Agency, 1980 (EPA-440/5-80-024). 5. KOPP, J. F. & KRONER, R. C. Trace metals in waters of the United States. Cincinnati, US Department of the Interior, 1967. 6. Toxicology of metals, vol. II. Washington, DC, US Environmental Protection Agency, 1977 (Environmental Health Effects Research Series). 7. HAMILTON, E. I. & MINSKY, M. J. Abundance of the chemical elements in man's diet and possible relations with environmental factors. Science of the total environment, 1: 375 (1973). 8. MEEHAN, W R. & SMYTHE, L. E. Occurrence of beryllium as a trace element m envuonmental matenals. Environmental science and technology, 1: 839 (1967). 9. TLVs-Threshold ltmtt values for chemical substances and physical agents m the workroom em•tronment with tntended changes for 1979. Cincinnati, Amencan Conference of Governmental Industrial Hygienists, 1979. 10. NICHIMURA, M. Clinical and experimental studies on acute beryllium disease. Nagoya ;ournal of medtcal science, 28: 17 (1966). II. NATIONAL RESEARCH COUNCIL. Drinking fVater and health. Washington, DC, National Academy of Sciences, 1977. 12. STOCKINGER, H. E. The toxicology of beryllium. Washington, DC, US Department of Health, Education and Welfare, 1972 (Publication 2173). 13. ScHROEDER, H. A. & MITCHELL, M. Lifetime studies in rats: effects of alummum, barium, beryllium and tungsten. Journal of nutritwn, 105: 420 ( 1975). 14. ScHROEDER, H. A. & MITCHELL, M. Lifetime effects of mercury, methylmercury and nine other trace metals on mice. Journal of nutrition, 105: 452 (1975). 15. MoRGAREIDGE, K. ET AL. Chronic feeding studies with beryllium sulphate in rats. Pittsburgh, Food and Drug Research Laboratories, Inc., 1975 (Final report to the Aluminum Company of America, 15219). 16. Some metals and metallic compounds. Lyon, International Agency for Research on Cancer, 1980 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 23), p. 190.
5. CADMIUM 5.1 General description
5.1.1 Sources
Cadmium-containing minerals are found in specific parts of the world, although the metal is uniformly distributed in trace amounts in the earth's crust. Practically all zinc ores contain small amounts of cadmium. Although rare, the predominant ore of cadmium is greenockite (cadmium blende), i.e., cadmium sulfide; this form is often associated with deposits of sphalerite (zinc sulfide). Cadmium production began slowly at the end of the nineteenth century. It is produced normally as a by-product of zinc extraction. The use of the element has been increasing steadily during this century, but it is only in the last 20 years that it has been of major interest. Cadmium has begun to contaminate the environment, and it has been found in air, food, soil, plants, and water. The principal uses of cadmium are in the fabrication of alloys and solders, metal plating, as pigments, as stabilizers in plastic materials, and in batteries.
5.1.2 Occurrence in water
The solubility of cadmium in water is influenced by the nature of the source of the cadmium and the acidity of the water. Surface-waters that contain more than a few micrograms of cadmium per litre have probably been contaminated by discharges of industrial wastes or by leaching from areas of landfill, or from soils to which sewage sludge has been added (J). Unpolluted waters generally contain less than 1 J.lg/litre (2-4).
The levels of cadmium in public water supplies are normally very low, since generally only tiny amounts exist in raw water and even where levels are somewhat elevated, many conventional water-treatment processes yvill remove much of the cadmium (5). Higher levels of cadmium in tap-water are associated with plated plumbing fittings, silver-base solders, and galvanized iron piping materials (1). 84
5.
CADMIUM
85
5.2 Routes of exposure 5.2.1 Drinking-water Drinking-water normally contains very low concentrations of cadmium, of the order of I Jlg/litre or less (1, 6-8); occasionally, levels up to 5 Jlg/litre have been reported (J) and on rare occasions levels up to 10 Jlg/litre have been detected (9). In some areas, well-water may contain elevated concentrations of cadmium (1). It is probable that the levels of cadmium could be higher in areas supplied with soft water of low pH, since this would tend to be more corrosive to any plumbing systems containing cadmium. The level of cadmium in a sample of water is likely to be a function of how long the water has been in contact with the plumbing, and as a consequence there is likely to be a variation in concentration when water is drawn at different times of day from the same tap. It would be difficult to define the average exposure unless a large number of samples were collected. Estimated daily exposure to cadmium via water, based on a water consumption of 2 litres per day, ranges from substantially less than }Jl g to over 10 Jlg per day. Such estimates are, of course, based on the presumption that all the cadmium from the water is ingested; it is likely, however, that in the preparation of certain beverages, e.g., tea, not all the cadmium will be consumed since some could become attached to the surfaces of the utensils, etc. 5.2.2 Food Most foodstuffs contain traces of cadmium; crops grown in polluted soil (from industrial contamination and from use of sewage sludge as fertilizer) or irrigated with polluted water may contain increased concentrations, as may meat from animals grazing on contaminated pastures. The kidneys and livers of animals concentrate cadmium, and people who eat these food items will tend to ingest more cadmium than those who do not (1). Shellfish also tend to accumulate the metal (1). An additional source of cadmium is phosphate fertilizer. A comprehensive review of the cadmium content of foodstuffs worldwide has been published (1). Most foodstuffs contain less than 0.1 mg/kg (wet weight). Typical dietary intakes range from 15 to 60 Jlg of cadmium per day (1, 10). The higher value is derived from a typical Japanese diet (9). Japan has the highest natural levels as well as the highest levels in polluted areas. Virtually nothing is known of the chemical form of cadmium in foodstuffs; such information would be valuable to ascertain the precise absorption of the metal when food is ingested. 5.2.3 Air The levels of cadmium in ambient air are generally low (11). Longterm average concentrations may vary from less than 0.001 to 0.5 Jlg/m 3
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depending on the degree of industrialization and presence of cadmiumemitting industries (10). A comprehensive review of the cadmium levels in the air worldwide has been published (1). It has been estimated that members of the general population will generally inhale less than 0.05 Jlg/day (10). For unusually polluted areas, maximum values as high as 3.5 Jlg/day have been estimated (1), and even in cities where the levels are 30 times as high as rural background values, this would still be small relative to the food source. The fraction of the particles containing cadmium that are deposited and retained in the lung varies with the particle size, but an average deposition of 25 % has been estimated (J) for the range of particles that exists in ambient air. 5.2.4 Smoking Cadmium is present in tobacco, one cigarette normally containing 1-2 Jlg. Because cadmium is volatile at elevated temperatures, some of the metal will be inhaled during smoking. Estimates have been made of the contribution made by various types of tobacco in different countries (1). Typically, 2-4 Jlg of cadmium per 20 cigarettes is inhaled; probably 50% will be deposited in the lungs (1). 5.2.5 Industrial exposure Workers in industry may inhale concentrations of cadmium ranging from a few micrograms up to several thousand micrograms per cubic metre of air (10); the highest levels, however, are associated mainly with exposures that occurred some years ago. 5.2.6 Relative significance of different routes of exposure From the information provided in sections 5.2.1-5.2.5, it can be seen that, for different individuals and population groups, there can be a wide range of exposures to cadmium from water, food, air, and occupational exposure, etc. For such exposures, a WHO Expert Committee (8) recommended that the weekly intake of cadmium should not exceed 0.5 mg per person. To give some idea of particular situations, a few simplified examples are given here. For low concentrations of cadmium in water (up to 1 Jlg/litre), the typical contribution made by water to overall intake for adults is less than 5 to 10% (based on dietary intakes of 15-60 Jlg/day and inhalation of 0.05 Jlg/day from ambient air). Since some information is available on the absorption of ingested and inhaled cadmium, it is relevant to consider the absorption of cadmium; some estimates for uptake in adults are provided in the table on page 87. No reliable information is available for children.
5.
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87
(a) Food intake 20 Jlg of Cd per day (6% absorption); inhalation 0.05 Jlgfday with 25% retention by the lungs and 64% absorption; drinking- water containing 1 Jlg of Cd per litre or 5 Jlg of Cd per litre (6% absorption) Weekly cadmium uptake (JLQ) Cadmium concentration in water Water only A1r only Food only Total water contribution (%)
Total
1 JLQ/Iitre 5 JLQ/Iitre
0.8 4.2
0.1 0.1
8.4 8.4
9.3 12.7
9 33
Smokmg 20 c1garettes per day would reduce the water contnbuuons to 5% and 21% respectively
(b) Food intake 50 Jlg of Cd per day (6% absorption); inhalation 0.05 Jlgfday with 25% retention by the lungs and 64% absorption; drinking- water containing 1 Jlg of Cd per litre or 5 Jlg of Cd per litre (6% absorption) Weekly cadmium uptake ( JLQ) Cadmium concentration in water Water only Air only Food only Total water contribution (%)
Total
1 JlQ/Iitre 5 JlQ/Iitre
0.8 4.2
0.1 0.1
21.0 21.0
21.9 25.3
4 17
Smok•ng 20 c1garettes per day would reduce the water contnbut1ons to 3% and 13% respect•vely
5.3 Metabolism Cadmium is fairly readily absorbed through ingestion or through the lungs. Alimentary absorption is affected by a number of factors, such as age, calcium, iron, zinc, and protein deficiency (I, 7, 10), and the chemical form of the cadmium ingested. As with lead (12), the state of the stomach is likely to influence the amount absorbed, with a fasting stomach probably providing the maximum uptake in contrast with a full stomach. In human subjects given labelled cadmium, between 4. 7% and 7% (mean 6 %) of orally administered cadmium was absorbed (I, 10). Dietary factors, such as iron, calcium, and protein deficiency, may increase the gastrointestinal absorption rate (13). In "iron-deficient women, up to 20% of ingested cadmium was found to be absorbed. Pulmonary absorption is dependent on the size and solubility of the particles containing cadmium and will be affected, too, by the depth and rate of breathing. Approximately 50% of 0.1-JLm particles will be retained by the lung in contrast to 20% for particles as large as 2 JLm (10). From information on the particle size distribution of cadmium in the general air and for tobacco smoke, it has been estimated that typically 25 % and 50% of the particles will be retained, respectively (10).
88
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Absorbed cadmium will enter the blood and become concentrated in certain parts of the human body (1, 14). Both the liver and the kidneys act as stores of cadmium (about 50% of any accumulated cadmium is found in these organs) (1). The cadmium is to a large extent bound to a protein of low relative molecular mass, known as metallothioneine (1 1); this metal-binding protein is believed to be involved in cadmium transport and absorption (1 5). Cadmium has a long biological half-time in the body (13-38 years) and accumulates with age (1). It has been found that, because the placenta acts as a fairly efficient barrier to cadmium, the newborn are virtually free of cadmium (approximately I Jlg only (1, 5)) in contrast to non-occupationally exposed people of age 50 who may have 10-50 mg (5) stored in their bodies. Grossly exposed industrial workers have been discovered to have levels of over 1000 mg (1). Blood levels are usually below 20 Jlg/litre for non-smoking members of the general population (1, 10). Cadmium in blood reflects recent exposure rather than body burden. Excretion of cadmium is usually rather slow, mainly via the urine (1). On a group basis, urinary cadmium is generally regarded as a good indication of the body burden. Cadmium interacts with other metals, especially zinc, and may influence the relative distribution of zinc in the body.
5.4 Health effects
Acute effects have been seen where food has been contaminated by cadmium from plated vessels; severe gastrointestinal upsets have been reported (1, 5). The acute oral lethal dose of cadmium for man has not been established, but it is estimated to be several hundred milligrams (16). Health effects have been demonstrated in industrial workers heavily exposed to cadmium oxide fumes and dust (10). Bronchitis, emphysema, anaemia, and renal stones have been reported (5). The renal cortex is generally accepted to be the critical organ for cadmium accumulation in man (8, 10, 17). The classical renal effects of cadmium poisoning are associated with proteinuria, glucosuria, and aminoaciduria (7). Where the exposure has been high, as was the case in Japan in the outbreak of Itai-Itai disease (a bone disease), irreversible renal injury occurred in those most severely exposed, especially in elderly women (1). There have been no reported effects from the low levels of cadmium that can be found in drinking-water, although contaminated beverages made from dispensers containing cadmium-plated fittings have caused acute effects in children (1). There have been many studies in which animals have been dosed with cadmium. There is evidence of hypertension after long-term low-level oral exposure and teratogenic, mutagenic and carcinogenic effects after injection of high doses (1). Short-term administration of cadmium in drinking-water at a level of 10 mg/litre has been found to cause partial
5.
CADMIUM
89
inhibition of the gastroi11testinal absorption of iron (1). Krasovsky's work indicates general toxic and gonadotoxic effects due to cadmium (18).
There are suggestions that there is a relationship between ingestion of cadmium and hypertension in man; however, at the present time, this relationship is inconclusive (19). The evidence that cadmium may be carcinogenic to man is rather weak (7, 20) although prolonged and heavy industrial exposures may constitute an increased risk of prostate cancer (21, 22). In the opinion of a WHO Study Group (15) the epidemiological studies on carcinogenicity of cadmium are not conclusive because most of them involved only a small number of workers. For this and other reasons the possible carcinogenicity of cadmium cannot be considered in deriving healthbased occupational exposure limits. The relationship between exposure level and cadmium concentration in blood is not yet sufficiently understood to derive a biological limit for blood with satisfactory precision. However, a WHO Study Group (15) agreed that the value of 10 ,ugjlitre of whole blood should be accepted as a tentative noadverse-effect level. It has been estimated that 3 mg of cadmium is the no-effect level for cadmium administered as a single oral dose to man (1). At the sixteenth meeting of the Joint FAO/WHO Expert Committee on Food Additives and Food Contaminants held in April 1972, it was recommended that the provisional tolerable intake of cadmium should not exceed 400-500 ,ug per week for an adult, i.e., 57-71 Jlg/day (9). Estimates have been made of a no-effect level for long-term exposure to cadmium (1). A threshold-effect dose of 200 Jlg/day has been suggested on the basis of epidemiological studies; this represents 12 ,ug/day absorbed (assuming 6% oral absorption). Taking the critical concentration for the human kidney cortex as between 200 and 250 mgjkg, it has been calculated that a person ingesting 248 ,ug/day (close to the 200 ,ugjday threshold) would achieve this critical kidney concentration at age 50 (1). With a daily intake within the range 5771 ,ug, as recommended in the report mentioned above (9), the renal cortex would receive about a quarter of the critical concentration of 200 mgjkg.
REFERENCES I. COMMISSION OF THE EuROPEAN COMMUNITIES. Criteria (dose/effects relationships) for cadmium. Oxford, Pergamon Press, 1978. 2. HIATT, V. & JUFF, J. E. The environmental impact of cadmium: an overview. International journal of environmental studies, 7: 277 (1975). 3. FLEISCHER, M. ET AL Environmental impact of cadmium: a review by the panel on hazardous trace substances. Envtronmental health perspectives, 7: 253 (1974). 4. FRIBERG, L. ET AL. Cadmium in the environment, 2nd ed. Cleveland, CRC Press, Inc., 1974.
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5. NATIONAL RESEARCH COUNCIL. Drinking water and health. Washmgton, DC, National Academy of Sciences, 1977. 6. Gutdelines for Canadtan drinking-water quality, 1978. Quebec, Supply and Services, 1979 (supporting documentation). 7. Ambient water quality criteria for cadmium. Washington, DC, US Environmental Protection Agency, Office of Water Regulations and Standards, Cntena and Standards Division, 1980. 8. WHO Technical Report Series, No. 505, 1972 (Evaluation of certam food addaives and the contaminants: mercury, lead and cadmiUm). 9. The hazards to health of persistent substances in water: annexes to a report of a WHO W'()rking Group. Copenhagen, WHO Regional Office for Europe, 1972. 10. COMMISSION OF THE EUROPEAN COMMUNITIES. Trace metals· exposure and health effects. Oxford, Pergamon Press, 1979. II. EnVIronmental health criteria for cadmiUm. Geneva, World Health OrganizatiOn, 1979 (interim report). 12. CHAMBERLAIN, A. C. ET AL. Report R.9198. Harwell, UK Atomic Energy Research Establishment, 1978. 13. FLANAGAN, P. R. ET AL. Increased dietary cadmium absorption m mice and human subjects with Iron deficiency. Gastroenterology, 74: 841 (1978). 14. UNDERWOOD, E. J. Trace elements in human and animal nutritiOn New York, Academic Press, 1977. 15. WHO Technical Report Series, No. 647, 1980 (Recommended health-based hmlls m occupational exposure to heavy metals: report of a WHO Study Group). 16. GLEASON, M. Clinical toxtcology of commercial products. Baltimore. Williams & Withams, 1969. 17. Toxicology of metals. vol. 11. Washington, DC. US Environmental Protection Agency, 1977 (Environmental Health Effects Research Series). 18. KRASOVSKY, G. N. ET AL. Toxic aRd gonadotropic effects of cadmium and boron relative to standards for these substances in drinking water. Environmental health perspectives, 13: 69 (1976). 19. NATIONAL INSTITUTE FOR OCCUPATIONAL SAFETY AND HEALTH. Criteria for a recommended standard. Occupational exposure to cadmium. Washington, DC, US Department of Health, Education and Welfare, 1977. 20. Cadmtum, nickel. some epoxides, miscellaneous industrial chemicals and general constderations on volatile anaesthettcs. Lyon, International Agency for Research on Cancer, 1976 (IARC Monographs on the evaluation of the carcmogenic nsk of chemicals to humans, vol. II). 21. KIPLING, M.D. & WATERHOUSE, J. A. H. Cadmium and prostatic carcinoma. Lancet, 1: 730 (1967). 22. LEMAN, R. A. ET AL. Cancer mortality among cadmium production workers. Annals of the New York Academy of Sctences, 271: 273 (1976).
6. CHROMIUM 6.1 General description 6.1. 1 Sources Most rocks and soils contain small amounts of chromium. The commonest ore is chromite in which the metal exists in the trivalent form; the ore is present in commercial quantities only in a few countries. Hexavalent chromium also exists naturally but infrequently. Chromium in its naturally occurring state is in a highly insoluble form; however, weathering, oxidation and bacterial action can convert it into a slightly more soluble form. Most of the more soluble forms in soil, especially any hexavalent chromium, are mainly the result of contamination by industrial emissions. Some contamination arises from the use of sewage sludge added to land. Contamination of air, water, and food has occurred as a result of man's use of chromium; traces of some natural chromium are present in food. Trivalent and hexavalent chromium occur in biological media, but only the trivalent form is stable, since hexavalent chromium is readily reduced by a variety of organic species (1).
The major uses of chromium are for chrome alloys, chrome plating, oxidizing agents, corrosion inhibitors, manufacture of chromium compounds, such as pigments, and in the textile, ceramic, glass, and photographic industries. 6.1.2 Occurrence in water Because of the low solubility of chromium generally, the levels found in water are usually low (9.7 Jlg/litre) (2); however, there are examples of contamination of water, in some cases serious, in which effluents containing chromium compounds have been discharged to rivers; the chromium may be in the trivalent or hexavalent form, either as a soluble salt or as insoluble particles and often as a chemical complex. The valency of the chemical form in natural waters is influenced by the acidity of the water (3). Trivalent chromium is converted to the insoluble hydroxide at neutral pH (4). Total chromium levels in raw water are usually 10 Jlg/litre or less; rarely do levels exceed 25 Jlg/litre except in highly contaminated situations (1-6). There is a tendency for the naturally occurring higher levels of chromium to be associated with waters of the greatest hardness (5). 91
92
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HEALTH-RELATED INORGANIC CONSTITUENTS
The levels of chromium in finished water entering the public supply are normally about the same as, or perhaps slightly lower than, those in raw source water. 6.2 Routes of exposure 6.2.1 Drinking-water Drinking-water normally contains very low concentrations of chromium (i.e., 5 Jlg/litre or less) (2). It would be very rare to find chromium levels as high as 20 Jlg/litre in tap-water. The level in tapwater may be a function of the period the water has been standing in contact with any plumbing fittings and as a consequence there could be variations in chromium levels when water is drawn at different times of the day at the same tap. Based on a water consumption of 2 Iitres per day, it can be estimated that the daily exposure to chromium in water might vary from substantially less than 10 Jlg to perhaps 40 Jlg per day on rare occasions. Since apparently trivalent chromium rarely occurs in drinking-water that is chlorinated, it is assumed that most waterborne chromium is in the hexavalent form (1). 6.2.2 Food Foodstuffs vary considerably in their chromium content, which ranges from 20 to 590 Jlgfkg. Because varying and conflicting figures have been quoted it is not possible to make a proper comparison between different foodstuffs; a review of the chromium content of 45 items of food has been published (5). Some seafoods appear to contain elevated levels of 0.02-0.21 mg/kg (8). Wines contain chromium and concentrations up to 60 Jlg/Iitre have been recorded (3). The chromium in foodstuffs is in both the trivalent and the hexavalent form. Some contamination of food could result during its preparation where plated or stainless steel utensils are used. There is a dearth of information on total dietary intake of chromium. In the USA it has been estimated to vary from 5 Jlg to 500 Jlg per day (2, 4, 9) and this range probably covers the vast majority of diets throughout the world. Average dietary chromium is probably about 100-300 Jlg/day. Virtually no information on the dietary intake for children exists. 6.2.3 Air There is only a limited amount of information on the levels of chromium in the air. Reported values (5) would suggest that mean concentrations in air in towns are typically about 0.02 Jlg/m 3 . In heavily industrialized areas however, concentrations of over 20 times this value have been recorded (5).
6.
CHROMIUM
93
Most of the chromium in the air will be in the form of fine particles, of which perhaps one half of those inhaled could become deposited in the respiratory tract. Based on a daily inspired volume of 22.8 m 3 and 50% alveolar retention, the daily quantity deposited in the lungs would be about 0.2 Jlg. 6.2.4 Smoking Cigarettes contain traces of chromium; a value of 1.4 Jlg per cigarette has been quoted (3) and some of this will be inhaled and absorbed. It is extremely difficult to give the precise exposure for smokers, but since only a small fraction of the chromium will be inhaled and only perhaps one half will be deposited in the lung, it can be estimated that the chromium retained in this way by smoking 20 cigarettes per day would not exceed a few micrograms per day. 6.2.5 Occupational exposure Levels of airborne chromium in a number of industrial situations, particularly in plating plants and where welding occurs, can be very much higher than in the ambient environment. Concentrations as high as hundreds of micrograms per cubic metre of air have been recorded. 6.2.6 Ingestion of dirt, dust, etc. Little is known about the levels of chromium in dust, but in general it is unlikely to be a very important source of exposure even where- young children have the opportunity to ingest dirt and dust. 6.2. 7 Relative significance of different routes of exposure From the discussion in section 6.2.1-6.2.6, it would appear that different individuals and population groups there is the opportunity a fairly wide range of exposure to chromium in water, food, air, and environment. To illustrate the potential relative contributions chromium in water in relation to overall intake, a few examples given here. 6.2. 7.1 Estimates of chromium uptake The estimates have been based on 10% absorption of chromium when ingested from food and water and 50% overall absorption and retention of chromium from respired air; it is further assumed that on a daily basis a person consumes 2 litres of water and breathes 20.0 m 3 of air per day containing chromium at a concentration of 0.02 J1g/m 3 . No reliable information is available for assessing the absorption of chromium by children or other sensitive groups. for for the of are
94
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(a) Food intake: 100 J19 of chromium per day Chromium concentration 1n water Weekly absorption of chromium (Jig) Water only Air only Food only Total Total water contribution ( %)
20 JiQ/Iitre 50 JlQ/Iitre 100 JlQ/Iitre
28 70 140
2 2 2
70 70 70
100 142 212
28 49 66
(b) Food intake: 300 J19 of chromium per day Chrom1um concentration in water Weekly absorption of chromium ( Jl9) Water only Air only Food only Total Total water contribution ( %)
20 JiQ/Iitre 50 JiQ/Iitre 100 JiQ/Iitre
28 70 140
2 2 2
210 210 210
240 282 352
12 25 40
6.3. Metabolism Chromium is absorbed through both the gastrointestinal and respiratory tracts. The amount absorbed differs in each system and depends on the form of chromium (1). Trivalent chromium is an essential form of the element for human beings. Hexavalent chromium is toxic. Discrepancies in values reported for chromium absorption from the digestive tract appear in the literature; exact values are not known. Trivalent chromium is poorly absorbed. From 0.1 % to 1.2% of trivalent chromium salts are absorbed, whereas 25 % of glucose tolerance factor (GTF), a chromium complex necessary for normal glucose tolerance, is absorbed. Natural chromium complexes in the diet seem to be more available for absorption than simple salts (1). It appears that at least 10% of the chromium in food is absorbed (10). Tissue chromium levels in rats exposed for one year to hexavalent chromium in drinking-water at a level of 25 mg/litre were approximately 9 times higher than the levels in tissues of rats similarly exposed to trivalent chromium (11). Hence it is assumed that the absorption rate for waterborne hexavalent chromium is at least 9 times that for trivalent chromium, i.e., approximately 10%, and elevated waterborne chromium is usually hexavalent. No quantitative information was found concerning the rates of absorption through the respiratory tract. Respiratory absorption would be expected to be dependent on particle size and solubility (2, 10). A reasonable estimate for the absorption of inhaled chromium is 50% of that inhaled.
6.
CHROMIUM
95
Chromium is distributed in human tissues in variable, low concentrations. Chromium levels in tissues other than the lungs decline with age {1). The largest stores of chromium in man are in skin, muscle, and fat; tissue levels are a function of sex, age, and geographical location (3). A homoeostatic mechanism, involving hepatic or intestinal transport mechanisms, prevents accumulation of excess trivalent chromium (1 2). Chromium is excreted slowly, mainly in the urine but also in the faeces. 6.4 Health effects Chromium appears to be necessary for glucose and lipid metabolism and for utilization of amino acids in several systems. It also appears to be important in the prevention of mild diabetes and atherosclerosis in humans {1). The harmful effects of waterborne chromium in man are associated with hexavalent chromium; trivalent chromium, which is regarded as a form of chromium essential to man, is considered practically non-toxic and no local or systemic effects appear to have been reported. People living in areas of the world where atherosclerosis is mild or virtually absent tend to have higher chromium levels in tissues than people from areas where the disease is endemic (13). Hexavalent chromium at 10 mgjkg of body weight will result in liver necrosis, nephritis, and death in man; lower doses will cause irritation of the gastrointestinal mucosa {14). Toxic effects have been observed in rats and rabbits when their drinking-water contained more than 5 mg of hexavalent chromium per litre (3), although in other studies up to 25 mg/litre produced no ill effects. A study of the effects of hexavalent chromium and cholesterol on the development of atherosclerosis in rabbits appears consistent with the hypothesis that chromium inhibits the development of experimentally induced atherosclerosis (1 5). Serum cholesterol levels, too, are higher in rats fed diets low in available chromium {1). Hexavalent chromium in high doses has been implicated as the cause of digestive tract cancers in man (3, 16), and there is firm evidence that there is an increased risk of lung cancer for workers who are exposed to high levels of chromium (3, 5). Two studies of the chromate (VI) pigment industry suggest a risk of lung cancer similar to that seen in the production industry in which there is a large excess risk of the disease (7), the greatest risk occurring in workers involved with processing either dichromate(VI) or chromium(VI) trioxide. Prostate cancer and maxillary sinus cancers have been reported in workers in other chromium-using industries (chromium platers); however, the risk of cancers at sites other than the lung cannot be assessed on the basis of the current data. Exposure to a mixture of chromium(VI) compounds of different solubilities (as found in the chromate production industry) carries the greatest risk for human beings. Epidemiological data do not allow an evaluation of the relative contributions to carcinogenic risk of
96
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metallic chromium, chromium(III) and chromium(VI), or of soluble versus insoluble chromium compounds (7). Other health effects related to industrial exposures have been reported, e.g., hexavalent chromium can produce cutaneous and nasal mucous-membrane ulcers and dermatitis (from skin contact) (4). The threshold level of exposure to hexavalent chromium needed to produce health effects is very unclear. According to an IARC monograph: "There is no evidence that at current levels of non-occupational exposure to chromium a health hazard exists" (17). REFERENCES I. TOWILL, L. E. ET AL. Reviews of the env~ronmental effects of pollutants, III: Chromium. Cincinnati, US Department of Commerce, National Technical Information Service, 1978 (PB-282-796). 2. NATIONAL RESEARCH CoUNCIL. Chromium. Washington, DC, National Academy of Sciences, I974. 3. Guidelines for Canadian drinking water quality, 1978. Quebec, Ministry of Supply and Services, 1979 (supporting documentation). 4. NATIONAL RESEARCH COUNCIL. Drinking water and health. Washington, DC, National Academy of Sciences, 1977. 5. COMMISSION OF THE EUROPEAN COMMUNITIES. Trace metals: exposure and health effects. Oxford, Pergamon Press, 1979. 6. KOPP, J. F. & KRONER, R. C. Trace metals in waters of the United States. Cincmnati, US Department of the Interior, 1967. 7. Some metals and metallic compounds. Lyon, International Agency for Research on Cancer, 1980 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 23), p. 303. 8. TEHERANI, D. K. ET AL. Determmation of heavy metals and selenium in fish from Upper Austrian waters. II. Lead, cadmium, scandium, chromium, cobalt, iron, zinc and selemum. Berichte der Oesterrezchischen Studiengesel/schaft fiir Atomenergie (1977) (SGAE No. 2797, pp. 1-21 (Chemical abstracts, 88, No. 49150e)). 9. UNDERWOOD, E. J. Trace elements in human and ammal nutritwn. New York, Academic Press, 1977. 10. FRIBERG, L. ET AL. Chromium. In: Handbook on the toxzcology of metals, Amsterdam, Elsevier/North-Holland BIOmedical Press, 1979. II. MACKENZIE, R. D. ET AL. Chrome toxicity studies. II. Hexavalent and trivalent chromium administered in drinkmg water to rats. A.M.A. archives of mdustrial health. 18: 232 (1958). 12. ScHROEDER, H. A. ET AL. Abnormal trace metals in man- chromium. Journal of chronic diseases, 15: 941 (1962). 13. ScHROEDER, H. A. The role of chromium in mammalian nutritiOn. American journal of clinical nutrllion, 21: 230 (1968). 14. KAUFMAN, D. B. E.T AL. Acute potassium dichromate poisonmg in man. American ;ournal of diseases of children, 119: 374 (1970). 15. NOVAKOVA, S. ET AL. [The content of hexavalent chromium in water sources and the effect on the development of experimental arteriosclerosis in warm blooded animals.] Gigiena i sanitari;a, 39(5): 78-80 (1974) (in Russian). 16. TELEKY, L. Krebs bei Chromarbeiten. Deutsche medzzinische Wochenschrift, 62: 1353 (1936). 17. Some inorganic and organometallic compounds. Lyon, International Agency for Research on Cancer, I973 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 2), p. 100.
7. CYANIDE 7.1 General description
7 .1.1 Sources Cyanide is present wherever life and industry occur. Both inorganic and organic forms of cyanide exist; the latter are normally classified as nitrites. Cyanides form part of life processes, in particular as metabolic intermediates. The commonest forms of cyanide include hydrogen cyanide (hydrocyanic acid in solution), those cyanide salts that are readily soluble in water, and metallocyanide complexes (1). The cyanide ion can combine with heavy metal ions to form complexes, some of which are very stable (2). Cyanide salts hydrolyse to give hydrocyanic acid (1). Cyanides are used in many industrial processes, e.g., in the production of acrylonitrile, adiponitrile, and methylmethacrylate. Cyanides are also used for extracting gold and silver, in the production of steel, for electroplating, and for the preparation of some intermediates in chemical synthesis (2). In some of these processes, contamination of the air and water can arise. The occasional use of cyanide for pest extermination may be a source of water contamination.
7.1.2 Occurrence in water Hydrocyanic acid dissociates to give the cyanide ion in water (1, 2); its dissociation is pH dependent, with the ionic form predominating above pH 8.2 (2). Conversion of cyanide to the much less toxic cyanate will occur at pH levels of 8.5 and above (3). In general, the levels in raw water appear to be low (i.e., less than 0.1 mg/litre) except in the case of serious contamination, mainly by industrial discharges to river or other sources (2). Metal-treating industries, coke and gas manufacture, and a variety of chemical producers can be major sources of cyanide contamination of water (4). Chlorination of potable water to a free chlorine residual under neutral or alkaline conditions will reduce the concentration of cyanide in the finished water to very low levels (3). Chlorination of water (at pH> 8.5) converts cyanides to innocuous cyanates (3, 5), which can ultimately be decomposed to carbon dioxide and nitrogen gas. 97
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7.2 Routes of exposure 7.2.1 Drinking-water Comprehensive information on levels of cyanide in drinking-water is not available; in general, it appears that the concentrations are well below maximum acceptable levels. 7.2.2 Food Most foods contain traces of cyanides. Some foods of plant ongm contain elevated natural levels of cyanide (e.g., almonds); cyanides can be found in fish living in contaminated waters (2). Cyanides are decomposed on heating and, thus, cooked foods will tend to contain lower levels (6). Typical daily intakes of cyanide from food do not appear to be properly known; in general, dietary intake is regarded to be low. The acceptable daily intake of cyanide residues via ingestion of fumigated foods has been set at 0.05 mg/kg of body weight (7). 7.2.3 Air Representative values do not appear to have been published, but it is generally considered that levels are extremely low. 7 .2.4 Other routes of exposure It has been reported that exposure to cyanide may be high in certain industrial situations (8). In such cases this may constitute a major exposure pathway.
7.2.5 Relative significance of different routes of exposure As insufficient information is available to deduce precise exposure from food sources (the main natural source of cyanide other than drinking-water), it is not possible to estimate reliably the relative contributions made by water and food. 7.3 Metabolism The cyanide ion is readily absorbed in animal species, and its highly poisonous effects are induced rapidly. Cyanide blocks oxidative processes in the cells of the carotid and aortic bodies, allowing anaerobic products, e.g., lactic acid, to accumulate in them, thus stimulating respiration. This reaction is due to the combination of cyanide with the catalytic iron group of cytochrome oxidase, preventing the donation of electrons to molecular oxygen. The absorption of oxygen by the cells is inhibited, i.e., cellular oxidation cannot proceed,
7.
CYANIDE
99
and the main supply of energy to the cells ceases. Without oxidation of glucose, neurones convert glucose to lactic acid at a greatly increased rate. In animals the increase in the lactic acid content of the brain, even with small doses of cyanide, may cause coma and convulsions with irreversible cerebral damage, although the rest of the body is unharmed (9). Low exposures to cyanide are not fatal to human beings who have an efficient detoxification system whereby the cyanide is converted to the thiocyanate ion, which is non-toxic at low levels (2). 7.4 Health effects A single dose of 50-60 mg for a human being is usually fatal (3). Exposures of 2.9-4.7 mg of cyanide per day are regarded as noninjurious to humans, owing to the highly efficient detoxification system in the human body in which the cyanide ion is converted to the relatively non-toxic thiocyanate ion through the rhodanese and thiosulfate enzyme system (1). Higher exposures may be fatal. On the basis of animal experiments, it has been calculated that an acceptable daily intake for man is 8.4 mg of cyanide (8).
REFERENCES I. Quality critenafor water. Washington, DC, US Envuonmental ProtectiOn Agency, 1976
{EPA-440/9-76-023). 2. Guidelines for Canadian drinkmg water quality, 1978. Quebec, Ministry of Supply and Services, 1979 {supporting documentation). 3. National interim primary drinking water regulations. Washington, DC, US Environmental Protection Agency, 1976. 4. Gorrs, R. M. ET AL. Treatment of industrial wastes at mumcipal water pollutwn control plants. Ontario, Ontario Water Resources Commission, 1966 {Proceedings, Ontano Industrial Waste Conference), p. 151. 5. Cyanides. In: Kirk, R. E. & Othmer, D. F., ed. Encyclopedia of chemical technology, 2nd ed. New York, John Wiley and Sons, 1965. vol. 6, p. 574. 6. LEDUC, G. ET AL. The use of sodium cyanide as a fish erad1cant in some Quebec lakes. Natura/isle canad1en, 100: I {1973). 7. Internatwnal standards for drinkmg water, 3rd ed. Geneva, World Health OrganizatiOn, 1971. 8 US Envuonmental Protection Agency. Water quality criteria; availability. Federal reg1ster, 44: 43667 {1979). 9. PASSMORE, R. & RossoN, J. S. A compamon to med1cal studies, vol. 2. Oxford, Blackwell Sc1entific Publications, 1970, pp. 15-25.
8. FLUORIDE 8.1 General description 8.1.1 Sources Fluorine is a fairly common element, representing about 0.3 g/kg of the earth's crust (1). It exists in the form of fluorides in a number of minerals, of which fluorspar, cryolite, and fluorapatite are the commonest; many rocks contain fluoride minerals. Fluorides are used industrially in the production of aluminium and are commonly present in phosphate fertilizers, bricks, tiles, and ceramics; they are also used in metallurgy (2). Fluorides are now frequently added to certain pharmaceutical products, including toothpastes and vitamin supplements (2). Owing to industrial activity, involving the use of so many fluorinecontaining substances, fluoride contamination 6f the environment is ubiquitous. Thus, plants, foodstuffs, and water all contain traces of fluoride.
8.1.2 Occurrence in water Traces of fluorides occur in many waters and higher concentrations are often associated with underground sources. In areas that are rich in fluoride-containing minerals, e.g., fluorapatite, well-waters may contain up to about 10 mg of fluoride per litre or even more (3, 4). The highest natural level reported is 2800 mgjlitre (1). Most waters contain below 1 mg of fluoride per litre (1). Occasionally, fluorides may enter a river as a result of industrial discharges.
8.2 Routes of exposure 8.2.1 Drinking-water The levels of fluoride in tap-water are very similar to those found in the source water, except where fluoridation of the supply is practised. In general, unfluoridated supplies contain less than 1 mg of fluoride per litre, but, depending on the type and situation of the source, may very occasionally contain up to 10 mg/litre (2). In most parts of the world such sources have by now been identified. Where fluoridation of water supplies is practised, fluoride concentration is normally within the range 0.6-l. 7 mgjlitre, ambient air temperature usually being the deciding 100
8.
FLUORIDE
101
factor (5). With a consumption of 2 litres of water per day, between 1.2 and 3.4 mg of fluoride per day could thus be ingested from drinkingwater in those areas where fluoridation is practised. Elsewhere, the daily exposure will range from a fraction of a milligram to perhaps 20 mg in very exceptional circumstances.
8.2.2 Air Fluorine-containing compounds are present in the air, mainly arising from industrial emissions. Concentrations in air will vary, depending on the type of industrial activity, but it has been estimated that the general exposure, equivalent to less than 1 Jlg/m 3 of air (2), is insignificant in comparison with ingested fluorine (6).
8.2.3 Food Virtually all foodstuffs contain at least traces of the element. All vegetation contains some fluoride, which is absorbed from soil and water. Some foods may contain high levels, particularly fish, some vegetables, and tea (1, 2). For example, the fluorine content of some fish can be as high as 100 mg/kg, and tea may contain more than twice that concentration, in contrast to most other foodstuffs, which rarely exceed 10 mg/kg (1). The use of fluoridated water in food processing plants can often double the level of fluoride in prepared foodstuffs. Estimates have been made in various countries to ascertain the daily dietary intake of fluoride, which for adults ranges from 0.2 to 3.1 mg (1). For children, e.g., the 1-3 years age group in the USA, an intake of about 0.5 mg/day has been estimated (1).
8.2.4 Other routes of exposure 8.2.4.1 Industrial exposure A number of industrial processes are known to release fluorinecontaining compounds into the air of the work place, and there are numerous documented examples of human exposure, in particular in aluminium smelters and glass manufacture (1). Such exposure to fluorine under these conditions (i.e., concentrations of up to several milligrams per cubic metre of air) could provide the major contribution to the total exposure. However, with improvements in working conditions, this route of exposure becomes far less significant. There is little information on the exposure to fluoride from tobacco smoke; however, it is not considered to be an important source of fluoride in comparison with the other routes of exposure.
102
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HEALTH-RELATED INORGANIC CONSTITUENTS
8.2.4.2 General population exposure Various products, such as toothpaste, tooth powders, mouth washes, chewing gum, vitamin supplements, and drugs, may contain added quantities of soluble fluoride, mainly in inorganic form. Such compounds are commonly added to dentifrices, typically at concentrations of about 1 g/kg (1). Studies have shown that significant quantities of fluoride can be absorbed by this route, and a possible absorption of about 50 Jlg of fluoride per "brushing" has been demonstrated (1). The use of topical applications of fluoride solutions can contribute to an increased absorption (7). Mouth washes can provide as much as 2 mg of fluoride (1). There is a range of different products that incorporate fluorides, including tablets containing sodium fluoride, used as an anticariogenic agent. Regular use of such tablets can provide up to about 1 mg of fluoride per day (1). 8.2.5 Relative significance of different routes of exposure The following table shows the influence of fluoridated water on the total dietary intake in adults. Food intake 1 mg of fluoride per day; drinking-water containing 0.5, 1.0, and 1.5mg of fluoride per litre; water consumption 2 litres per day Weekly intake of fluonde (mg) Fluoride concentration in water Water only Air only" Food only Total water contribution (%)
Total
0.5 mgjlitre 1 Omgjlitre 1.5 mgjl1tre
7.0 14.0 21.0
0.0 0.0 0.0
7.0 70 7.0
14.0 21.0 28.0
50 67 75
• Negligible contnbut1on from amb1ent a1r (mput by other routes of exposure not considered m these calculations)
Estimates of the uptake of fluoride from food and water are not provided here, since the efficiency of absorption is generally quite high (8) and figures quoted for intake, given above, would be fairly close to those relevant to absorption. 8.3 Metabolism Fluoride ingested with water is almost completely absorbed (J); fluoride in the diet is not as fully absorbed as from water, but the absorption is still rather high, although in the case of certain foods (e.g., fish and some meats) only about 25 'lo of the fluorides may be absorbed (8).
Absorbed fluoride is distributed rapidly throughout the body. It is retained mainly in the skeleton, and a small proportion is retained in the teeth (7). The amount of fluoride in bone increases up to the age of 55
8.
FLUORIDE
103
years (9). At high doses, fluoride can interfere with carbohydrate, lipid, protein, vitamin, enzyme, and mineral metabolism (1). Many of the symptoms of acute fluoride intoxication are the result of its binding effects with calcium (6). Fluoride is excreted primarily in the urine. The excretion is influenced by a number of factors, including the general health of the person and his previous history of fluoride exposure (2); the rate of retention decreases with age, and most adults can be regarded for practical purposes as "in balance" (6). Under this "steady state" condition, the fluoride present in the body is sequestered in calcified tissues; most of the remainder is present in plasma and thus available for excretion. Skeletal sequestration and renal excretion are the two major ways by which the body prevents the accumulation of toxic amounts of the fluoride ion.
8.4 Health effects Fluorine has been fairly conclusively demonstrated to be an essential element for some animal species (7); in particular, fertility and growth rate are improved as a result of relatively small doses of fluorine (7). Once fluoride is incorporated into teeth, it reduces the solubility of the enamel under acidic conditions and thereby provides protection against dental caries. There is good evidence to show that the presence of fluoride in water results in a substantial reduction of dental caries in both children and adults (2). The incidence of caries decreases as the concentration of fluoride increases to about 1 mg/litre, although mottling may sometimes occur even to an objectionable degree when the level rises to 1.5-2.0 mg/litre ( 6). Long-term consumption of water containing I mg of fluoride per litre may lead to such mottling in patients with long-standing renal disease or polydipsia ( 6), but only in persons in whom the teeth are under mineralization, i.e., children 0-7 years of age. Skeletal fluorosis has been observed in persons when water contains more than 3-6 mg of fluoride per litre depending on intake from other sources. Intakes of 20-40 mg of fluoride per day (or more where water contained in excess of 10 mg/1itre (J)) over long periods have resulted in crippling skeletal fluorosis (6). It has been accepted that 1 mg/litre is a safe level in relation to the fluoridation of water supplies, and the recommended control limits in water (5, 10) are around this figure, the exact concentrations depending on the air temperature. In high doses, fluoride is acutely toxic to man. Pathological changes include haemorrhagic gastroenteritis, acute toxic nephritis, and various degrees of injury to the liver and heart muscle (1). The acute lethal dose is about 5 g as sodium fluoride, i.e., about 2 g of fluoride (2). In animals, a variety of severe symptoms have been observed as a result of environmental exposure to fluoride in highly contaminated areas (7). Chronic effects from high exposure in man are primarily related to mottling of teeth and fluorosis, in which bone structure is affected, sometimes to a very alarming degree, producing serious crippling (1).
104
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Chronic effects on kidneys have also been observed, generally in persons with renal disorders (1). Other less common problems, including effects on the thyroid, are known as a result of high exposure {1). Initial signs and symptoms of intoxication are vomiting, abdominal pain, nausea, diarrhoea, and even convulsions (6). Epidemiological studies in areas where the concentration of fluoride in the water is naturally high have only rarely shown adverse effects (6); these have included mottling of teeth and skeletal fluorosis in areas of exceptionally high levels in water (2). In areas having optimal levels of fluoride in water, obvious effects (i.e., objectionable dental fluorosis) were observed in two children known to be suffering from diabetes insipidus (6). It has been suggested that mongolism and cancer are associated with elevated levels of fluoride in water. The idea that mongolism could be related to fluoride exposure stems from one limited study where the prevalence of mongolism registered at institutions was recorded in relation to fluoride levels in the water {11); the study has been very severely criticized, however, by the Royal College of Physicians in the United Kingdom (6). For nearly 30 years, various epidemiological studies have been carried out to assess whether there is a link between cancer and fluoride in water. In a few cases, claims were made that there was a positive association, but these have been criticized (2). It is now generally considered that there is no acceptable evidence whatever that fluoride in water is carcinogenic to human beings (2, 6, 12, 13). The question of sensitivity to fluoride has been raised (6). In general, claims that there are some people who are sensitive to fluoride have been dismissed. For example, no particular sensitivity has been recorded in the millions of tea drinkers who would ingest considerable quantities of fluoride from the infusion of tea leaf. However, the possibility of sensitivity or some idiosyncratic reaction cannot be completely dismissed (6), although from the evidence available the incidence of such cases would be expected to be small. Suggestions that fluoride is mutagenic or teratogenic or that it is related to birth defects have been thoroughly reviewed and have not proved justified (6).
REFERENCES I. F/uondes and human health. Geneva, World Health Organization, 1970 (Monograph
Senes, No. 59). 2. Gutdelines for Canadian drinking water quallly, 1978. Quebec, Ministry of Supply and Services, 1979 (supportmg documentation). 3. F/uondes. Washington, DC, Natwnal Academy of Sciences, 1971. 4. BuLusu, K. R. ET AL. Fluorides in water, defluoridatwn methods and their limitatiOns. Journal of the Institution of Engineers (India), 60 (1979). 5. International standards for drmking water. Geneva, World Health Organization, 1971. 6. NATIONAL RESEARCH COUNCIL. Drinking water and health. Washington, DC, National Academy of Sciences, 1977.
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7. UNDERWOOD, E. J. Trace elements in human and animal nutrition. New York, Academic Press, 1977. 8. NEWBURN, E. & ZIPKIN, I. Fluoride metabolism. Fluoride and dental caries. Springfield, IL, Charles Thomas, 1976. 9. JACKSON, D. & WEIDMANN, S. M. Fluorine in human bone related to age and the water supply of different regions. Journal of pathology and bacteriology, 76: 451 (1958). 10. European standards for drinking water. Geneva, World Health Organization, 1970. II. RAPAPORT, I. Nouvelles recherches sur le mongolisme. A propos du role pathogenique du fluor. Bulletin de l'Academie nationale de Medecine, 143: 367 (1959). 12. Some aromatic amines, anthraquinones and nitroso compounds, and inorganic fluorides used in drinking-water and dental preparations. Lyon, International Agency for Research on Cancer, 1982 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 27). 13. CLEMMESEN, J. The alleged association between artificial fluoridation of water supplies and cancer: a review. Bulletin of the World Health Organization, 61: 871-883 (1983).
9. HARDNESSa 9.1 General description
Hardness of water is not a specific constituent but is a variable and complex mixture of cations and anions. Hardness is predominantly due to calcium and magnesium, although strontium, barium, and other polyvalent ions contribute. Hardness is commonly expressed as mg of calcium carbonate equivalent per litre, and this is the unit adopted throughout this document. Several other units are used in various countries. Traditionally, hardness is a measure of the capacity of water to react with soap. It is often divided into carbonate (temporary) and non-carbonate (permanent) types of hardness. 9.1.1 Sources
Calcium and magnesium are common elements present in many minerals. Among the commonest sources of calcium and magnesium in water are limestones, including chalk (calcium carbonate). Calcium and magnesium are present in a great number of industrial products and they are common constituents of food. A minor contribution to the total hardness of water is made by such polyvalent ions as zinc, manganese, aluminium, strontium, barium, and iron, dissolved from minerals such as sphalerite, armangite, bauxite, strontianite, witherite, and phosphosiderite. 9.1.2 Occurrence in water
Although most calcium compounds are not easily soluble in pure water, the presence of carbon dioxide readily increases their solubility, and sources of water containing up to 100 mg of calcium per litre are fairly common (1-4); sources containing over 200 mg of calcium per litre are rare (1-4). Many salts containing magnesium are easily soluble and water sources containing levels of magnesium at concentrations up to 10mg/litre are common (1-4). Water sources rarely contain more than 100 mgjlitre (1-4), and calcium hardness usually predominates. The buffering capacity of water, normally described as alkalinity, is closely associated with water hardness. Thus, such anions as hydroxide, a A discussion of some other aspects of the effect of hardness on water quality will be found in Part V, section 5, p. 264.
106
9.
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107
bicarbonate, and carbonate have a significant influence, and so to a lesser degree do phosphate and silicate; molecular species of weak acids also contribute. 9.2 Routes of exposure 9.2.1 Drinking-water
Although water is sometimes artificially softened at the watertreatment plant, generally the raw water hardness is similar to that found in drinking-water piped to a household. Hardness may range from less than lOmgflitre to over 500mg/litre (4); water sources may have a hardness of less than 50 mgflitre; values above 500 mgflitre are relatively uncommon in most countries (2, 3, 5). 9.2.2 Food
Virtually all foods contain calcium and magnesium. Typical diets provide about 1000 mg of calcium per day (6) and 200-400 mg of magnesium (1, 7). The predominant source of ingested calcium and magnesium is normally food. Dairy products are a particularly rich source of calcium (2); magnesium tends to be associated more with meat and foodstuffs of plant origin (6). 9.2.3 Air, occupational exposure, and cigarette smoking
Although all these provide a route of exposure to man, the contribution in relation to food is trivial. 9.2.4 Relative significance of different routes of exposure
Only food and water provide important routes of exposure and therefore only these sources are considered. Some estimates of possible situations are given below. All the estimates relate to adults.
(a) Food intake: 1000mg of calcium per day Weekly intake of calcium (mg) Concentration of calcium in water 25mgflitre 100 mgflitre 200 mgtlitre Water only 350 1400 2800 Food only 7000 7000 7000 Total 7350 8400 9800
Water contribution (%)
5 17 29
Typ1cal water contnbut1on in relation to total input of calcium 1s about 5--20%.
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(b) Food intake: 200mg of magnesium per day Weekly intake of magnesium (mg) Concentration of magnesium in water 10 mgflitre 50mgflitre 100 mgjlitre Water only 140 700 1400 Food only 1400 1400 1400 Total 1540 2100 2800
Water contribution (%)
9 33 50
(c) Food intake: 400mg of magnesium per day Weekly intake of magnesium (mg) Concentration of magnesium in water 10 mgflitre 50mgjlitre 100 mgjlitre Water contribution (%) 5 20 33
Water only 140 700 1400
Food only 2800 2800 2800
Total 2940 3500 4200
Typical contribution from water in relation to the total onput of magnesium is about 5-20%.
Although some information on dietary intake is available for children, these contributions have not been calculated since it can be estimated that the relative proportions of the different routes of exposure in children are roughly similar to those in adults. Separate calculations have not been made for the uptake of calcium and magnesium. Although about 30% and 35% respectively of calcium and magnesium appears to be absorbed from the diet, little reliable information is available on the uptake of calcium from tap-water, and therefore only crude estimates of uptake would be possible.
9.3 Health effects
There is some suggestive evidence that drinking extremely hard water might lead to an increased incidence of urolithiasis. This has been suggested to account for urolithiasis in a small human population in the USSR where the local tap-water contained 300-500mg of calcium per litre (8); it has also been demonstrated in animals drinking extremely hard water (200-400 mg of calcium per litre), when they had been subjected to a high ambient temperature of 30 oc (9). The occurrence of drinking-water containing as much as 500 mg of calcium per litre must, however, be very rare. Thus, there appears to be no firm evidence that water hardness causes ill effects in man (2). Conversely there have been a number of studies, where the results suggest that water hardness protects against disease.
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109
9.3.1 Water hardness and cardiovascular disease In 1957 in Japan, it was demonstrated that there was a close association between death rates from strokes and the acidity of riverderived drinking-water (4). Since that time, a number of studies in various parts of the world have demonstrated that there is a highly statistically significant negative association between water hardness and cardiovascular disease (3, 10). In most studies, the calcium concentration has shown the strongest correlation, but the magnesium content of the water has been indicated as the most significant correlating factor in some Canadian studies (4, 7). Some small-scale studies, however, have not confirmed the relationship (1 1-1 3). Uncertainty remains about the magnitude of a possible "water effect" and the extent to which confounding factors might account for it, e.g., factors such as air temperature, rainfall, latitude, longitude, socioeconomic factors, town type, and air pollution. Nevertheless, a strong statistical association can be demonstrated when allowance is made for a number of confounding variables (2, 14). A special international scientific colloquium on this particular subject was held in 1975 (3). In a recent retrospective largescale study (14) of 253 towns in Great Britain, after allowing for climatic conditions and certain social factors, mortality from stroke and ischaemic heart disease was found to be strongly related to water hardness, but only up to about 170 mgjlitre as CaC0 3 . Of many water factors analysed, the correlation with hardness and the calcium content of water was high, although other water parameters, many of which are intercorrelated with hardness, also provided strong statistical associations. Such retrospective studies are limited in what they can achieve and research is continuing; particular hope is placed on prospective studies of cardiovascular risk factors in selected population groups to assess the significance of a wide range of water parameters (14). Several hypotheses have been proposed in an attempt to account for the relationship, but there is at present no definite evidence that hardness, or its major constituents calcium and magnesium, are involved. The two most quoted hypotheses relate to (a) a constituent (or constituents) in hard water being protective in some way and (b) a substance (or substances) in soft water (e.g., metals leached from piping materials) promoting the disease. In the case of the protective hypothesis it is often considered that the diet provides an adequate supply of calcium and magnesium, although for magnesium there is the possibility of a dietary deficiency in some situations (7). However, the presence of other elements, e.g., lithium, chromium, vanadium, and silicon, could have a protective role (15, 16). Lead and cadmium, which can be leached from plumbing material, have been suggested as possibly promoting the disease, but there is no firm evidence that they are involved in this way (3).
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9.3.2 Water hardness and other diseases The results of several studies have suggested that a variety of other diseases are correlated with the hardness of water. These include certain nervous system defects, anencephaly, perinatal mortality, and various types of cancer (2, 17-20). Although some of these findings have been demonstrated in different countries (2), there is still considerable doubt about their significance. These associations may merely reflect disease patterns that can be explained by social, climatological, and various environmental factors, rather than by the hardness of water.
REFERENCES I. NATIONAL RESEARCH COUNCIL Drinkmg water and health. Washington, DC, National Academy of Sciences. 1977. 2. Guidelines for Canadian drinking water quality. Quebec, Mimstry of Supply and Services, 1979 (supporting documentatwn). 3. AMURIS, R. ET AL., ed. Hardness of drinking water and public health. Oxford, Pergamon Press, 1975 (Scientific colloquium, Luxembourg, 1975). 4. MARIER, J. R. ET AL. Water hardness, human health, and the importance of magnesium. Ottawa, Canada. National Research Council, 1979. 5. Qualuy crl/eria for water. Washington, DC, US Environmental ProtectiOn Agency, 1976. 6. WHO Techmcal Report Senes, No. 532, 1973 (Trace elements m human nutrition: report of a WHO Expert Committee). 7. NERI, L. C. & JOHANSEN, H. L. Water hardness and cardiovascular mortality. Annals of the New York Academ}' of Sciences, 304: 203 (1978). 8. BOKINA, A. I. ET AL. [Hygienic assessment of drinking water hardness as a factor favounng the development of urolithiasis.] Gtgiena i samtari)a, 30(6): 3 (1965) (m Russian). 9. BOKINA, A. I. & YURIEVA, V. K. [Shifts of certam biochemical mdices in persons after long-term use of hard drinkmg water.] Gigiena i sanitari)a, 31(12): 33 (1966) (in Russian). 10. KOBAYASHI, J On the geographical relationship between the chemical nature of river water and death-rate from apoplexy. Berichte des Ohara lnstituts jiir Landwirtschaftliche Biologie, 2: 12 (1957). II ALLWRIGHT, S. P. A. ET AL Mortality and water hardness in three matched communities 10 Los Angeles. Lancet, 2: 860 (1974). 12. BIERENBAUM, M. L. ET AL Possible toxic water factor 10 coronary heart disease. Lancet, 1. I 008 (1975). 13 MEYERS, D. lschaemic heart disease and the water factor. A vanable relationship. Brttish journal of pret•entil'e and soc tal medicine, 29: 98 ( 1975). 14. PococK, S. J. ET AL Bntish regwnal heart study: geographic variations in cardiovascular mortality, and the role of water quality. British medical ;ourna/, 280: 1243 (1980). 15. VooRs, A. W. Lithium in the drinking water and atherosclerotic heart death: epidemiOlogical argument for a protective effect. Amencan journal of epidemwlogy, 92: 164 (1970). 16. SCHWARTZ, K Silicon, fibre, and atherosclerosis. Lancet, 1: 454 (1977). 17. STOCKS, P. Incidence of congenital malformations m the regwns of England and Wales. British journal of prevenlll'e medicine, 24: 67 (1970). 18. HART, J. T. The distribution of mortality from coronary heart disease in South Wales. Journal of the Royal College of General Practitioners, 19: 258 (1970). 19. FEDRICK, J. Anencephalus and the local water supply. Nature, 227: 177 (1970). 20. LOWE, C. R. ET AL. Malformations of the central nervous system and softness of local water supplies. Brttish medical Journal, 2: 357 (1971).
10. LEAD 10.1 General description 10.1.1 Sources
Lead is a natural constituent of the earth's crust at an average concentration of about 16 mg/kg (1). It is present in a number of minerals, the principal one being galena (lead sulfide); most countries have lead deposits of one sort or another. Lead has been widely used for many centuries, and in many places some contamination of the environment has occurred as a result of the mining and smelting processes used or from the use of products made from it. Consequently it is present in air, food, water, soil, dust, and snow. Lead in the environment exists almost entirely in the inorganic form, but small amounts of organic lead result from the use of leaded gasoline and from natural alkylation processes that produce methyl lead compounds (2). Lead is used widely for a variety of purposes, including the manufacture of acid accumulators, alkyl lead compounds for gasoline, solder, pigments, ammunition, caulking, and cable sheathing. Its use as roofing materials and piping materials, including pipes used for potable water, is currently being discontinued and discouraged. 10.1.2 Occurrence of lead in water
The natural lead content of lake and river water worldwide has been estimated to be 1-10 J.tg/litre (1, 3). Although higher values have been recorded where contamination has occurred, particularly from industrial sources, such situations are relatively rare, since there are a number of natural mechanisms that control the levels. The concentrations in finished water (i.e., water after treatment) prior to its distribution are generally lower than in source waters since lead is partially removed by most conventional water-treatment processes (4). The levels in drinkingwater, however, can be much higher owing to the use of lead service pipes running from the street to a dwelling, from lead plumbing, and/or lead-lined storage tanks (4, 5). Particularly high lead levels can result when the water is aggressive, soft, or has a low pH. These conditions tend to produce the highest levels (1, 4, 6). Lead pipes have not been used extensively throughout the world, but in some towns and cities in certain countries lead is still being widely used. This occasionally results in undesirably high levels of lead in the tap-water (1, 4). Ill
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10.2 Routes of exposure 10.2.1 Drinking-water In most countries, the levels of lead in domestic tap-water are relatively low, i.e., normally well below 10-20 Jlg/litre. In some places, however, they may be rather high. For example, in Scotland, the water is extremely soft, has a low pH, and lead plumbing and lead-lined storage tanks are common. As a result, it has been estimated that in over 10% of houses in Scotland (population about 5 million), the firstdraw water contains more than 300 Jlg of lead per litre (7). Values in excess of 2000 Jlg/litre have been recorded in some places in the world (8). It is very difficult, however, to define precisely the average exposure in terms of the concentration of lead in the water because of the very wide variation of levels produced at the tap. These levels depend critically on factors such as the stagnation time of the water in a lead service pipe or in household plumbing. Even in the same water supply area, there may be considerable home-to-home variations in the level of lead because of differences in the length of pipe, water-use patterns, and types of deposit that have built up. Based on a water consumption of 2 litres per day, calculations show that the daily intake of lead from water varies from 10 to 20 Jlg to 1 mg or more. Such estimates are based on the assumption that all lead is consumed. It has been shown, however, that in the preparation of certain beverages, e.g., tea (9), not all of the lead in the water will appear in the beverages prepared. On the other hand, tap-water is also used for cooking and food preparation, which provides an additional opportunity for ingesting lead from domestic tap-water. 10.2.2 Food Lead is present in a wide variety of foodstuffs. The amounts vary depending on the type of food. For instance, canned foods (3, 10, 11) tend to contain the highest levels if lead solders have been used in the manufacture of the can. Many fresh vegetables, cereals, and fruits contain small quantities of lead as a result of some limited absorption of the metal from the soil in which they are grown, and because of the deposition on surfaces from lead in the air. Lead is also present in milk and dairy products and in wine (10). Because of the large differences in individual diets, precise calculations of lead intake are not possible. Estimates of typical daily intakes range from less than 100 to over 500 Jlg of lead (1, 5, 8, 10, 12); the worldwide average for adults is about 200 Jlg/day. Estimates of levels of lead in the diet appear to have been falling in recent years (8, 10). Women generally eat less than men and their intake of lead in food is consequently lower. It has been estimated that children aged 1-5 years ingest about 90 Jlg of lead per day (10). Additional lead in food can arise from contamination
10.
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113
by cooking vessels, such as pots that have soldered joints and some glazed earthenware utensils. Some lead comes from tap-water used for preparing food. In general, the major source of ingested lead is food.
10.2.3 Air In rural areas, average levels of lead of 0.1 J.J.g/m 3 of air can be found (3, 10, 12). Average city levels are typically in the range 0.5-2 J.J.g/m 3 (8, 10). The levels in any one particular area will depend on the type and the extent of the emission sources (e.g., traffic, industry) and the natural dispersion conditions of the area (prevailing weather conditions). At present, most of the airborne lead in non-industrialized cities comes from motor vehicle traffic. In general, people living near busy highways will be most exposed. In some industrialized areas, average ambient air levels as high as 6 J.J.g/m 3 have been recorded (3, 8). Deposited lead from the air contaminates soil, and levels of 2 g/kg have been reported (8); in highly contaminated soils, levels of over 10 gjkg have been recorded (8). Most of the lead in the air is in the form of fine particles. When these particles are inhaled, only 20-60% will be deposited in the respiratory system (8). On the basis of a daily respired volume of air of 15-22.8 m 3 (1, 8) a typical daily intake for an urban dweller (exposed to 1 J.l.g of lead per cubic metre of air and with 40% retention) would be 6-9 J.l.g. Much of this retained lead will ultimately be absorbed.
10.2.4 Occupational exposure Levels of airborne lead in industrial work areas can be much higher than in the general environment. Levels of up to 100 J.l.gfm 3 of air are not uncommon (8).
10.2.5 Smoking Small quantities of lead are found in tobacco, but m general the exposure from this source is relatively small.
10.2.6 Ingestion of soil and dust and chewing of paint Soil, dust, and particularly household paints (especially old paint) contain elevated levels of lead. Because of the "hand-to-mouth" activity (pica) of young children, they may be subject to significant exposure from such sources (1, 3-8). Despite a host of studies analysing the relative contribution of soil, dust, and paint to lead exposure in children, the precise contributions from these sources remain uncertain (6).
10.2. 7 Relative significance of different routes of exposure For different individuals and population groups the exposure to lead from water, food, air, etc., can vary significantly. Since the relative
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contributions of each of these sources can also vary widely it is not possible to provide comprehensive information for a wide range of circumstances_ A range of some environmental situations has, however, recently been considered and numerous estimates provided (10)_ To give some idea of possible situations, a few simplified examples are given here_ The last two columns of the tables give estimates of the relative contribution of water to the total intake of lead and to the uptake of lead by the body. An important group omitted in the following tables is that of infants (up to one year old). The available information about lead in their diet and its absorption is insufficient to allow reasonable estimates to be made. It is likely, however, that the contribution of daily lead via drinking-water in this group is as high, or perhaps higher, than that for the 1-5-year-old children, which is given in the tables. No account is taken in the calculations of the contributions made by cigarette smoking, occupational exposure, or various other sources. 10.2.7.1 Estimates of weekly intake and uptake of lead by the body in adults The tabular examples below assume that an average person consumes 2 litres of water per day, and breathes air at the rate of 20m 3 jday (10). Lead absorption from individual sources is assumed not to be influenced by the uptake from other sources; absorption from food and water is assumed to be 10% of the intake and there is assumed to be complete absorption of the 40% retained from inhalation.
(a) Daily intake of 100 Jlg of lead per day in food; 1.0 Jlg of lead per m 3 of air Weekly 1ntake of lead (m9) Lead concentration in water Water only Air only Food only Intake ratio: waterjtotal (%) Uptake rat1o: waterjtotal (%)
Total
20 119/l1tre 50 119/l1tre 100 !19/htre
0.28 0 70 1 40
0.14 0.14 0.14
0.70 0.70 0.70
1.12 1.54 2.24
25 47 63
18 35 52
(b) Daily intake of 300 Jlg of lead per day in food; 1.0 Jlg of lead per m 3 of air Weekly mtake of lead (m9) Lead concentration in water Water only A1r only Food only Intake ratio: waterjtotal (%) Uptake rat1o. water/total (%)
Total
20 !19/litre 50 !19/htre 100 !19/litre
0.28 0.70 1.40
014 0.14 0.14
210 2.10 210
2.52 2.94 3.64
11 24 40
10 20 34
10.
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115
10.2.7.2 Estimates of weekly intake and uptake of lead by the body in children aged 1-5 years
The example below assumes that water is consumed at the rate of I litre per day and air is breathed at the rate of 4.7 m 3 /day {10). The basic assumptions for lead absorption are the same as for adults except that the absorption from food and water is 50% of the intake (4, 8). Daily intake of 93 J19 of lead per day in food; 1.0 J19 of lead per m 3 of air Weekly intake of lead (mg) Lead concentration in water 20 llQ/Iitre 50 llQ/Iitre 100 ll9/litre Water only 0.14 0.35 0.70 Air only 0.03 0.03 0.03 Food only 0.65 0.65 0.65 Intake ratio: water/total (%) 18 35 51 Uptake ratio: water/total (%) 17 35 51
Total 0.82 1.03 1 38
10.3 Metabolism An important consideration is the proportion of the lead in drinkingwater that is actually absorbed when ingested. Although little is known about the absorption of fine particles of lead present in·tap-water, some information exists on the intestinal uptake of lead from aqueous solutions containing dissolved lead. Generally a figure of about I 0 % (4, 5, 8) is regarded as typical of the fraction absorbed for an aduJt,a but this value depends on whether the water is consumed on a full or an empty stomach. For example, humans who fast for 6 hours before and after an oral dose of lead ions have markedly increased absorption (e.g., 50% or more) (13, 14). This finding has been confirmed in mice {15). Other factors influence the absorption of lead from the gastrointestinal tract, such as the presence of elements such as calcium, phosphorus, iron, copper, and zinc in the diet, and the age and physical state of the subject {1, 6, 8, 11). Pulmonary absorption depends on the size of lead particles and on the depth and rate of breathing (5, 8). Some large particles are deposited on the mucous lining of the respiratory tract, and some are ultimately swallowed (8, 10). Pulmonary retention is typically 40% (5, 8, 10). Absorbed lead enters the blood and is distributed to soft tissues and bone. After prolonged exposure an equilibrium is reached between the blood and soft tissues. In contrast, bone has the ability to accumulate lead with time. Postmortem studies indicate that the skeletal burden of lead increases with age; in fact, about 90 % of the total body burden lies • For children, the absorption seems to be higher; values up to children under 5 years of age (4. 8). 50~.
have been regarded as typical for
116
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in the bone (8, 11). The respective half-lives of lead in blood, soft tissues, and bone have been estimated to be 2-4 weeks (5, 8, 14), 4 weeks (14), and 27.5 years (10). Lead passes through the placenta easily and fetal blood has almost the same lead concentration as maternal blood. Lead also passes the bloodbrain barrier, although the brain does not accumulate lead (16). Relationships between lead intake and blood lead level have been extensively studied and reviewed (6). It appears that there is a curvilinear relationship between air lead and blood lead; as lead exposure increases the corresponding blood lead increments become smaller (17). The interrelationship between lead exposure from water and blood level is also curvilinear, and this has been demonstrated in human beings drinking household tap-water where lead was present in high concentrations, i.e., exceeding 50 J.tg/litre (18, 19). It should be stressed, however, that at the present time there is considerable uncertainty regarding the precise relationship between lowlevel exposure to lead and levels in the blood; consequently, great care needs to be exercised when using data derived from such relationships. Lead is excreted in urine, faeces, sweat, hair, fingernails, and toenails. The metabolism of lead has been reviewed in two WHO publications (8, 16).
10.4 Health effects Lead in high doses has been recognized for centuries as a cumulative general metabolic poison. Some of the symptoms of acute poisoning are tiredness, lassitude, slight abdominal discomfort, irritability, anaemia and, in the case of children, behavioural changes (8). Such symptoms are difficult to quantify and currently there is considerable interest in various possible subtle effects, including neurophysiological ones, caused perhaps by exposure to low levels of lead (5, 11). Lead at low levels can reduce the activity of an enzyme, porphobilinogen synthase (EC 4.2.1.24) (8). This enzyme is involved in normal haeme synthesis at the stage of conversion of aminolevulinic acid to porphobilinogen; a decrease in the activity of this enzyme may be used as an index of exposure to lead. Lead also has an affinity for amino acids containing sulfur. In addition, lead has a tendency to bind to mitochondria, leading to interference in the regulation of oxygen transport and energy generation (10). Significantly higher lead blood levels (> 400 J.tg/litre) have been found in mentally retarded children (20). Many animal studies have been carried out on the effects of lead on the haemopoietic, nervous, renal, cardiovascular, and reproductive systems (8, 10, 11). While the results of these studies cannot be directly extrapolated to man, they do provide valuable dose-response data that are not available from epidemiological human studies. A wide range of epidemiological and clinical studies have been
10.
LEAD
117
conducted on human subjects. including retrospective studies to ascertain the possible causes of mortality and morbidity in lead-exposed populations and studies on the effects on specific organs and systems (8). Of particular relevance are the studies of levels of lead in water in relation to levels of lead found in the blood of both adults and children, and possible associated subtle behavioural effects in children (6). Many of these studies have shown that there is a small increase in lead in the blood where the levels in water are relatively high, but in general the blood levels attributable to water are not high in relation to acceptable levels of blood lead for individuals and population groups (6, 8). For example, for adults a mean increase of 25 Jlg of lead per litre of blood has been estimated to be associated with drinking-water containing an average concentration of 100 Jlg of lead per litre (6). Increments of about 40 and 50 Jlg of lead per litre of blood have been estimated for young children and pregnant women, respectively (6, 10). The interpretation of such figures must be considered with care, however, owing to the curvilinear relationship between lead intake and lead in the blood. Population group mean exposure limits for lead from all environmental sources combined has been specified as 200 Jlg per litre of blood, but for individuals values are within the range 300-350 Jlg/litre. For individual children, the Centers for Disease Control in the USA and also the American Academy of Pediatrics (6, 10) have recommended an exposure limit value of 300 Jlg per litre of blood. Lower lead values may need to be considered in the future and 250-300 Jlg litre has been proposed for individuals (6, 10). Typical values for the lead levels in the general population are not known very accurately, but for nonoccupationally exposed adults figures of below 200 Jlg per litre of blood are often quoted. In the USA (6) it has been estimated that 99.5% of children would have < 300 Jlg of lead per litre of blood if the geometric mean level was maintained at 150 Jlg/litre. Many of the studies on behavioural effects in children relate to situations where there are high levels of lead in the environment generally rather than from the lead in water specifically. In a few instances there is a tenuous suggestion of some association between adverse effects and levels of lead in water (8, 10). In several studies, chromosomal aberrations were found in peripheral lymphocytes of lead-exposed populations whose blood lead levels ranged from 100 to 1000 Jlg/litre. Negative results were obtained in other studies in which blood lead levels ranged from 40 to 500 Jlg/litre (21). Lead is not known to be essential for the functioning of biological systems and the general view is that where possible the exposure to lead should be kept as low as possible. Although in 1972, an FAO/WHO Expert Committee recommended that the maximum intake of lead should be 3 mgjweek (0.05 mgjkg of body weight) for adults (12), no corresponding value was suggested for children. The situation will be different for children (including infants), because absorption of lead is
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higher than for adults (12); children also have a higher susceptibility, due in part to their rapid growth rate (4). Pregnant women and developing fetuses also appear to be more sensitive to lead because of increased maternal food intake and changes in hormonal status (8). Should the intake in the diet be greater than 220 Jl.g of lead per day (using the assumptions in the tables on page 114), the 3-mg weekly limit for adults would be exceeded. The health effects of lead have been reviewed by three WHO groups (8, 16, 22).
REFERENCES I. Guidelines for Canadian drinking water quality, 1978. Quebec, Ministry of Supply and Services, 1979 (supporting documentation). 2. HARRISON, R. M. & LAXEN, D. P. H. Natural source of tetra-alkyl lead in air. Nature, 275: 738 (1978). 3. The hazards to health and ecological effects of persistent substances in the env~ronment: report of a working group. Copenhagen, WHO Regional Office for Europe, 1973. 4. NATIONAL RESEARCH CouNCIL. Drmking water and health. Washington, DC, National Academy of Sciences, 1977. 5. Toxicology of water. vol. II. Washington, DC, US Environmental Protection Agency, 1977 (Environmental Health Effects Research Series). 6. US Environmental Protection Agency. Ambient water quality criteria for lead. Washington, DC, Criteria and Standards Division, Office of Water Planning and Standards. 1980 (EPA 440/5-80-057). 7. Lead m drmking water, a survey m Great Britain. London, Department of the Environment, 1977 (Pollution paper No. 12). 8. Lead. Geneva, World Health Organization, 1977 (Environmental Health Criteria 3). 9. ZOETEMAN, B. C. J. & BRINKMANN, F. J. J. In: Amavis, R. et al., ed., Hardness of drmkmg water and pubbc health, Oxford, Pergamon Press, 1975. 10. DRILL, S. ET AL. The environmental lead problem. An assessment of lead in drinking water from a multi-medw perspective. Washington, DC, US Environmental Protection Agency, 1979. II. UNDERWOOD, E. J. Trace elements th human and animal nutrition. New York, Academic Press, 1977. 12. WHO Technical Report Series, No. 505, 1972 (Evaluation of certain food additives and the contaminants: mercury, lead and cadmium). 13. CHAMBERLAIN, A. C. ET AL. Investigations into lead from motor vehicles. Harwell, Oxfordshire, Atomic Energy Research Establishment, Environmental and Medical Sciences Division, 1978 (AERE- R9198). 14. RABINOWITZ, M. ET AL. Studies of human lead metabolism by use of stable isotope tracers. Environmental health perspectives, 7: 145 (1974). 15. GARBER, B. T. & WEI, E. Influence of dietary factors on the gastrointestinal absorption of lead. Toxicology and applied pharmacology, 27: 685 (1974). 16. WHO Technical Report Series, No. 647, 1980 (Recommended health-based limits in occupational exposure to heavy metals). 17. HAMMOND, P. B. & BELILES, R. P. Metals: lead. In: Doull, J. et al., ed., Casarett and Doull's toxicology: the basic science of poisons, 2nd ed. New York, Macmillan, 1980. 18. MOORE, M. R. ET AL. Contribution of lead in water to blood-lead. Lancet, 2: 661 (1977). 19. THOMAS, H. F. ET AL. Relationship of blood lead in women and children to domestic water lead. Nature, 282: 712 (1979).
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119
20. MooRE, M. R. ET AL. A retrospective analysis of blood-lead in mentally retarded children. Lancet, 1: 717 (1977). 21. Some metals and metallic compounds. Lyon, International Agency for Research on Cancer, 1980 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 23). 22. Health hazards from drinking-water: report of a Working Group. Copenhagen, WHO Regional Office for Europe, 1977 (ICP/PPE 005).
11. MERCURY 11.1 General description 11.1.1 Sources The major source of mercury in the environment IS the natural degassing of the earth's crust, quantities ranging between 25 000 and 150 000 tonnes of mercury per year being released. In addition, many industrial activities not directly related to mercury production or use contribute significant amounts of this element to the environment; these include burning of fossil fuels, smelting of various metals, cement manufacture, and waste disposal. In addition, mercury is used in chloralkali plants (producing chlorine and sodium hydroxide), in paints as preservatives or pigments, in electrical switching equipment and batteries, in measuring and control equipment (e.g., thermometers, medical equipment), in dentistry, and in agriculture (especially as seed dressings). Mercury can exist in the environment as the metal, as monovalent and divalent salts, and as organomercurials, the most important of which is methyl mercury. Methyl mercury may be produced from inorganic mercury by microorganisms found in aquatic sediments and sewage sludge; other microorganisms can demethylate mercury back to inorganic mercury. Fish and mammals absorb and retain methyl mercury to a greater extent than inorganic mercury; it is methyl mercury that accumulates along food chains (1). 11.1.2 Occurrence in water Rainwater in Sweden is reported to contain approximately 300 ng of mercury per litre (1). In most surface-waters, mercuric hydroxide and chloride are the predominant mercury species; levels are generally less than 0.001 mg/litre (2-5). In polluted rivers and lakes, levels of up to 0.03 mg/litre have been reported (6). Inland waters in the Federal Republic of Germany contain mercury at concentrations of about 400 ngjlitre whereas values ranging between 100 and 1800 ng/litre were found in rivers (7). Levels of mercury in drinking-water are usually very low (8). For example, median levels in several Canadian provinces are approximately 0.0002 mg/litre (9). Inorganic mercury can be controlled in water treatment by iron and alum coagulation (10). Seven hundred water 120
11.
MERCURY
121
samples collected from the drinking-water reservoirs in the Federal Republic of Germany indicate that the purest drinking-water contained less than 0.00003 mg/litre (11). 11.2 Routes of exposure 11.2.1 Drinking-water
Where there is no evidence of mercury contamination, levels of mercury in freshwater bodies are less than 0.0002 mg/litre. Levels of mercury in ambient waters may be significantly reduced by conventional treatment processes and further losses may occur during the preparation of beverages. Therefore, the estimate that the intake of mercury from drinking-water would not normally exceed 0.1 Jlg daily (1) may be on the high side (4). 11.2.2 Food
Food is the main source of mercury in non-occupationally exposed populations and fish and fish products account for most of the methyl mercury in food (1). The average daily intake of mercury from food is in the range 10-12Jlg (1, 12, 13) but in regions where ambient waters have become contaminated with mercury and where fish comprises a high proportion of the diet, the intake from food may be much higher. 11.2.3 Air
Because metallic mercury and organomercurial compounds have relatively high vapour pressure, they can be released to the atmosphere by volatilization. Ambient air levels, except in polluted areas, appear to be of the order of 0.02 Jlg/m 3 . Assuming an ambient air level of 0.05 Jlg/m 3 , the average daily intake of metallic mercury vapour would amount to about 1 Jlg/day, of which about 80% is retained (1).11.2.4 Industrial exposure
A large number of occupations or trades involve exposure to mercury (14). Especially significant are the mining industry, the chloralkali
industry, and the manufacture of some scientific instruments; mercury levels in air may attain values as high as 5 mg/m 3 (1). Accepting the time-weighted average threshold limit value of 0.05 mg/m 3 proposed by the American Conference of Government Industrial Hygienists (ACGIH), the occupational exposure would lead to a daily average intake of 500 Jlg of mercury or less, assuming a pulmonary ventilation of 10m 3 /day at work (1).
122
III.
HEALTH-RELATED INORGANIC CONSTITUENTS
11.2.5 Relative significance of different routes of exposure
For adults the relative significance of mercury in drinking-water is illustrated in the following table:
Weekly intake (uptake) of mercury ( /19) Food (a) 70 (b) 140 (c) 70 (d) 140 Air Water Total Ratio: water/total ( %)
(5.6) (11.2) (5.6) (11.2)
2.8 2.8 7.0 7.0
(2.2) (2.2) (5.6) (5.6)
14 14 14 14
(2.1) (2.1) (2.1) (2.1)
87 157 91 161
(9.9) (15.5) (13.3) (18.9)
16 9 15 9
(21.0) (13.6) (15.8) (11.1)
AssumptiOns (1) Food mtake10 pg/day (a. c) or 20 pgfday (b,d) wnh 8% of mgested mercury absorbed (2) Atr contammg 0 02 pg/m' (•. b) or 0 05 pgfm' (c, d) wrth total ventrlatron of 20 Om 3 fday and 80% retent1on (3) Dnnktng-water contam1ng 0 001 mg/lltre With mtake of 2 lltres per day and 15% of Ingested mercury absorbed
No reliable information is available for children. 11.3 Metabolism
Mercury serves no beneficial physiological function in man. The various physical and chemical states (metal, inorganic compounds. organomercurial compounds) each have intrinsic properties that dictate independent toxicological assessment (14). Absorption of inorganic mercury compounds from food is about 7-8% of the ingested dose; in contrast, gastrointestinal absorption of methyl mercury is practically complete. Absorption of inorganic mercury compounds from water may be 15% or less {15) whereas the methyl mercury is almost completely absorbed. Inorganic mercury compounds are rapidly accumulated in the kidney, the main target organ for these compounds {15). Absorbed methyl mercury rapidly appears in the blood where, in man, 80-90% is bound to red cells; demethylation of methyl mercury to inorganic mercury occurs at a slow but significant rate. The greater intrinsic toxicity of methyl mercury compared with inorganic mercury is due to its lipid solubility, which permits it to cross biological membranes more easily than inorganic mercury, especially into brain, spinal cord, and peripheral nerves, and across the placenta. Mercury salts are excreted from the kidney, liver, intestinal mucosa, sweat glands, salivary glands, and through milk; the most important routes are via the urine and faeces {15). 11.4 Health effects
The major effects of mercury poisoning take the form of neurological and renal disturbances, which are primarily associated with organic and
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MERCURY
123
inorganic mercury compounds respectively (16). According to Krasovsky (17), in addition to producing general toxic effects, mercury causes gonadotoxic and mutagenic effects and disturbs the cholesterol metabolism. The toxicology of mercury compounds has been reviewed in depth in the WHO Environmental health criteria series (J). It was not possible to identify even approximate minimum effect values for inorganic, aryl, and alkoxyalkylmercurials. There exists no evidence that inorganic mercury is carcinogenic. Alkylmercurials are embryotoxic and teratogenic in laboratory animals (9).
REFERENCES I. Mercury. Geneva, World Health Organization, 1976 (Environmental Health Criteria 1).
2. HoLDEN, A. V. Present levels of mercury in man and his environment. In: Mercury 3. 4. 5.
6.
7.
8. 9. 10. II.
12. 13. 14.
15. 16. 17.
contamination in man and his environment, Vienna, International Atomic Energy Agency, 1972, p. 143 (Technical Report Series No. 137). WIKLANDER, L. Mercury in ground and river water. Grundfoerbaettring, 21: 151 (1968). WERSHAW, R. L. Sources and behaviour of mercury in surface water. In: Mercury in the environment, Washington, DC, US Geological Survey, 1970 (Professional Paper No. 713). VoEGE, F. A. Levels of mercury contamination in water and its boundaries. In: Proceedings of the symposium on mercury in man's environment, Ottawa, Royal Society of Canada, 1971, p. 107. Investigations of mercury in the St. Clair River- Lake Erie systems. Washmgton, DC, US Department of the Interior, 1970 (Report of the Federal Water Quality Administration), p. 108. ScHRAMEL P. ET AL. Some determinations of Hg, As, Se, Sb, Sn and Br in water, plants, sediments and fishes in Bavarian rivers. Internatwnal journal of enVIronmental 5tudies, 5: 37 (1973). NATIONAL RESEARCH COUNCIL. Drinking water and health. Washington, DC, National Academy of Sciences, 1977. Guidelines for Canadian drinking water quality, I978. Quebec, Ministry of Supply and Services, 1979 (supporting documentation). Gu1dance for the issuance of variances and exemptions. Washington, DC, Office of Drinking Water, US Environmental Protection Agency, 1979. BouQUIAUX, J. In: Proceedings of an international symposium on the problems of contammatwn of man and his environment by mercury and cadm1um. Luxembourg, Commission of the European Commumties, 1974, p. 23. MERANGER, J. C. & SMITH, D. C. The heavy metal content of a typical Canadian diet. Canadwn journal of public health, 63: 53 (1972). NEILSEN-KuosK, F. Absorption of mercury vapour from the respiratory tract in man. Acta pharmacologica et toxicolog1ca, 23: 250 (1965). KEY, M. M. ET Al., ed. Occupational diseases-a guide to their recognition. Washington, DC, US Department of Health, Education and Welfare, 1977 (NIOSH publicatiOn 77-181), pp. 370-373. WHO Technical Report Senes, No. 647, 1980 (Recommended health-based limits m occupational exposure to heavy metals). SWEDISH ExPERT GROUP. Methyl mercury in fish. A toxicologic-epidemiologic evaluation of risks. Nordisk hygienisk Tidskrift, Suppl. 4 (1971). KRASOVSKY, G. N. ET AL. [The need for revising the existing hygienic standard for mercury in water.) Gigiena i sanitarija (2): 20 (1981) (in Russian).
12. NICKEL 12.1 General description 12.1.1 Sources Nickel is ubiquitous; typical soils contain between 10 and IOO mg of nickel per kg (1). The chief ores are mainly arsenides and sulfides. The processing of minerals as well as the production and use of nickel has caused environmental contamination. Nickel is used as a component in some alloys and for metal plating, for catalysts, for batteries, and in certain fungicides; the use of the metal in food processing equipment can give rise to some contamination of food. 12.1.2 Occurrence in water Many nickel salts are water-soluble, therefore contamination of water can arise; significant problems are associated with industrial discharge to rivers of effluents containing nickel compounds. Levels as high as I mg/litre have been reported in surface-waters (1), although the levels are generally much lower, e.g., 5-20 Jlg/litre (2). In the USSR, underground water supplies have been found to contain nickel at levels up to 0.13 mgjlitre (3). 12.2 Routes of exposure 12.2.1 Drinking-water Some nickel is removed by conventional water treatment, so that the levels in treated water are generally lower than in untreated water (2). Few comprehensive surveys of nickel levels in tap-water have been identified. Limited data suggest that concentrations of 2-5 Jlg/litre are fairly typical (1-5); elevated levels may sometimes be found (6), especially where nickel-plated plumbing fittings are used. Exceptionally, levels in drinking-water up to 0.5 mg/litre have been reported (4, 7). Assuming a consumption of 2 litres of water per day, human exposure to nickel from this source would not normally exceed I0-20 Jlg/day. 12.2.2 Food Nickel is present in most foodstuffs, but at levels below (and often well below) I mgjkg (8). Little is known about the chemical form of 124
12.
NICKEL
125
nickel in food, although it is probably partly complexed with phytic acid (8). Dietary contributions have been reported ranging from less than 200
to 900 Jlgfday (1, 2, 8- 10). A typical diet might contribute about 400 Jlg/day. Nickel concentrations of 100 Jlgflitre and 50 Jlg/litre have been reported in wines and beers respectively (J). 12.2.3 Air Few data have been reported concerning the nickel content of air. However, it would appear that levels in air are generally less than 0.5 Jlg/m 3 . Higher levels were reported in the past, mainly associated with industrial areas (1, 2). A typical urban air concentration is 0.2Jlg/m 3 . 12.2.4 Other routes of exposure 12.2.4.1 Industrial exposure Levels up to 400 Jlg of nickel per m 3 of air have been reported in the industrial environment (J), although generally exposure levels in industry are much lower. In some cases the major route of exposure of a worker could be his industrial environment. 12.2.4.2 Smoking It has been reported that about 10-20% of the nickel content of cigarettes (typically around 3 Jlg per cigarette) can be inhaled (I, 2). This appears to be mainly as a volatile nickel compound, nickel carbonyl (2). A typical weekly intake for someone who smokes 20 cigarettes a day might be 40-80 Jlg of nickel.
12.2.5 Relative significance of different routes of exposure Estimates of weekly human exposure to nickel for adults are given below: (a) Weekly intake from food, air, and water Nickel concentration in water Weekly intake of nickel (JlQ) Water only Air only Food only Total Ratio: water/total (%)
50 JlQ/1 itre 7 5 JlQ/1 itre 100 JlQ/Iitre 150 JlQ/Iitre
700 1050 1400 2100
28 28 28 28
3150 3150 3150 3150
3878 4228 4578 5278
18.0 25.0 30.6 40.0
Assumpttons Da1ly mckel mtake from food. 450
JJ.Q, vent1lat1on 20m3 of a1r dally, mckel content. 0 2 1J.Q/m3
126
Ill.
HEALTH-RELATED INORGANIC CONSTITUENTS
(b) Weekly absorption from food, air, and water Nickel concentration in water Weekly absorption of n1ckel (Jig) Water only A1r only Food only Total Rat1o: water;total (%)
50 JIQ/htre 75 JIQ/Iitre 100 JIQ/Iitre 150 JIQ/Iitre
7.0 10.5 14.0 21.0
14 14 14 14
31.5 31.5 31.5 31.5
52.5 56.0 59.5 66.5
13 3 18 8 23.5 31.6
Assumptions Inhaled ntckel 50% absorptton, mgested mckel 1 % absorptton
12.3 Metabolism Nickel is almost certainly essential for animal nutntwn, and consequentially it is probably essential to man (1 1). Absorption of nickel through the gastrointestinal tract seems to be very low, i.e., I '1o or even less {1), although higher absorption values have been reported, i.e., 10% (8). There is little evidence of accumulation of nickel by various tissues (1 2). No significant accumulation of nickel was observed in rats fed nickel in drinking-water at a concentration of 5 mgjlitre. It is clear that at least in the animal body a mechanism controls excessive intake of nickel. Certain disease states in man do give rise to elevated nickel in tissue; the reasons, however, are not understood (8). Nickel is readily excreted, mainly in the faeces, with smaller quantities in the urine. Significant amounts can also be discharged in sweat (8). 12.4 Health effects Nickel is a relatively nontoxic element. The levels of nickel usually found in food and water are not considered a serious health hazard (1, 8); however, high doses (1600 mgjkg in the diet) were shown in early animal studies to cause minimal toxic effects (decreased number of mice pups weaned) (8). Such effects were not substantiated in later threegeneration reproduction studies. Rats and mice were given drinkingwater containing 5 mg of nickel per litre throughout their lifetime without adverse effects (1 3). Certain nickel compounds have been shown to be carcinogenic in animal experiments (1, 6). However, soluble nickel compounds are not currently regarded as either human or animal carcinogens (2). As in the case of other divalent cations, nickel can react with DNA, and at high concentrations can result in DNA damage as shown in in vitro mutagenicity tests (IARC, personal communication). Dermatitis is most commonly associated with industrial exposure. However, the same effects have been observed from dermal contact with coinage or jewellery. High-level occupational exposures have been associated with renal problems, and effects such as vertigo and dyspnoea have been observed (1).
12.
NICKEL
127
Studies using large numbers of patients have investigated the role of contact dermatitis in eczema of the hands; between 4% and 9% of the patients were found to respond positively to nickel patch tests (14). Women appear to be more sensitive to nickel than men by a factor of ten (15). These studies are of limited value in that they examined eczema patients and cannot fully reflect the true incidence of sensiti~ation or contact dermatitis in the general population (16).
REFERENCES I. COMMISSION OF THE EUROPEAN COMMUNITIES. Trace metals. exposure and health effects. Oxford, Pergamon Press, 1979. 2. NATIONAL RESEARCH COUNCIL. Drinking water and health. Washington, DC, National Academy of Sciences, 1977. 3. SIDORENKO, G. I. & ITSKOVA, A. I. Nickel. Moscow, Medicina. 1980. 4. DURFOR, C. N. & BECKER, E. Publzc water supplies of the 100 largest cities in the United States, 1962. Washington, DC, Government Printing Office, 1964 (Geological Survey Water Supply Paper 1812). 5. KOPP, J. F. & KRONER, R. C. Trace metals m waters of the United States. Cincinnati, US Department of the Interior, 1967.
6. Cadmium, nickel, some epoxides, miscellaneous industrial chemicals and general considerations on volatile anaesthetics. Lyon, International Agency for Research on Cancer, 1976 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. II). 7. KoPP, J. F. The occurrence of trace elements in water. In: Hemphill, D. P., ed.
8. 9.
10. II.
Proceedings of the Third Annual Conference on Trade Substances in Environmental Health, 1969, Columbia, University of Missouri, 1970, pp. 59-73. UNDERWOOD, E. 1. Trace elements in human and animal nutrition. New York, Academic Press, 1977. HAMILTON, E. I. & MINSKI, M. J. Abundance of the chemical elements in man's d1et and possible relations with environmental factors. Science of the total environment, 1: 375 (1973). MASIRONI, R. How trace elements in water contribute to health. WHO chronicle, 32: 382 (1978). WHO Technical Report Series, No. 532, 1973 (Trace elements in human nutrition: report of a WHO Expert Committee).
12. ENVIRONMENTAL PROTECTION AGENCY. Water quality critena; availability. Federal register, 44: 43684 (1979). 13. SCHROEDER, H. A. ET AL. Long-term effects of nickel in rats: survival, tumors, interactions with trace elements and tissue levels. Journal of nutrition, 104: 239 (1974). 14. WILKINSON, D. S. ET AL. The role of contact allergy in hand eczema. Transactions of the St. John's Hospital Dermatological Soc1ety, 56: 19-25 (1970). 15. FISHER, A. A. & SHAPIRO, A. Allergic eczematous contact dermatitis due to metallic nickel. Journal of the American Medical Association, 161: 717-721 (1956). 16. KAALTER, K. ET AL. Low mckel d1et in the treatment of patients with chrome nickel dermatitis. British journal of dermatology, 98: 197-201 (1978).
13. NITRATE AND NITRITE Nitrate and mtnte are considered together because conversion from one form to the other occurs in the environment. The health effects of nitrate are generally a consequence of its ready conversion to nitrite in the body. Concentrations in water are expressed as mg/litre for nitrate-nitrogen (nitrate-N) and nitrite-nitrogen (nitrite-N). 13.1 General description 13.1.1 Sources
Nitrates are widely present in substantial quantities in soil, in most waters, and in plants, including vegetables (1). Nitrites also occur fairly widely, but generally at very much lower levels than nitrates (1). Nitrates are products of oxidation of organic nitrogen by the bacteria present in soils and in water where sufficient oxygen is present. Nitrites are formed by incomplete bacterial oxidation of organic nitrogen (2). One of the principal uses of nitrate is as a fertilizer; most other nitrogen-containing fertilizers will, however, be converted to nitrate in the soil (2). Nitrates are also used in explosives, as oxidizing agents in the chemical industry, and as food preservatives (2). The main use of nitrites is as food preservatives, generally as the sodium or the potassium salt (1). Some nitrates in the environment are produced in the soil by fixation of atmospheric nitrogen (bacterial synthesis). Some nitrates and nitrites are formed when oxides of nitrogen produced by the action of lightning discharge or via man-made sources are washed out by rain (2). Nitrates and some nitrites are also produced in the soil as a result of bacterial decomposition of organic material, both vegetable and animal. Because nitrates and nitrites are widespread in the environment, they are found in most foods, in the atmosphere, and in many water sources. 13.1.2 Occurrence in water
Fertilizer use, decayed vegetable and animal matter, domestic effluents, sewage sludge disposal to land, industrial discharges, leachates from refuse dumps, and atmospheric washout all contribute to these ions in water sources (1, 3, 4). Changes in land use may also give rise to increased nitrate levels. Depending on the situation, these sources can contaminate streams, rivers, lakes, and groundwater, especially wells (3). 128
13.
NITRATE AND NITRITE
129
Contamination may result from a direct or indirect discharge, or it may arise by percolation over a period of time, sometimes after many years. The levels of nitrates in polluted water are almost invariably very much higher than the levels of nitrites (4). Levels of nitrate in water are typically below 5 mg of nitrate-N per litre, but levels exceeding 10 mgjlitre occur in some small water sources. In chlorinated supplies, levels of nitrite are often less than the limit of detection, i.e., < 0.005 mg of nitrite-N per litre (2-4), but relatively high levels may occur in unchlorinated water. Very high nitrite levels are usually associated with water of unsatisfactory microbiological quality. A number of studies have revealed levels of nitrate in the range of 20 to over 200 mg of nitrate-N per litre (3), but this is rare. Most of the higher levels of nitrate are found in groundwater (3); nitrates in surfacewaters tend to get depleted by aquatic plants (4). Increases in the levels of nitrates in water are associated with the application of nitrogen fertilizers. Marked seasonal variations can occur in concentration in rivers and high levels may occur, particularly after heavy rainfall following severe drought periods (5). The levels in groundwater tend to be much more steady during the year.
13.2 Routes of exposure 13.2.1 Drinking-water Because none of the conventional water treatment and disinfection practices modify the levels of nitrate to any appreciable extent, and since nitrate concentration is not changed markedly in water distribution systems, the levels in tap-water are often very similar to those for source waters. For nitrite, the levels in tap-water are likely to be markedly lower than in source waters because of oxidation during water treatment, particularly when water is chlorinated (2). It is very difficult to define a range and average exposure to nitrate or nitrite in water because the concentrations vary widely depending on the water source. For the majority of the world's population the exposure is likely to be well below 5 mg of nitrate-N per litre (3). For small populations, often in remote areas, the levels may be up to 100 mg or more of nitratenitrogen per litre. Assuming a consumption of 2 litres of water per day, the exposure is typically under 20 mg of nitrate-N per day, but in rare circumstances it could be over five times this value (3).
13.2.2 Food Considerable quantities of nitrates and lesser amounts of nitrites are present in certain foods and in general the major source of human intake of both nitrates and nitrites is food (4). In certain crops the nitrate levels may be as high as 100 mgjkg. Certain vegetables, including cabbage, celery, lettuce, potatoes, several root vegetables, and spinach, contain
130
Ill.
HEALTH-RELATED INORGANIC CONSTITUENTS
relatively high levels of nitrate, but only small quantities of nitrite (4). Nitrates and nitrites are added as preservatives to certain foods, particularly for certain meats and cheeses (2). A further important source of ingested nitrates and nitrites is saliva (4); human beings secrete about 10 mg of nitrate-N per day, of which about 2 mgfday is reduced to nitrite (4); the nitrate in the saliva is essentially derived from food sources (6), mainly vegetables. The variation in the quantities of nitrates and nitrites ingested from the diet is extremely high (4). For example, individuals who eat few vegetables and little cured meat will have a very much lower intake (4). Various typical dietary inputs have been estimated as ranging from about 120 mg to 230-300 mg of nitrate per day (no estimate for nitrites) (4, 7, 8). As infants are the most sensitive group of the population to nitrate, it is important to define their exposure; a daily intake from food alone for a two-month-old infant has been estimated to be approximately 25 mg of nitrate-N (2).
13.2.3 Air Apart from natural sources of oxides of nitrogen and nitrates in the air, there are some important man-made sources, in particular the products of combustion of fossil fuels (coal, oil, gas) and from the chemical industry. Inorganic and organic forms of nitrates occur; these and the oxides of nitrogen will be inhaled and some will be absorbed by the respiratory system, producing a mixture of nitrates and nitrites in the body. It has been estimated, for areas with elevated levels of nitrogen compounds in the air, that if all these compounds were absorbed by an adult, it would amount to an intake of approximately 0.1 mg of nitrateN per day.
13.2.4 Other routes of exposure 13.2.4.1 Industrial exposure Oxides of nitrogen are fairly common industrially, and although aerosols of nitrates and nitrites may occur, their levels would generally be low. A maximum of 5 parts of nitrogen dioxide in 10 6 parts of air is permitted in the USA in industrial situations for an eight-hour shift (9); this could represent an exposure of up to about 30 mg of nitrate-N per day in extreme circumstances. 13.2.4.2 Smoking Oxides of nitrogen and nitrate aerosols are produced in a burning cigarette, but in comparison with food and water, the quantities contributed by smoking tobacco are regarded as relatively small (2).
13.
NITRATE AND NITRITE
131
13.2.5 Relative significance of different routes of exposure
For different individuals, there can be a wide range of possible exposures to nitrates and nitrites in water, food, and air. General air pollution appears to be a relatively unimportant source, making food, saliva, and water the main normal sources (2). Two levels of food nitrate are considered and, ignoring contributions from saliva, the proportions of nitrate from water containing 10 mg of nitrate-N per litre have been calculated. The estimates are based on 100% absorption from food and water, with a consumption of 2 litres of water per day for adults and I litre of water per day for children. Negligible intakes from inhalation, cigarette smoking, occupational exposure, and other sources are assumed. Weekly uptake of nitrate-N (mg) Water only• Food only Rat10: water ;total (%)
Daily intake of nitrate-N m food (mg) Adults-case 1· Adults-case 2: Children-case 3: a 10 mg nltrate-N per litre
Total
20 70 25
140 140 70
140 490 175
280 630 245
50 22 28
13.3 Metabolism
The metabolism of ingested nitrate is not fully understood; it seems that absorption takes place in the upper portion of the small intestine and that excretion is primarily, if not exclusively, through the kidney (4). It is well known that nitrate is absorbed in the upper gastrointestinal tract and concentrated ultimately into saliva by the salivary glands (4). The nitrate metabolism in man has not been studied fully, and the results of animal experiments are not very reliable when extrapolated to man (4). Both nitrates and nitrites are very readily absorbed by the body. A very important consideration is the fact that nitrate can be readily converted in vitJo to nitrite as a result of bacterial reduction (4). Exposure to high levels of nitrate has been shown to give rise to large increases in the concentration of salivary nitrite (4). However, there can be marked variations between individuals; this can be a function of differences in their oral microflora and various constituents in their diet (4). Reduction of nitrate to nitrite also occurs elsewhere in the body, including the stomach; little conversion takes place unless the pH is greater than 4.6 (4). In infants, where the stomach acidity is normally very low, about pH 4 or higher (3, 4), a high yield of nitrite is obtained. In contrast, the acidity of an adult stomach is pH I- 5, and less conversion of nitrate occurs (4).
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HEALTH-RELATED INORGANIC CONSTITUENTS
The formation of nitrite is especially important for two reasons. Firstly, it can oxidize haemoglobin to methaemoglobin, a pigment that is incapable of acting as a carrier of oxygen. Secondly, under certain conditions, nitrites may react in the human body with secondary and tertiary amines and amides (commonly derived from food and other sources) to form nitrosamines, some of which are considered to be carcinogenic (1). This process occurs in acidic solution within the pH range 1-5 (4) characteristic of the normal acidity range in a human stomach. The reaction rate is greatest at pH 3.5 or less.
13.4 Health effects 13.4.1 Methaemoglobinaemia Normally, 1-2% of the body's haemoglobin is in the methaemoglobin form, but when the proportion is in excess of 10% (3), clinical effects are detectable (methaemoglobinaemia); 30-40% leads to anoxia. It has been well documented that, in some countries, water supplies containing high levels of nitrate have been responsible for cases of infantile methaemoglobinaemia and death (2, 10). The extent of the worldwide problem has been reviewed in a WHO document.a It has been recommended that water supplies containing high levels of nitrate (> 100 mg of N0 3 per litre) should not be used for the preparation of infant foods; alternative supplies having a low nitrate content, even to the extent of using bottled water, have been recommended (5). The susceptibility of infants to nitrate has been attributed to their high intake relative to body weight (3), to the presence of nitrate-reducing bacteria in the upper gastrointestinal tract, and to the greater ease of oxidation of fetal haemoglobin (present in this form for the first few months of life) (11). The problem of methaemoglobinaemia does not arise in adults. Increased sensitivity may also occur when infants suffer from gastrointestinal disturbances, which increase the numbers of bacteria that can convert nitrate to nitrite (3, 12). The reconstitution of powdered milk, as opposed to other forms of milk, has also been regarded as increasing the sensitivity to the nitrate content in water. The stomach pH in infants, being about neutral, enables bacterial growth to occur in both the stomach and upper intestine. Infants, in contrast to adults, are also deficient in two specific enzymes that can convert methaemoglobin back to haemoglobin (3). The most common cause of infantile methaemoglobinaemia is excessive levels of nitrate in water used for the reconstitution of baby food (3). Prolonged boiling of water may exacerbate the problem by increasing the nitrate levels owing to evaporation. The vast majority of a Cyanosts of mfants produced by htgh mtrate concentrallon m rural wells. WHO Expert Committee on Maternal and Child Health, 1949 (Unpublished document WHO/MCH/13 19)
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133
cases of infantile methaemoglobinaemia have been associated with the use as a source of water of private wells that were microbiologically contaminated (3). A large number of studies have been carried out on the levels of nitrate in water giving rise to methaemoglobinaemia, but there are conflicting conclusions regarding the threshold level for an effect (3, 13). Cases of infantile methaemoglobinaemia have not been reported in areas where the drinking-water consistently contains less than 10 mg of nitrate-N per litre (2). Many infants have consumed much higher levels than this without developing the disease (2); only 2.3% of all cases appear to be associated with nitrate levels of between 10 and 20 mg of nitrate-N per litre of water {1). There is therefore some doubt about the effects at these concentrations. However, although clinical manifestations of infantile methaemoglobinaemia may not be apparent at these levels, undesirable increases in levels of methaemoglobin in the blood do occur (2).
There is also a suggestion that pregnant women are at greater risk than the general adult population {1), but further work is needed to confirm this. Although methaemoglobinaemia is well recognized and is unlikely to be a problem in areas with adequate medical facilities, it may be more important in the developing areas where such facilities are lacking.
13.4.2 Carcinogenicity of nitrosamines
Since ingested nitrates can be readily converted to mtntes, either in the mouth or elsewhere in the body where the acidity is relatively low (high pH), it is possible that nitrosamines, some of which may be carcinogenic, will be produced. It has been shown that the formation of nitrosamines may be increased in individuals with bladder infections and people suffering from achlorohydria (a condition of low stomach acidity) (14). In the case of bladder infections, it is probable that the nitrosamines produced there would be absorbed into the blood (2). Although tests on animals have shown that a number of nitrosamines are carcinogenic, there is no direct evidence of their carcinogenicity in man (1, 4, 15). There have been several such studies, but the overall evidence that nitrate in water might relate to cancer remains inconclusive (4, 16). Evidence of carcinogenicity from nitrate via the formation of nitrosamines rests with epidemiological studies, there being no appropriate animal studies that relate to humans. In a review of gastric cancer in China {17), the Putian Prefecture of the Fujian Province was found to have the highest mortality from this disease (120-147 per 100 000 males). Data showed that in this area the levels of both nitrate and nitrite in drinking-water and in vegetables were higher than in the low-risk areas. A study of the epidemiology and etiology of the disease is proceeding in both high- and low-risk areas.
134
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REFERENCES I. Nitrates, nitrites and N-nitroso compounds. Geneva, World Health Organization, 1978 (Environmental Health Criteria 5). 2. Guidelines for Canadian drinking water quality, 1978. Quebec, Ministry of Supply and Services, 1980 (supporting documentation). 3. Nitrates in water supplies. Report by the International Standing Comm1ttee on Water Quality and Treatment. Aqua, 1: 5-24 (1974). 4. NATIONAL RESEARCH COUNCIL. Drmking water and health. Washington, DC, National Academy of Sciences, 1977. 5. Royal Commission on Environmental Pollution. London, HM Stationery Office, 1979, Chapter 4. 6. TANNENBAUM, S. R. ET AL. Nitnte in human saliva. Its possible relationship to nitrosamine formation. Journal of the National Cancer Inslltute, 53: 79 (1974). 7. WHITE, J. W. Relative sigmficance of dietary sources of mtrate and mtrite. Journal of agricultural and food chemtstry, 23: 886 (1975). 8. PHILLIPS, W. E. J. Change in nitrate and nitrite content of fresh and processed spinach during storage. Journal of agricultural and food chemistry, 16: 88 (1968). 9. TLVs-Threshold limit values for chemical substances and physical agents in the workroom environment with intended changes for 1976. Cincinnati, American Conference of Governmental Industrial Hyg1emsts, 1976. 10. Health effects of nitrates m water. Cincinnati, US Environmental ProtectiOn Agency, 1977 (EPA-600;1-77-030). II. BETKE, K. ET AL. Vergleichende Untersuchungen iiber die Spontanoxydation von Naberschnur- und Erwachsenenhiimoglobin. Zeitschrift fur Kinderheilkunde, 77: 549 (1956). 12. SHUVAL, H. I. & GRUENER, N. Epidemiological and toxicological aspects of mtrates and nitrites in the environment. American journal of public health, 62: I 045 ( 1972) 13. FRAZER, P. & CHILVERS, C. Health aspects of mtrate m drinking water. Netherlands, 1980 (paper presented at the Internatwnal Symposium on Water Supply and Health). 14. HILL, M. J. ET AL. Bacteria, nitrosamines and cancer of the stomach. British journal of cancer, 28: 562 (1973). 15. Some N-mtroso compounds. Lyon, International Agency for Research on Cancer, 1978 (IARC Monographs on the evaluatwn of the carcinogenic risk of chemicals to humans, Vol. 17). 16. FRAZER, P. ET AL. Nitrate and human cancer: a review of the evidence. International journal of epidemiology, 9: 3 (1980). 17. Xu GUANG-WEI. Gastric cancer in Chma: a rev1ew. Journal of the Royal Society of Medicme, 74: 210 (1981).
14. SELENIUM 14.1 General description 14.1.1 Sources As a result of geochemical differences, levels of selenium in soil and vegetation vary within broad limits (1). The chemical form of selenium, and thus its solubility, is another decisive factor both as regards its entry into the food chain and its presence in water. Environmental processes can decrease selenium solubility, converting soluble selenate compounds, or some selenites, into compounds of very low solubility, such as elemental selenium or selenides (or even selenites of certain metals) (2). Selenium is usually present in water as selenite or selenate, the chemical form being influenced by such factors as pH and by the presence of salts of certain metals such as iron (2). 14.1.2 Occurrence in water Several reviews (1-7) of data from different parts of the world indicated that the selenium content in most surface-water samples analysed was well below I 0 pg/litre. Only 2 of 535 samples from the major watersheds of the USA studied over a four-year period exceeded this value, 14pg/litre being the maximum level found. Non-seleniferous regions of the USSR reported values ranging in general from a few tenths to several pg per litre, the maximum reported value being 5.1 pg/litre. The selenium level in 22 surface-waters in Argentina ranged from less than 2 to 19 pg/litre, with a median value of 3 pgjlitre. One study analysing 42 samples of surface-water from Colorado, USA, reported values ranging from less than 1 pgjlitre to 400 pgjlitre, with a median value of lpgjlitre. Waters draining in the USSR Ural Mountain areas nearest to pyrite deposits were also found to contain similar high values. Irrigation drainage from seleniferous soils increases selenium levels in surface (river) waters. Water from some springs and shallow wells contains selenium at levels exceeding 100 pg/litre; levels as high as 330 pg/litre have been reported from some wells in a seleniferous area of South Dakota, USA (2, 5). 14.2 Routes of exposure With the exception of occupational exposure, where the exposure through air and dermal contact is of particular significance, the general population is exposed to selenium mainly through food. 135
136
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HEALTH-RELATED INORGANIC CONSTITUENTS
14.2.1 Drinking-water
The level of selenium in tap-water samples from various public water supply systems in Canada and the USA, and some village supplies in Australia and the Federal Republic of Germany, only exceptionally exceeded 0.01 mgjlitre (2-8). Higher selenium levels are found in seleniferous areas, particularly in well-water. Systematic studies on selenium in drinking-water in high selenium areas of the world are scarce. Thus, in countries like the USA or Canada, the contribution made by selenium in drinking-water would usually not exceed 5-l 0% of the daily dietary intake and would generally, in most localities, be much lower than this value. Drinking-water in general does not represent the only or the main source of selenium exposure for the resident population in seleniferous areas. 14.2.2 Food
Dietary intake of selenium depends on food consumption patterns and selenium levels in foodstuffs, the latter being determined mainly by the character of the foodstuff and by geochemical conditions. Vegetables and fruits generally represent a poor dietary source of selenium, in contrast to grain, grain products, meat (particularly internal organ meat), and seafood, which contain substantial selenium levels, usually well above 0.2 mgjkg on a wet weight basis. The chemical composition of the soil and its selenium content have a marked influence on the selenium content in grain from different countries, ranging from 0.04 mg/kg to 21 mgjkg (4, 6-9). Recently reviewed data from studies on daily dietary intakes of selenium in different countries range from 56 Jlg/day in New Zealand (low selenium region) to over 320 Jlg/day in Venezuela, a com:itry with very high selenium levels in soils and vegetation (7, 9). On the other hand, daily intakes of selenium as low as 20 Jlg/day have been recorded in some apparently healthy individuals in New Zealand (10). Nutritional surveys indicate that, in countries such as the USA or Canada, "typical" diets would provide about 100-200 Jlg of selenium per day in the adult population (7, 9). 14.2.3 Air
Available data on selenium levels in the ambient air and in tobacco indicate that respiratory exposure does not contribute significantly to the daily intake of selenium in the general population (4, 7). 14.3 Metabolism
Soluble selenium salts, such as sodium selenite, are readily absorbed in the gastrointestinal tract of rats. Absorption exceeded 95 ',l'o whether the
14.
SELENIUM
137
diet contained 20 JJ.g or 4000 JJ.g of selenium per kg (J 1). About 93 % of selenium was absorbed by man when milligram doses of sodium selenite were administered in aqueous solutions. Thus people, like rats, exhibit no homoeostatic control limiting gastrointestinal absorption of large amounts of selenite (12). Absorbed selenium is widely distributed in organs and tissues, with high levels present in the liver and kidneys. Selenium penetrates through the placenta and also into the milk, the extent depending on the chemical form (4, 6, 13). Within the body, two primary metabolic pathways predominate. One is direct incorporation into or binding by proteins. The other, reduction followed by methylation, is responsible for the production of dimethylselenide and trimethylselenonium ions. When its rate of formation exceeds the rate of further methylation to a urinary metabolite trimethylselenonium ion, the volatile dimethylselenide is exhaled. Under the conditions of exposure prevailing in the general population, urinary selenium excretion predominates (3, 13). The rate of selenium elimination depends on the chemical form in which selenium is administered and on the selenium nutritional status. Available human data indicate that selenium administered as selenite is excreted more rapidly from the body than when given in organic form, such as selenomethionine (14). In rats, the biological half-time of selenium decreased with increased dietary selenium levels (15).
14.4 Health effects and dose-response relationships
Selenium has been identified as an essential nutrient in several animal species (8, 16). Certain endemic diseases of farm animals have been identified in areas with low selenium levels and selenium supplementation has been highly effective in preventing these diseases. Higher selenium doses are, of course, toxic, resulting in other diseases in farm animals (J, 2). 14.4.1 Studies on human populations
There is growing evidence that selenium is essential for human health. Manifest effects of selenium deficiency appear only under extreme conditions of long-term exposure to locally produced diets having extremely low selenium levels. Recent studies on Keshan disease suggest that this myocardial disease of children (17-19) could be induced by low-level selenium intake. Smith et al. (20) and Smith & Westfall (21) studied a group of farmers living in seleniferous regions of the USA who were consuming essentially locally produced foodstuffs and were exposed to selenium levels as high as 200 JJ.g/kg of body weight per day. Signs and symptoms observed were rather unspecific or vague. Nevertheless, in a group of 100 subjects exposed to high selenium intake, gastrointestinal disturbances, icteric
138
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HEALTH-RELATED INORGANIC CONSTITUENTS
discoloration of the skin, and bad teeth were observed in 31, 28, and 27 subjects, respectively. Another study was conducted by Jaffe in Venezuela, a country with high selenium levels in the soil and vegetation (22, 23). In this study, children with very high selenium exposure levels were compared with a group of children from Caracas, where the blood selenium levels and urinary selenium excretion were lower. The blood selenium levels found in the first group of children were the highest reported so far in the general population worldwide (8). In this study, Jaffe recognized that the two Venezuelan groups differed not only in selenium intake, but in other variables, including nutritional status and parasite infestation. Nausea, dermatitis, and pathological changes in the nails were more frequent among the children in the high seleniferous area than in the group from Caracas. However, observed changes, in particular growth retardation and anaemia, may have been due to other factors. Several studies have attempted to relate high dental caries prevalence in different population segments to high selenium exposure or, on the other hand, lower cancer incidence in certain areas to higher exposure levels to selenium (for review see 4, 24). However, these studies have not excluded the involvement of other variables and have been criticized also from other points, particularly as regards selenium exposure (25, 26).
14.4.2 Animal studies
In most animal species studied, basic dietary requirements approximate 0.04-0.10 mgjkg of food (8). However, vitamin E deficiency substantially increases the demand for selenium, and several other nutritional interactions need to be considered as well (4, 13). Dietary selenium levels of 5 mgjkg of food or more may cause chronic intoxication, and in seleniferous areas this value has been considered as the dividing line between toxic and non-toxic feeds (4). This conclusion is based both on field experience with farm animals raised in seleniferous areas and on many available experimental animal studies. The main effects of excessive selenium intake in animals include reduced body growth, decreased survival, and damage to the liver and other organs; in some cases, there has also been damage to the myocardium, kidneys, and pancreas. Some animal studies reported effects following long-term exposure to dietary selenium levels lower than those mentioned above (4). Proliferation of the hepatic parenchyma was reported to be more prevalent in rats fed a semi purified diet supplemented with 0. 5-2.0 mg of selenium per kg of feed in comparison with control rats (27). Increased concentration of glutathione in the blood, decreased activity of succinate dehydrogenase in the liver, and some impairment of the excretory function of the liver have been associated with long-term
14.
SELENIUM
139
low-level exposure to sodium selenite. A number of behavioural effects were observed at this exposure level as well (28). Certain discrepancies exist between two reports on the effects of high selenium levels in drinking-water. Water containing selenite at the level of 2 mg of selenium per litre resulted in 50% mortality after less than 3 months in male rats, with less pronounced toxic effects in females (29). On the other hand, sodium selenite at a level of 3 mg of selenium per litre had no effect on the survival of male rats (30). Selenium compounds have been shown to be less toxic to animals kept on a higher dietary intake of selenium (31, 32). Experiments with monkeys fed a cariogenic diet and exposed to drinking-water containing sodium selenite at a level of 2 mg of selenium per litre for 15 months followed by 1 mgjlitre for 45 months indicated a cariogenic effect of selenium during tooth development, but not when the teeth were exposed posteruptively (33). There is insufficient evidence to support the claim of carcinogenic effects of high selenium intake in experimental animals (34). The criticism of deficiency in the design of the experiments or in the statistical evaluation of the results does not apply to a recent study revealing carcinogenic effects of long-term exposure to selenium disulfide in mice (5). On the other hand, several reports indicate that doses of selenium higher than nutritionally essential have a preventive effect on the development of cancer in experimental animals (4). A review of the anticarcinogenic effects of selenium in experimental animals has recently been published (35).
REFERENCES I. ROSENFELD, I. & BEATH, D. A. Se/emum: geobotany, biochemistry, toxicity and nutrition. New York, Academic Press, 1964. 2. NATIONAL RESEARCH COUNCIL. Drinkmg watl'r and health. Washington, DC, National Academy of Sciences, 1977. 3. MuTH, 0. H., ed. Selenium in biomedicine. Westport, CN, Avi Publishmg Company Inc., 1967. 4. Selenium. Washington, DC, National Academy of Sciences, 1976. 5. Ambient water quality critena for selemum. Washington, DC, US Environmental ProtectiOn Agency, 1980. 6. ERMAKOV, V. V. & KOVALSKIJ, V. V. Biological significance of selemum. Moscow, Nauka Publishing House, 1974. 7. NATIONAL RESEARCH COUNCIL. Drinking water and health, vol. 3. Washington, DC, National Academy of Sciences, 1980. 8. WHO Technical Report Series, No. 532, 1973 (Trace elements in human nutrition). 9. LEVANDER, 0. A. In: Proceedings of the symposium on selenium- tellurium m the environment, Pittsburgh, lndustnal Health Foundation, Inc., 1976. 10. STEWART, R. D. H. ET AL. Quantitative selemum metabolism in normal New Zealand women. British journal of nutrition, 40: 45 (1978). II. BROWN, D. G. ET AL. Effect of dietary selenium on the gastromtestinal absorption of "Se0 3 in the rat. International journal for vitamin and nutrition research, 42: 588 (1972).
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12. THOMSON, C. D In: Trace elements m human and ammal health and disease in New Zealand. Hamilton, Waikato Umversity Press, 1977. 13. DIPLOCK, A. T. Metabolic aspects of selenium actiOn and toxicity. C RC critical reviews m toxicology, 4: 271 (1976). 14. GRIFFITHS, N. M. ET AL. The metabolism of (75 Se) selenomethionine in four women. British ;ournal of nutrition, 35: 373 (1976). 15. BURK, R. F., JR. ET AL. Influence of dietary and mjected selenium on whole-body retention, route of excretion, and tissue retention of 7 'Seo 2 - in the rat. Journal of nutritwn, 102: 1049 (1972) 16. SCHWARZ, K. ET AL. Introduction. Symposmm on nutritional significance of selemum (Factor 3). Federation proceedmgs, 20: 665 (1961 ). 17. KESHAN DISEASE RESEARCH GROUP OF THE CHINESE ACADEMY OF MEDICAL SCIENCES. ObservatiOns on the effects of sodium selenite in the prevention of Keshan disease.
Chinese medzcal journal., 92: 471 (1979). 18. KESHA"' DISEASE RESEARCH GROUP OF THE CHINESE ACADEMY OF MEDICAL SCIENCES. Epidemiological studies on the etiologic relationship of selenmm and Keshan disease.
Chinese medzcal journal, 92: 477 ( 1979). 19. XIAOSHU CHEN ET AL. The relatiOns of selenium and Keshan disease. Bwlogzcal trace element research, 2: 91 (1980). 20. SMITH, M. I. ET AL. The selenium problem in relatiOn to public health. A preliminary 21. study to determme the possibility of selenium intoxication m the rural populatiOn living on seleniferous soil. Public health reports, 51: 1496 (1936). SMITH, M. I. & WESTFALL, B. B. Further field studies on the selenium problem in relation to public health. Public health reports, 52: 1375 (1937). JAFFE, W. G. ET AL. Estudio clinico bioquimico en nil1os escolares de una zona selinifera. Archivos latinoamericanos de nutncion, 22: 595 (1972). JAFFE, W. G. In: Proceedings of the symposium on selenium- tellurium in the em•ironment, Pittsburgh, Industrial Health Foundation, Inc .. 1976. GLOVER, J. ET AL. In: Friberg, L. et al., ed. Handbook on the toxicology of metals, Amsterdam, Elsevier/North-Holland Biomedical Press, 1979. ALLAWAY, W. H. An overview of distribution patterns of trace elements in soil and plants. Annals of the New York Academy of Sciences, 199: 17 (1972). SCHWARZ, K. In: Muth, 0. H., ed. Selenium in biomedzcine. Westport, CN, Avi Publishing Company Inc., 1967, pp. 225-226. HARR, J. R. ET AL. In: Muth, 0. H., ed. Selenium in biomedicine. Westport, CN, A vi Publishing Company Inc., 1967. PLETNIKOVA, I. P. BIOlogical effects and safe concentrations of selenium m drinking water. Hygzene and sanitation, 35: 176 (1970). ScHROEDER, H. A. & MITCHENER, M. Selenmm and tellurium m rats. Effects on growth survival and tumors. Journal of nutrition, 101: 1531 (1971). PALMER, I. S. & OLSON, 0. E. Relative toxicities of selenite and selenate in the drinking water of rats. Journal of nutntwn, 104: 306 (1974). JAFFE, W. G. & MoNDRAGON, M. C. Adaptation of rats to selenium intake. Journal of
22. 23.
24. 25. 26. 27. 28. 29. 30. 31.
nutrition, 97: 431 (1969). 32. JAFFE, W. G. & MONDRAGON, M. C. Effects of ingestion of orgamc selenium m adapted and non-adapted rats. British ;ournal of nutrition, 33: 387 (1975). 33. BowEN, W. H. The effects of selenium and vanadium on caries activity in monkeys
(M. irus). Journal of the Irish Dental Association, 18: 83 (1972). 34. Some aziridines, N-, S- and 0-mustards and selenium. Lyon, International Agency for Research on Cancer, 1975 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 9), p. 245. 35. JACOBS, M. M. Effects of selenium on chemical carcinogens. Preventive medicine, 9: 362-367 (1980).
15. SILVER 15.1 General description
15.1.1 Sources
Silver occurs naturally in elemental form and as various ores, such as argentite and horn silver; it is also associated with lead, gold, copper, and zinc ores. Silver is present in the earth's crust at a concentration of about 0.1 mgfkg (1). It is used as a component of various alloys and solders, in photography, electrical equipment, electroplating, for fungicides, silverware, jewellery, coins, and dentalware (1). Because of their bacteriostatic properties, silver salts are used for water disinfection and as prophylactic agents (2).
15.1.2 Occurrence in water
Levels of silver in natural waters are very low (1). There is insufficient information to provide precise levels in water, but published data (1-3) suggest that few water sources contain more than 1 J-tg of silver per litre and levels greater than 10 J-Ig/litre are rare.
15.2 Routes of exposure 15.2.1 Drinking-water
A number of conventional water-treatment practices have been shown to be effective in removing silver from water and, consequently, many treated waters contain very low levels of silver (3). However, because some metals (such as lead and zinc) used in distribution systems may contain traces of silver, and also because in some countries silver oxide is used to disinfect water supplies, silver levels in tap-water may sometimes be elevated. Levels exceeding 50 J-tg/litre have been recorded on rare occasions, particularly when silver-containing, point-of-use water purifiers have been employed to obtain drinking-water (2). The average levels in tap-water are low and certainly less than 1 J-tgflitre. Assuming a consumption of 2 litres per day, the average daily exposure from drinking-water would thus not be likely to exceed 2 J-tg. 141
142 15.2.2 Food
Ill.
HEALTH-RELATED INORGANIC CONSTITUENTS
Few published data are available on the silver content of foodstuffs or diet, although most foods seem to contain trace amounts (less than 1 mg/kg)_ The reported exception is mushrooms, which may contain several hundred mg/kg (3)_ Various diets have been estimated to provide from 1 to 80 Jlg of silver per day (3, 4), although, where silver utensils are used, the amounts ingested could be very much higher_ Vegetables cooked in water containing silver absorb the metal very effectively (3). There is inadequate information to provide a precise average dietary intake, but estimated values within the range 20-80 JJ.g/day might be reasonable. 15.2.3 Air Little information has been published on the silver levels in air, but values up to 0.1 Jlg/m 3 of air have been reported in the USA (5, 6). Industrial emissions of silver are controlled for economic reasons and, as the levels present in fossil fuels are reported to be very low, the silver content of ambient air will be correspondingly low. A typical level in urban air would not be expected to exceed 0.05 JJ.g/m 3 and the exposure from this source is negligible. 15.2.4 Other routes of exposure 15.2.4.1 Industrial exposure Little information is available on industrial levels. Except in brazing operations and the manufacture of silver varnish, industrial exposures are considered to be very low. 15.2.4.2 Use of pharmaceutical preparations The topical use of certain products containing silver can result in significant exposure (7, 8). 15.2.5 Relative significance of different routes of exposure On the basis of the considerations in sections 15.2.1-15.2.4, the average daily intake of silver may be roughly assessed. Assuming that the silver concentration in water is 1 Jlgflitre and the daily water consumption is 2 litres, the intake from water is 2 Jlg of silver daily. Assuming also that the silver intake from air is negligible and the daily diet contains 20-80 Jlg of silver, the average daily intake from all sources will be about 22-82 Jlg. Using a silver-containing bacteriostatic point-of-use water purifier, producing a concentration of silver in the drinking-water of 50 JJ.g/litre,
15.
SILVER
143
together with a daily diet containing as much as 80 Jlg of silver, the daily intake may be as high as 180 Jlg per person. Generally, the levels in drinking-water are low, usually not in excess of I Jlg/litre, and thus the intake of silver from this source, in relation to normal total daily intake from all sources, is unlikely to be more than 5 '/'~. Precise data for the absorption of silver by human beings is not available and no figures for uptake can be provided.
15.3 Metabolism Relatively little is known about the absorption and metabolism of silver in humans (1) except that individuals and individual organs absorb the metal selectively (3). Animals seem to absorb about 10% of any ingested silver (2). Silver can be detected in various organs; the liver and spleen especially seem to concentrate the metal (2). In humans, more than 50 'I~ of the body burden can be found in the liver 16 days after exposure (9). Inhaled silver is also absorbed to a slight extent (2). Silver combines with the sulfhydryl component of some enzyme systems and other biologically important chemical groups, thus influencing the precipitation of proteins and inactivating some enzyme systems (10). Animal experiments have also shown that silver interacts metabolically with copper and selenium (4). Most of the absorbed silver is excreted almost exclusively with the faeces, and only small quantities are permanently retained by the tissues, the exception being the skin, where larger amounts of silver can accumulate (3). Silver that is available for excretion has a biological lifetime in the body ranging from a few days to a few weeks (2).
15.4 Health effects There is no evidence that silver is essential to the human organism. Cases of fatal poisoning have been recorded, but only with extremely high doses. The main effect of silver is discoloration of skin, hair, and fingernails (argyria). This has been detected when silver arsphenamine has been administered as a medication (3). A single dose of I g of silver, injected as silver arsphenamine, can produce this effect (1 1). The effect has also been observed in workers industrially exposed to silver; the condition is rarely encountered nowadays, however (3). It is possible that argyria may occasionally mask some mild systemic effects (2, 3). There is no evidence that ingested silver is carcinogenic (12). Pathological changes have been observed in the kidneys and liver of rats consuming water with silver concentrations of 400 J.tg/litre and above. It is difficult, however, to extrapolate these results to man. If it is assumed that the first appearance of argyria has no significant health effect, then the discoloration could be used to estimate a safe exposure level. The minimum dose for human beings that might induce argyria is
144
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1000 mg of silver (13). A lifetime (70 years) exposure to 1000 mg of silver would be equivalent to a continuous daily exposure to 40 J-Ig of silver. However, because silver is continuously excreted and only approximately 10% is absorbed, the daily exposure level needed to cause argyria over a lifetime could be as high as 400 ,_,g.
REFERENCES I. Guidelines for Canadian drinking water quality, 1978. Quebec, Ministry of Supply and Services, 1980 (supporting documentation). 2. Toxicology of metals, vol. II. Washington, DC, US Environmental Protection Agency, 1977 (Environmental Health Effects Research Series). 3. NATIONAL RESEARCH COUNCIL. Drinkmg water and health. Washington, DC, National Academy of Sciences, 1977. 4. UNDERWOOD, E. J. Trace elements in human and animal nutrition. New York, Academic Press, 1977. 5. GREENBERG, R. R. ET AL. Composiuon and size distributions of particles released in refuse incineratiOn. Environmental science and technology, I2: 566 (1978). 6. RAGAINI, R. C. ET AL. Environmental trace metal contamination in Kellogg, Idaho, near a lead smelting complex. Enmronmental science and technology, II: 733 (1977). 7. PARISER, R. J. Generalized argyria. Clinicopathologic features and histochemical studies. Archives of dermatology, I44: 373 (1978). 8. MARSHALL, J. P. & ScHNEIDER, R. P. Systemic argyria secondary to topical stlver nitrate. Archives of dermatology, ll3: 1077 (1977). 9. NEWTON, D. & HoLMES, A. A case of acctdental inhalation of Zn-65 and Silver-110. Radiation research, 29: 403 (1966). 10. GooDMAN, L. S. & GILMAN, A. The pharmacological basis of therapeutics, 5th ed. New York, Macmillan, 1975. II. HILL, W. B. & PILLSBURY, D. M. Argyria. The pharmacology of silver. Baltimore, MD, Williams and Wilkins, 1939. 12. ENVIRONMENTAL PROTECTION AGENCY. Water quality criteria; availability. Federal register, 44: 15 964 (1979).
16. SODIUMa 16.1 General description 16.1.1 Sources Sodium is present in a number of minerals, the principal one being rock salt (sodium chloride). Seawater contains relatively high levels of sodium. Overall, sodium represents about 26 gjkg of the earth's crust (1). Sodium and its salts are used for a wide variety of purposes, including the de-icing of roads, in the paper, glass and soap industries, in the pharmaceutical and general chemical industries, for the treatment of water, in the food industry, and for culinary purposes. It is widely present, sometimes in substantial quantities, in soils, plants, water, and many foods. Most countries have significant mineral deposits of sodium. Considerable amounts are excreted by humans and it is a common constituent of domestic sewage.
16.1.2 Occurrence in water The sodium ion is ubiquitous in water owing to the high solubility of its salts and the abundance of mineral deposits. Seawater contains about 10 g of sodium per litre. The highest freshwater levels are found in lowland rivers and in groundwater. Upland streams and associated reservoirs will tend to have a relatively low sodium content. Particularly elevated levels of sodium are associated with groundwater in areas where there is an abundance of sodium mineral deposits or where there has been contamination from saline intrusion (sea and estuarine sources) or other forms of pollution (2). Near coastal areas, windborne sea spray can make an important contribution, either by fallout on to land surfaces where it drains to the water source or from washout by rain from the air to surface-water sources (2). Discharge of effluents (domestic, commercial, industrial wastes) to rivers is another important source of sodium in water. Levels in such rivers would be a function of many factors, including river flow rate and sodium concentration in the effluent. Substantial concentrations can sometimes be detected in some rivers, especially at times of low river flow. a A discussion of some other aspects of the effect of sodium on water quality will be found in Part V, section II, p. 286
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16.2 Routes of exposure 16.2.1 Drinking-water
In most countries, the majority of water supplies contain less than 20 mg of sodium per litre, but in some countries sodium levels can exceed 250 mg/litre. Apart from saline intrusion and natural contamination, salt used in de-icing roads, water-treatment chemicals, domestic water softeners, sea-salt spray and sewage effluents can all contribute significant quantities to water. Water-treatment chemicals, such as sodium fluoride, sodium silicofluoride, sodium hydroxide, sodium carbonate, sodium bicarbonate, and sodium hypochlorite provide, as individual chemicals, a relatively small contribution, but collectively the amounts can be significant; levels of 30 mgjlitre might result in some instances (2). Domestic water softeners can provide levels of over 300 mgjlitre, but in general the levels are much lower (2). Most people are exposed to less than 50 mg of sodium per day by drinking tap-water (based on a consumption of 2 litres per day). As sodium salts are very soluble, virtually all the sodium present in water, whether consumed directly, in the preparation of beverages, or incorporated into food, will be absorbed. 16.2.2 Food
Sodium is naturally present in all foods. The way food is processed governs critically the final sodium content. For example, frozen peas contain much more sodium than fresh peas (2). Fresh fruit and vegetables contain from less than 10 mgjkg to 1 g/kg, in contrast with cereals and cheeses, which may contain 10-20 g/kg (2). Milk contains a relatively high proportion of sodium, i.e., 1.5 gjlitre (2). Bottled waters can also sometimes contain similar levels of sodium (2). The estimation of the daily intake of sodium from food is difficult because of the wide variations of concentrations in foods and the fact that many people add salt to their food. In western Europe and North America, the overall dietary sodium chloride consumption is estimated to be 5-20 gjday (2-8 g of sodium per day), with an average of around 10 gjday (4 g of sodium) (2). For medical reasons, some people need a special lowsodium diet calling for a sodium intake of less than 2 g of sodium per day (3). In the case of artificial infant feeding, regulations requiring a reduction of the sodium content of infant food have been widely introduced. 16.2.3 Air
The level of inhaled sodium in ambient air and factory atmospheres is small relative to the amount of sodium absorbed into the body from the diet.
16.
SODIUM
147
16.2.4 Relative significance of different routes of exposure
For different individuals there can be a fairly wide range of exposures to sodium. Generally, by far the major source is food. Since sodium is readily absorbed after ingestion, uptake can be considered to be equivalent to the exposure to the element. Uptakes for various categories of dietary and water sources have been calculated and are given in the tables below. The information for adults is based on published data on the dietary input of sodium (2). Calculations for infants (0-2 months) are based on an estimate of 250 mg of sodium per day and for children 1-5 years on an estimate of 2000 mg/day (2). 16.2.4.1 Weekly uptake of sodium from water and diet in adults (a) Special restricted diet-500mg of sodium per day Sodium concentration in water Weekly uptake of sod1um (mg) Water only Food only Total Ratio· Water/total ( %)
20 mgjlitre 50 mg/litre 100 mgjlitre 200 mg/litre
280 700 1400 2800
3500 3500 3500 3500
3780 4200 4900 6300
7 17 28 44
(b) Relatively low sodium diet-2000 mg per day Sodium concentration in water Weekly uptake of sodium (mg) Water only Food only Total Rat1o: water/total ( %)
20 mgjlitre 50 mgjlitre 100 mgjlitre 200 mg/litre
280 700 1400 2800
14000 14000 14000 14000
14280 14 700 15400 16800
2 5 9 17
(c) Typical sodium diet-5000 mg of sodium per day Sod1um concentration m water Weekly uptake of sod1um (mg) Water only Food only Total Ratio. Water/total ( %)
20 mgjlitre 50 mgjlitre 100 mgjlitre 200 mgjlitre
280 700 1400 2800
35000 35000 35000 35000
35280 35700 36400 37800
1 2 4 7
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Ill.
HEALTH-RELATED INORGANIC CONSTITUENTS
(d) Total intake restricted to 500 mg of sodium per day Sodium concentration in water 20 mg/litre 50 mg/litre 100mgjlitre 200 mg/litre Weekly uptake of sod1um (mg) Water only 280 700 1400 2800 Food only 3220 2800 2100 700 Total 3500 3500 3500 3500 RatiO: waterjtotal ( %) 8 20 40 80
All the above calculations are based on a water consumption of 2 litres per day. 16.2.4.2 Weekly uptake of sodium from water and diet in children (a) Infants aged 0---2 months-250mg of sodium per day from food Sodium concentration 1n water 20 50 1 00 200 8
Weekly uptake of sod1um (mg) Water only• 140 350 700 1400 Food only 1 750 1 750 1 750 1 750 Total 1 890 2100 2450 3150 Ratio: waterjtotal ( %) 7
mgjlitre mgjlitre mgjlitre mgjlitre
17 29 44
Based on a water consumption of 1 0 litre per day
(b) Children aged 1-5 years-2000 mg of sodium per day from diet Sodium concentration in water 20 50 100 200 8
Weekly uptake of sod1um (mg) Water only• 210 525 1050 2100 Food only 14000 14000 14000 14000 Total 14210 14525 15050 16100 RatiO. waterjtotal ( %) 1 4 7 13
mgjlitre mg/litre mg/litre mg/litre
Based on a water consumptiOn of 1 5 htres per day
16.3 Metabolism The metabolism of sodium has been studied extensively because of its physiological properties and its importance to the body (3). Reference is made here only to the general aspects of the metabolism of sodium. It is the most abundant cation of plasma and extracellular fluid in man. It is present in bones, in cells, and in most tissues. The level of sodium in extracellular fluid is carefully maintained by the kidney under the
16.
SODIUM
149
influence of endocrine, cardiovascular, and autonomic regulatory mechanisms. The total amount of sodium in extracellular fluid thus determines the volume of these fluids (3). Control of sodium balance is achieved through a complex interrelated system involving both nervous and hormonal systems (1). Increases in the plasma sodium concentration stimulate the osmoreceptors in the hypothalamic centre, regardless of fluid volume, with the resultant sensation of thirst (1). In hot climates and during heavy work, a substantial loss of sodium occurs by perspiration and additional salt may be needed to make up the loss (1). The intake of sodium is not physiologically controlled. More than 90% of the amount in food is absorbed (2). Intake is normally dependent primarily on diet. The minimum sodium chloride requirement is about 120 mgjday (approximately 50 mg of sodium in this form) (2). 16.4 Health effects 16.4.1 Acute effects In general, sodium salts are not acutely toxic substances because of the efficiency with which mature kidneys excrete sodium (1). Excessive intake of sodium chloride causes vomiting and the elimination of much of the salt. Acute effects may include convulsions, muscular twitching and rigidity, and cerebral and pulmonary oedema (1). The effects on infants, in contrast to adults, are different because of the immaturity of infant kidneys (2). Acute effects and death have been reported in cases of accidental overdoses of sodium chloride (2). Severe deterioration of chronic congestive heart failure can result from excessive salt intake, and ill effects due to high levels of sodium in drinking-water have been documented (2). 16.4.2 Hypernatraemia Infants with severe gastrointestinal infections can suffer from fluid loss leading to dehydration and raised sodium levels in the plasma (hypematraemia); permanent neurological damage is common under such conditions (2). Evidence that a raised sodium intake is a factor in "sudden infant death" is limited and indirect and is not generally thought to be conclusive (2). However, for healthy infants and children, the total sodium intake should be kept as low as possible (2). Modem infant feeding practice using cows' milk added to solid feed has been suggested as one of the causes of hypematraemia (2). The situation could be exacerbated if tap-water containing high levels of sodium was also incorporated into the feed (2). The concentration of sodium in cows' milk is about three times that in human breast milk (2). The immature kidneys of infants are not as effective as those of adults in
150
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HEALTH-RELATED INORGANIC CONSTITUENTS
maintaining plasma osmolarity so that precautions are needed, such as the introduction of regulations requiring a reduction of sodium in infant foods. 16.4.3 Raised sodium intake and hypertension
There has been considerable scientific controversy for some time on this issue (2), and there are persuasive scientific arguments for the hypothesis that salt consumption affects the development and level of hypertension. 16.4.3.1 Evidence from animal experiments Hypertension has been clearly demonstrated in different species of animals given high levels of sodium chloride in their diet (2). Despite the usual reservations about extrapolating animal results to humans, the consistency of the animal data suggests the validity of such a procedure. 16.4.3.2 Evidence from human volunteer studies There is no conclusive evidence that elevated blood pressure is related to high-salt diets fed to volunteers. However, there is some doubt about these short-term studies in relation to hypertension. Most people in western societies ingest a high-salt diet from infancy, yet persistent hypertension is uncommon until the fourth decade (2). 16.4.3.3 Evidence from epidemiological studies (a) High-sodium diet population studies
A particularly striking contrast is that between some non-westernized groups and western populations. Non-westernized groups have lowsodium diets and a very low prevalence of hypertension, with no increase in blood pressure with age (2). Although one may be tempted to conclude a causal relation, there are a number of other differences between the two populations that might account for such a contrast (2). However, the strong consistency between these results and those of other studies (2) gives further support to there being a direct link between raised sodium intake and hypertension. (b) Sodium intake studies via drinking-water
Recently completed epidemiological studies in the USA and the Netherlands have demonstrated that schoolchildren (particularly girls) living in areas with moderate levels of sodium in the drinking-water (128-161 mg/litre) had slightly higher blood pressures (3-5 mmHg) than those living in areas with low levels of sodium (28 mgjlitre) (4-6). A somewhat similar study in the USSR of people in the age range 16-60 years demonstrated a similar relationship between sodium in water and blood pressure (7).
16.
SODIUM
151
In a study in the USA (6), children (10-11 years old) living in the same community where the drinking-water had a high sodium content (108 mgjlitre) were grouped into triads and matched by systolic blood pressure. Two of the three groups were supplied with bottled water containing 108 mg of sodium per litre and the third with bottled water of low sodium content (8 mgjlitre). The results for females were consistent with the results of the other two reported studies in the USA; girls provided with bottled water of low sodium content had lower blood pressures than those in the other two groups. The blood pressure differences for boys followed an identical trend for the first six weeks, but were not consistent for the remainder of the study. 16.4.4 Relationship between sodium in water and other diseases Although there is an association between hypertension and some diseases, such as coronary heart diseases, genetic differences in susceptibility, possible protective minerals (potassium and calcium), and methodological weaknesses in experiments make it difficult to quantify a relationship. Sodium levels in drinking-water are generally only a small contributor to dietary sodium. No firm conclusions can at present be drawn on the importance of sodium in drinking-water and its possible association with disease. The relevance of ingested sodium from all sources was reviewed by a WHO Working Group in 1978 (2). One of the recommendations of the group was that in areas where levels of sodium exceeded 20 mgjlitre, public health authorities should be notified because certain people (patients with hypertension or congestive heart failure) need to restrict their overall dietary intake of sodium. As any action to be taken depends upon local conditions and policies, no specific level based on health considerations is recommended in the present guidelines (see Part V, section 11.4, for a guideline value based on taste threshold). REFERENCES I. Guidelines for Canadian drinking water quality, 1978. Quebec, Ministry of Supply and Services, 1980 (supporting documentation). 2. Sodium, chlondes and conductwity in drinkmg·water. Copenhagen, WHO Regional Office for Europe, 1979 (EURO reports and studies, No. 2). 3. NATIONAL RESEARCH COUNCIL. Drinking water and health. Washington, DC, National Academy of Sciences, 1977. 4. CALABRESE, E. J. & TUTHILL, R. W. Elevated blood pressure and high sodium levels in the public drinking water. Archives of environmental health, 35: 200 (1977). 5. TuTHILL, R. W. & CALABRESE, E. J. Elevated sodium levels in the public drinking water as a contributor to elevated blood pressure levels in the community. Archives of em•1ronmental health, 37: 197 ( 1979). 6. TUTHILL, R. W. & CALABRESE, E. J. Drinking water sodium and blood pressure in children· a second look. Amencan journal of public health, 71. 722-729 (1981). 7. FATULA, M. I The frequency of arterial hypertension among persons using water with an elevated sodium chloride content. Sovetskaja medicina, 30: 123 (1967).
PART IV.
HEALTH-RELATED ORGANIC CONSTITUENTS
1. CHLORINATED ALKANES One of the major uses of chlorinated alkanes in the chemical industry is as an intermediate in the production of other organochlorine compounds. They are therefore produced in large quantities and consequently many are found in both raw and finished drinking-water. Of the large number of chlorinated ethanes known to be produced commercially, only I ,2-dichloroethane can be clearly labelled as a carcinogenic hazard on the basis of the available data. 1.1 Carbon tetrachloride 1.1.1 General aspects Carbon tetrachloride (CCI 4 ) is a haloalkane with a wide range of industrial and chemical applications. At room temperature it occurs as a heavy, colourless liquid with a density of 1594 g/litre. It is relatively nonpolar, miscible with alcohol, acetone and most organic solvents, and soluble in water to the extent of 800 mgjlitre at 25 °C. Approximately 423 000 tonnes (932. 7 x 10 6 pounds) are produced at 11 plant sites in the USA. 0 The major portion of this production is used in the manufacture of fluorocarbons (95% in 1973), which are used primarily as aerosol propellants. Accidental CC1 4 spills have occurred and as much as 63.6 tonnes were discharged into the Ohio River in February 1977, resulting in surface-water concentrations as high as 340 Jlg/litre (1). It is also frequently found in contaminated groundwater. Its general occurrence in raw water supplies results in reported levels of 2-3 Jlg/litre in finished drinking-water. Hydrolytic decomposition, as a means of removal from water, appears to be insignificant as compared with evaporation. 1.1.2 Routes of exposure 1.1.2.1 Water In the National Organic Reconnaissance Survey (NORS) study performed in the USA by the EPA (2), CC1 4 was found in 10 'J~ of the drinking-water supplies at levels less than 2-3 Jlg/litre. In New Orleans, • JOHNS, R. Air pollution assessment of carbon tetrach/oruk. Prepared under contract for the US Envtronmental ProtectiOn Agency. McLean, VA., Mitre Corp., 1976.
155
156
IV.
HEALTH-RELATED ORGANIC CONSTITUENTS
CCl 4 was found in both human blood plasma and drinking-water. CCl 4 has been found to be an occasional contaminant of chlorine used in the disinfection of drinking-waters, but it is not produced in drinking-water as a result of the chlorination process itself. 1.1.2.2 Food
CCl 4 has been detected in a variety of foodstuffs with levels ranging from 0.1 to 20 Jlg/kg. McConnell et al. (3) provide a summary of the different food categories that have been shown to be particularly susceptible to CCl 4 contamination. They noted that there is no evidence of significant bioaccumulation of CCl 4 via the food chain to higher trophic levels. The primary means of contamination of food with CC1 4 is through the use of CC1 4 as a fumigant. 1.1.2.3 Air
Carbon tetrachloride has been measured extensively in the atmosphere; consequently, there is a good understanding of its atmospheric distribution. The occurrence of CCl 4 in the atmosphere is due largely to the fact that it is such a volatile compound. The sources of CCl 4 have been found to be primarily man-made (4-7). The atmospheric distribution is approaching homogeneity. Some high concentrations have been reported in urban air; however, they are generally close to the background level of 0.00078-0.00091 mg(m 3 found in the continental air mass. 1.1.3 Metabolism 1.1.3.1 Absorption
CCl 4 is readily absorbed through the lungs and more slowly but still completely absorbed through the gastrointestinal tract (8). It can also enter the body by penetration through the skin. The rate and amount of absorption is enhanced by the simultaneous ingestion of fat (8) and alcohol (8- 10). In an investigation of the absorption of CCl 4 from the gastrointestinal tract of dogs, Robbins (JJ) found that considerable quantities were absorbed from the small intestine, less from the colon, and little from the stomach. 1.1.3.2 Distribution
Nielsen & Larsen (8) found high concentrations of CCl 4 in animal testicular fatty tissues, liver, brain, bone marrow, and kidneys. Robbins (JJ) studied the distribution of CCl 4 in dogs after oral administration. The highest concentration of CCl 4 was found in the bone marrow. The amount found in the liver, pancreas, and spleen was one-fifth of the amount found in bone marrow. From work done by Recknagel &
1.
CHLORINATED ALKANES
157
Litteria (12), it would appear that the organ distribution of CC1 4 varies with the route of administration, its concentration, and the duration of exposure. At the cellular level, McLean et al. (13) found CC1 4 in all cell fractions, with higher concentrations in the ribosomes. 1.1.3.3 Biotransformation
When CC1 4 is administered to mammals, it is metabolized to a small extent, but the majority is excreted through the lungs. The metabolites include chloroform, hexachloroethane, and carbon dioxide. Research has revealed that these metabolites play an important role in the overall toxicity of CC1 4 (14).
1.1.4 Health effects The toxicity of CC1 4 to humans is not often recognized (15). Acute and subacute toxicity has resulted from oral, dermal, and inhalation exposures, with adverse effects on the skin, circulation, respiration, blood, and the function of the kidneys, liver, eyes, and pancreas. In many instances of acute poisoning, the patient develops signs of liver injury within a few days. The patient becomes jaundiced and the liver becomes enlarged and tender. As liver injury develops, and sometimes in its absence, injury of the kidneys may be observed; at times this may dominate the clinical picture and it is often responsible for early death (16). In general, hepatic complications are a more frequent response to CC1 4 toxicity than are renal complications. Changes in blood parameters, visual acuity, and the pancreas have also been noted. The clinical picture of chronic CC1 4 poisoning is much less characteristic than that of acute poisoning. Reports of pathological changes in persons dying from CC1 4 poisoning are generally limited to findings in the liver and kidneys. There are very few reports on the mutagenesis of CC1 4 . Kraemer et al. (17) found that CC1 4 was not mutagenic in the Salmonella typhimurium or Escherichia coli reversion tests; however, halogenated hydrocarbons are usually negative in the Ames test. The available data appear to be sufficient to permit the conclusion that CC1 4 is a carcinogen to laboratory animals (18-24). The standard for a permissible level of human exposure is based on the study performed in the USA by the National Cancer Institute (24) on trichloroethene, in which CC1 4 was utilized as a positive control. In this work, CC1 4 was found to be carcinogenic in the B6C3-Fl mouse. Although other studies have been conducted, inadequate dose-response information was obtained or the experiments were of too short a duration to allow utilization of the data for risk estimates. Because there are doubts regarding the mechanism of tumorigenesis in the liver of this strain of mouse with agents that are known hepatotoxins (such as CC1 4 ),
158
JV.
HEALTH-RELATED ORGANIC CONSTITUENTS
the appropriateness of a no-threshold model for extrapolation is questionable. Alternative extrapolation models do not exist for epigenetically induced cancer. The need for guidance on this frequent contaminant of drinking-water dictates a conservative approach. Consequently, a linear multistage extrapolation model was used to derive a tentative guideline value for CC1 4 of 3 JLg/litre. This level should give rise to less than 1 additional cancer per 100 000 population for a lifetime of exposure, assuming a daily consumption of drinking-water of 2 litres.
REFERENCES 1. AMERICAN CHEMICAL SociETY. CCI 4 spill causes along Ohio River. Chemical and engineering news, 55: 7 (1977). 2. Ambient water qua/zty critena for carbon tetrachloride. Washington, DC, US Environmental Protection Agency, 1980 (EPA 440/5-80-026). 3. McCONNELL, G. ET AL. Chlorinated hydrocarbons m the environment. Endeavour, 34: 13 (1975). 4. ALTSCHULLER, A. P. Average tropospheric concentration of carbon tetrachloride based on mdustrial production, usage and emissions. Environmental sc1ence technology, 10: 596 (1976). 5. LovELOCK, J. E. ET AL. Halogenated hydrocarbons in and over the Atlantic. Nature, 247: 194 (1974). 6. WILKNISS, P. E. ET AL. Atmospheric trace gases in the southern hem1sphere. Nature, 245: 45 (1973). 7. SINGH, H. B. ET AL. Atmospheric carbon tetrachlonde: Another man-made pollutant. Science, 192: 1231 (1976). 8. NIELSEN, V. K. & LARSEN, J. Acute renal failure due to carbon tetrachloride poisoning. Acta medica Scandinavica, 178: 363 (1965). 9. FoLLAND, D. S. ET AL. Carbon tetrachloride tox1city potentiated by Isopropyl alcohol. Investigation of an industrial outbreak. Journal of the Amencan Med1cal Assocwtwn, 23: 1853 (1976). 10. MooN, H. D. Pathology of fatal carbon tetrachloride poisoning with special reference to histogenesis of the hepatic and renal lesions. American ;ournal of pathology, 26: 1041 (1950). 11. ROBBINS, B. H. The absorption, distributiOn, and excretiOn of carbon tetrachloride m dogs under various conditwns. Journal of pharmacology, 37: 203 (1929). 12. RECKNAGEL, R. 0. & LITTERIA, M Bwchem1cal changes in carbon tetrachloride fatty liver: ConcentratiOn of carbon tetrachloride m liver and blood. Amencan journal of pathology, 36: 521 (1960). 13. McLEAN, A S. M. ET AL. Cellular necros1s in the liver mduced and modified by drugs and other agents. International review of experimental pathology, 4: 127 (1965). 14. GORDIS, E. Lipid metabolites of carbon tetrachloride. Journal of clinical im•est1gallon, 48: 203 (1969). 15. VoN OETTINGEN, W. F. The halogenated hydrocarbons of industrial and toxicological importance. In: Browning, E., ed. Elset'ler monographs on toxic agents. New York, Elsev1er Publishing Co., 1964. 16. VON 0ETTINGEN, W. F. The halogenated aliphat1c, olefimc, cyc/1c, aromatic, and aliphatic-aromatiC hydrocarbons mcludmg the halogenated insectiCides, the1r tOXICity and potential dangers. Washington, DC. Department of Health, Education & Welfare, 1955. 17. KRAEMER, M. ET AL. S. typhimurium and E. coli to detect chemical mutagens. NaunynSchmiedebergs archives of pharmacology, 284: 46R (Abstract).
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159
18. Some halogenated hydrocarbons. Lyons, International Agency for Research on Cancer, !979 (IARC Monographs on the evaluation of carcinogenic risk of chemicals to humans, vol. 20). 19. EDWARDS, J. Hepatomas in mice induced with carbon tetrachloride. Journal of the National Cancer Institute, 2: 197 (1941). 20. EDWARDS, J. & DALTON, A. Induction of cirrhosis of the liver and hepatomas in mice with carbon tetrachloride. Journal of the National Cancer Institute, 3: 19 (1942). 21. EDWARDS, J. ET AL. InductiOn of the carbon tetrachloride hepatoma in strain L mice. Journal of the National Cancer Institute, 2: 297 (1942). 22. ESCHENBRENNER, A. B. & MILLER, E. Studies on hepatomas-size and spacing of multiple doses in the induction of carbon tetrachloride hepatomas. Journal of the Natwnal Cancer lnstllute, 4~ 385 (1943). 23. EscHENBRENNER, A. B. & MILLER, E. Liver necrosis and the induction of carbon tetrachloride hepatomas in strain A mice. Journal of the National Cancer Institute, 6: 325 (1946). 24. NATIONAL CANCER INSTITUTE. Carcinogenesis bioassay of trichloroethylene. Washington, DC, US Department of Health, Education & Welfare, 1976 (CAS No. 79-01-6, NCI-CG-TR-2).
1.2 1,2-Dichloroethane 1.2.1 General aspects 1,2-Dichloroethane (CH 2 Cl-CH 2 Cl) is a liquid with a relative density of 1.25, and a threshold odour limit of 2 mgjlitre (1). It is used extensively as a solvent for a great many organic chemicals, as an intermediate in chemical synthesis, and as an insecticide. Production of I ,2-dichloroethane in the USA in 1976 was 3.63 x I 0 6 tonnes (8 x I 0 9 pounds) (2).
1.2.2 Routes of exposure 1.2.2.1 Drinking-water As a result of the use of 1,2-dichloroethane in industry, it is a component of industrial effluents and has been detected in raw and finished drinking-water in the USA. 1,2-Dichloroethane was found in the water of 28 cities in the USA at levels up to 6 ,ug/litre (3). 1.2.2.2 Food 0.4% of I ,2-dichloroethane exposure results from the consumption of aquatic organisms, which exhibit an average bioconcentration potential of 1.2-fold. The remaining 99.6% of exposure results from drinkingwater (4). 1.2.2.3 Industrial exposure In the USA, the National Institute for Occupational Safety and Health (NIOSH) (2) estimated that 4.5 million workers are exposed to I ,2-dichloroethane by inhalation and dermal routes.
160 1.2.2.4 Air
IV.
HEALTH-RELATED ORGANIC CONSTITUENTS
1,2-Dichloroethane has been detected in urban air at levels between 0.04 and 38 JJ.g/m 3 (5). As the result of manufacture, storage, and distribution it was calculated in the USA in 1974 that the emissions of 1,2-dichloroethane to the ambient air amounted to approximately 74 x I 06 kg, i.e., about 1.8% of the total production.
1.2.3 Metabolism Few data are available concerning the metabolism of 1,2-dichloroethane. It is known that this substance is readily soluble in the lipids of the brain; this property promotes the influence of I ,2dichloroethane on the nervous system (J).
1.2.4 Health effects Data on the toxicity of 1,2-dichloroethane are related mainly to its inhalation exposure in occupational conditions; it acts as a narcotic and causes damage to the liver, kidneys, and cardiovascular system (7). Insignificant symptoms of intoxication were found at concentrations under 4 mg of I ,2-dichloroethane per m 3 in air (J). On the basis of the available information, NIOSH recommended that occupational exposure to I ,2-dichloroethane should not exceed 20 mg/m 3 determined as a time-weighted average for up to a 10-hour working day with a 40-hour working week. Peak concentrations should not exceed 60 mgjm 3 as determined by a 15-minute sample. The exposure standard recommended by the Occupational Safety and Health Administration (OSHA) of the USA is 200mgjm 3 , while the standard for occupational conditions in the USSR is 10 mg/m 3 . The LD 50 of I ,2-dichloroethane administered orally to white rats was 1120 ± 142 mgjkg of body weight (7). Epidemiological studies have not disclosed a relationship between exposure to I ,2-dichloroethane and cancer. However the compound is carcinogenic in animal tests, inducing a statistically significant number of squamous cell carcinomas of the prestomach and haemangiosarcomas of the circulatory system in male rats, mammary adenocarcinomas in female rats and mice, and endometrial tumours in female mice (6, 7). The ambient water quality criterion for I ,2-dichloroethane in the USA was calculated by applying a linearized multistage model to the data from the appropriate bioassay of the National Cancer Institute. I ,2-Dichloroethane is a known mutagen. It was mutagenic in the Ames Salmonella assay for the strains TA 1530 and 1535, and was also mutagenic for the E. coli DNA polymerase-deficient system (8). It induced highly significant increases in somatic mutation frequencies in Drosophila melanogaster (9). Morphological and chlorophyll mutations in eight varieties of peas were induced by treatment of the seeds with I ,2-dichloroethane (10).
I.
CHLORJNA TED ALKANES
I6I
Chioroacetaidehyde, a postulated metabolite of I,2-dichloroethane, is mutagenic in Salmonella typhimurium TA IOO (see 11). The guideline value for I ,2-dichloroethane is based on the induction of circulatory system haemangiosarcomas in male Osborne-Mendel rats given oral doses of I,2-dichloroethane over a period of 78 weeks (10). The concentration of 1,2-dichloroethane in water, calculated to keep the lifetime cancer risk below w-s, is 9.4 jig/litre, i.e., approximately 10 jig/litre.
REFERENCES I. ZOETEMAN, B. C. ]. Sensory assessment and chemical composition of drinking-water. Leidschendam, Netherlands, Institute of Water Supply, 1978. 2. NATIONAL INSTITUTE FOR OCCUPATIONAL SAFETY AND HEALTH. Ethylene dichloride ( 1,2-dichloroethane). Washington, DC, Department of Health, Education and Welfare, 1978 (Current intelligence Bulletin 25 (NIOSH) publication No. 78149). 3. SYMONS, J. M. ET AL. National organics reconnaissance survey for halogenated organics. Journal of the American Water Works Association, 67: 634 (1975). 4. Ambient water quality criteria for chlorinated ethanes. Washington, DC, Environmental Protection Agency, 1980 (440/5-80-029). 5. 0KUNO, T. ET AL. [Gas chromatography of chlorinated hydrocarbons in urban air.) Hyogo-ken kogai kenkyusho kenkyu hokoku, 6: 1-6 (Chemical abstracts, 87.72564 f) (1974). 6. NATIONAL CANCER INSTITUTE. Bioassay of 1,2-dichloroethane for possible carcinogenicity. Washington, DC, US Department of Health, Education & Welfare, 1978 ((NIH) 78-1305). 7. Some halogennted hydrocarbons. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 20). 8. BREM, H. ET AL. The mutagenicity and DNA-modifying effect of haloalkanes. Cancer research, 34: 2576 (1974). 9. NYLANDER P. 0. ET AL. Mutagenic effects of petrol in Drosophila melanogaster. I. Effects of benzene and of 1,2-dichloroethane. Mutation research, 57: 163 (1978). 10. K!RICHEK, Y. F. Effect of 1,2-dichloroethane on mutations in peas. Uspehi himii mutageneza se., 232 (1974). II. Some monomers, plastics and synthetic elastomers, and acrolein. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 19).
2. CHLORINATED ETHENESa This group of compounds is used widely in a variety of industrial processes as solvents, softeners, paint thinners, dry-cleaning fluids, intermediates, etc. Because of their wide use, they are often found in raw and treated drinking-water. They are known to occur in groundwater at concentrations of a few milligrams per litre. Because of their high volatility, they are usually lost to the atmosphere from surface-water and therefore generally occur at lower concentrations. The compounds that are of interest within this group are those for which there are indications of carcinogenic activity in experimental animals. This group includes the well-known human carcinogen vinyl chloride. The occurrence of vinyl chloride in drinking-water seems to be primarily associated with the use of poorly polymerized poly(vinyl chloride) water-pipes, a problem that can be more appropriately controlled by product specification than by the setting of a guideline level. b 2.1 Vinyl chloride 2.1.1 General aspects Vinyl chloride is mainly used for the production of poly(vinyl chloride) (PVC) resins which, in turn, form the most widely used plastics in the world. Minor uses (less than 5% of total production) are as an intermediate in the manufacture of methyl chloroform and as a comonomer with vinylidene chloride in the production of vinylidene chloride-vinyl chloride copolymers, which are widely used in food packaging and as coatings. Vinyl chloride was formerly used as an aerosol propellant and as a refrigerant, but these uses appear to have been discontinued (J). The largest use of PVC is in the production of piping and conduits; other important uses are in floor coverings, in consumer goods, in electrical applications, and in transport applications (1). Vinyl chloride is volatile and readily passes from solution into the gaseous phase under most laboratory and ecological conditions. Low concentrations have been detected in effluents discharged by chemical a Formerly known as chlorinated ethylenes.
b US National Sanitation Foundation Standard No. 14. P/a.st1c piping components and related materials, revised December 1980, permits IOmg of vmyl chloride monomer per kilogram of p1pe for drinking·water.
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and latex manufacturing plants (1) and in drinking-water as a result of leaching from PVC pipes used in water distribution systems (2). A number of product standards exist which specify a quality of PVC waterpipe that limits the quantity of free vinyl chloride monomer (VCM) present. Provided pipes of this quality are used, the concentration of the vinyl monomer likely to be present in the drinking-water will be small compared with the value that would be derived by applying the same linear multistage extrapolation model that was used in the case of other carcinogenic organic substances (20 JLg/litre based upon an acceptable risk of less than one additional cancer case per 100 000 population over a lifetime).
2.1.2 Routes of exposure 2.1.2.1 Water Vinyl chloride in samples of wastewater from seven areas in the USA (associated with PVC-vinyl chloride manufacturing plants) ranged from 0.05 to 20 mg/litre (3). The highest concentration of vinyl chloride detected in finished drinking-water in the USA was 10 JLg/litre (4). In a five-city survey in the USA, concentrations of vinyl chloride up to 1.4 JLg/litre were detected in drinking-water taken from distribution systems constructed with PVC pipe (2). 2.1.2.2 Food Small quantities of vinyl chloride are ingested as a result of migration into foods from PVC packaging materials. Studies by the Food and Drug Administration of the USA indicated that up to 20 mg of vinyl chloride per kg was present in alcoholic beverages packaged in PVC containers (5). Vinyl chloride has been found at concentrations up to 14.8 mgjkg in edible oils, butter, and margarine packaged and stored in PVC containers (1). Many countries now limit the content of entrained vinyl chloride monomer in PVC packaging materials and prohibit the use of such materials for products containing alcohol or edible oils. 2.1.2.3 Air Vinyl chloride is a gas at normal atmospheric temperature and pressure and it occurs in the vicinity of vinyl chloride and PVC industries. Concentrations of up to 8.8 mg/m 3 have been detected in the air near vinyl chloride manufacturing plants (1). Vinyl chloride was formerly used as a propellant for many aerosol products, such as pesticides, hair sprays, and deodorants; consumers repeatedly using such products were undoubtedly exposed to moderately high concentrations. Vinyl chloride has been detected in concentrations of 1-3 mg/m 3 in the air in the interior of new automobiles (1).
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2.1.3 Metabolism
2.1. 3.1 Absorption Vinyl chloride is readily absorbed following oral administration (6, 7) or inhalation (7). 2.1.3.2 Distribution In studies on the distribution of vinyl chloride in rats, the greatest concentrations were detected in the liver, kidneys, and spleen (6, 8). 2.1.3.3 Biotransformation Vinyl chloride is metabolized by microsomal mixed-function oxidases (predominantly through the P-450 system) to chloroethene oxide, which can rearrange spontaneously to chloroacetaldehyde. A major route of metabolism of chloroacetaldehyde involves oxidation to chloroacetic acid; this is either excreted as such or bound to glutathione which, after further enzymic degradation, is excreted. A number of other routes of metabolism of chloroacetaldehyde are also involved (9). 2.1.3.4 Elimination The kinetic parameters and half-lives for the elimination of vinyl chloride after inhalation and intravenous injection have been reported (7). Rats given 250 Jlg of vinyl chloride per kg of body weight by the intragastric route eliminated more than 96 '/'0 within 24 hours (3. 7 '/'~ exhaled as vinyl chloride, 12.6% as carbon dioxide, 71.5% as urinary metabolites, and 2.8% in faeces) (10). 2.1.4 Health effects
2.1.4.1 Acute and subacute toxicity The principal response to acute exposure to vinyl chloride is one of central nervous system depression. Pathological findings at necropsy include congestion and oedema of the lungs and hyperaemia of the liver and kidneys (11). 2.1.4.2 Carcinogenicity The carcinogenicity studies in animals and epidemiological observations in man have been studied and reviewed. Carcinogenic effects have been demonstrated in rats, mice, hamsters, and rabbits following ingestion or inhalation; tumours were produced at several sites, including angiosarcomas of the liver. Vinyl chloride produces angiosarcomas in the liver of man, as well as tumours of the brain, lung, and haematolymphopoietic system (1). The International Agency for
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Research on Cancer regards the evidence on vinyl chloride as sufficient to support a causal association between exposure and cancer {12). 2.1.4.3 Mutagenicity The mutagenicity of vinyl chloride and several of its metabolites has been reviewed {1). It is mutagenic in a number of biological systems, including Salmonella typhimurium, Escherichia coli K12 bioauxotropic strain, several species~of yeast, germ cells of Drosophila, and Chinese hamster V79 cells. The mutagenic action appears to be dependent on metabolic activation. 2.1.4.4 Teratogenicity Skeletal abnormalities have been observed in mice and rats exposed during gestation (13).
REFERENCES I. Some monomers, plastics and synthetic elastomers, and acrolein. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 19). 2. DRESSMAN, R. C. & McFARREN, E. F. Determination of vinyl chloride migration from polyvinyl chloride pipe into water. Journal of the American Water Works Association, 70: 29 (1978). 3. Preliminary assessment of the environmental problems associated with vinyl chloride and polyvinyl chloride. Springfield, VA, US Environmental Protection Agency, 1974 (EPA 560/4-74-001). 4. SAFE DRINKING WATER COMMITTEE. Drinking water and health. Washington, DC, National Academy of Sciences, 1977, p. 794. 5. ANON. FDA to propose ban on use of PVC for liquor use. Food chemical news, 14 May: 3-4 (1973). 6. WATANABE, P. G. ET AL. Fate of ( 14C) vinyl chloride after single dose administration in rats. Toxicology and applied pharmacology, 36: 339 (1976). 7. WITHEY, J. R. Pharmacodynamics and uptake of vinyl chloride monomer administered by various routes to rats. Journal of toxicology and environmental health, 1: 381 (1976). 8. BoLT, H. M. ET AL. Disposition of [1,2- 14C) vinyl chlonde in the rat. Archives of toxicology, 35: 153 (1976). 9. PLUGGE, H. & SAFE, S. Vinyl chloride metabolism. A review, Chemosphere, 6: 309 (1977). 10. GREEN, T. & HATHAWAY, D. E. The biological fate in rats of vinyl chloride in relation to its carcinogenicity. Chemico-biological interactions, 11: 545 (1975). II. PATTY, F. A., ed. Industrial hygiene and toxicology. Vol. II, New York, Interscience, 1963. 12. Chemicals and industrial processes associated with cancer in humans. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, suppl. I). 13. JoHN, J. A. ET AL. The effects of maternally-inhaled vinyl chloride on embryonal and foetal development in mice, rats and rabbits. Toxicology and applied pharmacology, 39: 497 (1977).
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2.2 1,1-Dichloroethene 2.2.1 General aspects Of the three isomers of dichloroethene, 1,1-dichloroethene (1,1-DCE) is the most widely used in the chemical industry. It is an intermediate in the synthesis of methylchloroform and the production of poly(viny1idene chloride) (PVDC) copolymer. PVDC polymers are used as barrier coatings in the packaging industry and Saran, a 1,1-DCE containing polymer, is widely used in the food packaging industry. 1, 1-DCE has a water solubility of 2500 mg/litre; its octanol/water partition coefficient has been reported as 5.37, indicating that it should not accumulate significantly in animals.
2.2.2 Routes of exposure 2.2.2.1 Water In the USA, the National Organics Monitoring Survey of the EPA (1) reported detecting 1,1-DCE in drinking-water, but did not quantify its occurrence. One source of 1,1-DCE could be the decomposition of 1,1,1trichloroethane which has occasionally been detected in drinking-water at concentrations of about 1 J.Lg/litre (2, 3). Dichloroethene has been found in some European groundwaters. 2.2.2.2 Food 1,1-DCE copolymer food wrappers find extensive use but unfortunately no data are available on the extent to which the unreacted monomer migrates into the wrapped food. The possibilities of other human contact via the diet appear remote. 2.2.2.3 Air The major exposure via the inhalation route is occupational. The threshold limit value (TLV) is 40 mgjm 3 of air and corresponds to an exposure of 280 mg/day for workers in an industry using or manufacturing DCE (4).
2.2.3 Metabolism 2.2.3.1 Absorption On the basis of studies of related compounds, such as trichloroethene, it is assumed that virtually 100% of ingested 1,1-DCE may be systemically absorbed (5, 6).
2. 2.2.3.2 Distribution
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I67
In studies on l,I-DCE distribution in rats (7), the largest concentrations were found in the kidneys, followed by the liver, spleen, heart, and brain. Blood concentrations were high relative to tissue concentrations. Data from subcellular distribution studies suggest substantial binding of I, I-DCE metabolites to macromolecules and associations with lipids. 2.2.3.3 Biotransformation Liebman & Ortiz (8) identified the formation of chloracetic acid from 1,1-DCE. It appears that chloroethenes are metabolized through epoxide intermediates, which are reactive and may form covalent bonds with tissue macromolecules (9). In intact test animals, a large portion of systemically absorbed 1,1-DCE is metabolized. The relationship between 1,1-DCE metabolites and their toxicity is not well understood.
2.2.4 Health effects 2.2.4.1 Acute, subchronic, and chronic toxicity I, I-DCE, like other chlorinated ethenes, possesses anaesthetic properties. Kidney and liver damage in rats and guinea-pigs exposed to air containing 1,1-DCE was reported by Prendergast (10). Differences were observed between intermittent and continuous exposures at similar I,IDCE concentrations and total exposure times. Continuous exposure, at lower concentrations than intermittent exposure, produced increased mortality. Oral administration of single doses of 200-400 mg of I,IDCE per kg of body weight had strong effects on liver enzyme activities. Only one epidemiological study has been published in which workers exposed to I, I-DCE were examined (11). No abnormal findings could be associated with I, I-DCE exposure in a population of 138 workers; measured concentrations in the work places ranged from 9 to 280 mg/m 3 (time-weighted averages). 2.2.4.2 Mutagenicity I,I-DCE has been shown to be mutagenic in Salmonella typhimurium strains TA 1530 and TA 100 (12) and E. coli KI2 (13). Henschler (9) and his associates have suggested that the mutagenic and presumably carcinogenic activities of the chloroethene series are related to the unsymmetrical chlorine substitution of the respective epoxide intermediates. Such substitution would result in less stable and more reactive intermediates than those derived from symmetrically substituted epoxides. The finding of increased mutation rates in bacterial systems has not yet been confirmed in mammalian systems.
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2.2.4.3 Teratogenicity Teratogenic effects of the DCEs do not appear to have been evaluated. 2.2.4.4 Carcinogenicity Maltoni and his coworkers (14, 15) have reported the effects of inhalation exposures to 1,1-DCE. At concentrations of 100 mg/m 3 of air, 25 of 300 Swiss mice developed kidney adenocarcinomas, while no adenocarcinomas were observed in control animals. A significant increase in mammary adenocarcinomas in Swiss mice inhaling 100 mgfm 3 and in Sprague-Dawley rats exposed to 600 mg/m 3 was also observed. Lee et al. (16) observed a small increase in hepatic haemangiosarcomas in animals exposed to 1,1-DCE in a concentration of 220 mgfm 3 for 4 hours a day, 5 days a week for 7-12 months. Rampy et al. {17) exposed Sprague-Daw1ey rats to drinking-water containing 200 mg of 1,1-DCE per litre for 2 years and to 100 and 300 mg/m 3 by inhalation. He found no evidence of increased tumours in animals treated with 1, 1-DCE. In view of the demonstrated insensitivity of Sprague-Dawley rats in the Maltoni (15) study, however, these data are not considered to alter the interpretation of positive results in Swiss mice. There is some evidence that rats are in general sensitive to the carcinogenic effects of low molecular weight chlorinated hydrocarbons (18). Epidemiological studies on workers exposed to vinylidine chloride are inadequate for evaluation (19).
2.2.5 Derivation of criterion 2.2.5.1 Existing standards Existing standards in the USA are for occupational exposures via the inhalation route. The TLV as established by the American Conference of Governmental Industrial Hygienists (4) is 40 mg of 1,1-DCE per m 3 of air in working-places. This value allows for a daily exposure of 286 mg of 1,1-DC E. This standard was established on the basis of the work of Prendergast et al. (10) described above. 2.2.5.2 Carcinogenic risk limit 1,1-DCE has been shown to produce mammary tumours in both mice and rats, and kidney adenocarcinomas in mice. Additionally, 1,1-DCE has been shown to be mutagenic in the Ames assay, a qualitative indicator of carcinogenic activity. Given this information, a linear multistage extrapolation model was applied to determine a limit that
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would give a calculated risk of less than one additional case of cancer per 100 000 population assuming a daily consumption of 2 lit res of drinking-water by a 70-kg man. The value obtained was 0.3 J.lg/litre. This value is lower than that based on non-carcinogenic risks.
REFERENCES I. Statement of basis and purpose for an amendment to the national interim primary drinking water regulations on a treatment technique for synthetzc organics. Washington, DC, US Environmental Protection Agency, 1978. 2. Preliminary assessment of suspected carcinogens in drinking water. Washington, DC, US Environmental Protection Agency, 1975. 3. List of organic compounds identified in US dnnking water. Cincinnati, OH, US Environmental Protection Agency, 1978. 4. TLVs-Threshold limit values for chemzca/ substances and physical agents in the workroom environment with intended changes for /976. Cincinnati, American Conference of Governmental Industrial Hygienists. 1976. 5. McKENNA, M. J. ET AL. The fate of (' 4 C) vinylidene chloride following inhalation exposure and oral administration in the rat. Proceedings of the Society of Toxicology, 206 (1977). 6. McKENNA, M. J. ET AL. Pharmacokinetics of vinylidene chloride in the rat. Environmental health perspectives, 21: 99-106 (1977). 7. JAEGER, R. L. ET AL. 1,1-dichloroethylene hepatotoxicity: proposed mechanism of action of distribution and binding of 14C radio-activity following inhalation exposure in rats. Environmental health perspectives, 21: 113-120 (1977). 8. LEIBMAN, K. C. & ORTIZ, E. Metabolism of halogenated ethylenes. Environmental health perspectives, 21: 91-98 (1977). 9. RENSCHLER, D. Metabolism and mutagenicity of halogenated olefins-A comparison of structure and activity. Environmental health perspectives, 21: 61-64 (1977). 10. PRENDERGAST, J. A. ET AL. Effects on expenmental animals of long-term inhalation of trichloroethylene, carbon tetrachloride, I, I, 1-tnchloroethane, dichlorodiftuoromethane, and 1,1-dichloroethylene. Toxicology and applied pharmacology, 10: 270-289 (1967). II. OTT, M. G. ET AL. A health study of employees exposed to vinylidene chloride. Journal of occupational medicine, 18: 735 (1976). 12. BARTSCH, H. ET AL. Tissue-mediated mutagenicity of vinylidene chloride and 2chlorobutadiene in Salmonella typhimurium. Nature, 255: 641 (1975). 13. GREIM, H. ET AL. Mutagenicity in vitro and potential carcinogenicity of chlorinated ethylenes as a function of metabolic oxirane formation. Biochemical pharmacology, 24: 2013 (1975). 14. MALTONI, C. ET AL. Carcinogenicity bioassays of vinylidene chloride. Research plan and early results. Medzcina de /avoro, 68: 241 (1977). 15. MALTONI, C. Recent findings on the carcinogenicity of chlorinated olefins. Environmental health perspecllves, 21: I (1977). 16. LEE, C. C. ET AL. Inhalation toxicity of vinyl chloride and vinylidene chlonde. Environmental health perspectives, 21: 25 (1977). 17. RAMPY, L. W. ET AL. Interim results of a two-year toxicological study in rats of vinylidene chloride incorporated in the drinking water or administered by repeated inhalation. Envzronmenta/ health perspectives, 21: 33 (1977). 18. NATIONAL CANCER INSTITUTE. Bioassay of tetrachloroethylene for possible carcinogenicity. Washington, DC, US Department of Health, Education and Welfare. 1977 (Technical Report Series No. 13. (NIH) 77-813). 19. Some monomers, plastics and synthetic elastomers, and acrolein. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 19).
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2.3 Trichloroethene 2.3.1 (;eneral aspects
Trichloroethene (1,1,2-trichloroethene; TCE) is a clear colourless liquid, with the empirical formula C 2 HC1 3 • It is used mainly as a degreasing solvent in metal industries. TCE has also been used as a household and industrial dry-cleaning solvent, an extractive solvent in foods, and as an inhalation anaesthetic during certain short-term surgical procedures (1). The volatilization of TCE during production and use is the major source of environmental levels of this compound. TCE has been detected in air, in food, and in human tissues (2). Its detection in rivers, municipal water supplies, the sea, and aquatic organisms indicates that TCE is widely distributed in the aquatic environment (2-4). TCE is not expected to persist in surface-water because of its volatility. However, it has been found as a frequent contaminant of groundwater.
2.3.2 Routes of exposure
2.3.2.1 Water The US National Organics Monitoring Survey observed TCE in drinking-water in 4 of 112 cities in March-April 1976, in 28 of 113 cities in May-July 1976, and in 19 of the cities in November 1976 to January 1977; the mean concentrations were 11 Jlgflitre, 21 Jlgflitre, and 1.3 Jlg/litre, respectively. TCE in water may occur as a result of direct contamination or from atmospheric contamination by rainfall (2). TCE may also be formed during the chlorination of water (5, 6). 2.3.2.2 Food There is little information concerning the occurrence of TCE in foodstuffs. In England, TCE has been observed at concentrations up to 10 Jlg/kg in meats and up to 5 Jlg/kg in fruits, vegetables, and beverages (3). Packets of tea were found to contain 60 Jlg of TCE per kg. Little TCE would be expected in other foodstuffs, except in the case of ground and instant coffee and in spice extracts, when TCE has been used as a solvent. 2.3.2.3 Air By far the most significant exposures of humans to TCE are confined to a relatively small industrial population (7). Other inhalation exposures are associated with the use of cleaning fluids containing TCE, but the hazards of such exposure would be acute.
2. 2.3.3 Metabolism 2.3.3.1 Absorption
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TCE is readily absorbed by all routes of exposure. This would be predicted on the basis of its physical and chemical properties (8). Most human data concerning TCE absorption have been obtained with inhalation as the route of exposure because of interest in the compound as an industrial toxicant and its use as an anaesthetic. Absorption of TCE following ingestion has not been studied in humans. In rats, 72-85% and 10-20% of the total orally administered dose could be accounted for in expired air and urine, respectively, with less than 0.5% appearing in the faeces (9). This indicates that at least 80% (and probably more) of ingested TCE is systemically absorbed. 2.3.3.2 Distribution
The distribution of TCE in the body is as would be expected on the basis of its chemical and physical properties (8). In guinea-pigs it was observed that the concentrations in the ovaries tend to be about 50%, and in other tissues about 25 %, of the concentration observed in fat. Laham (10) demonstrated transplacental diffusion of TCE in humans. The ratio of fetal blood concentration to maternal blood concentration varies between 0.52 and 1.90. 2.3.3.3 Biotransformation
The metabolism of TCE appears to be central to its long-term deleterious effects. In a qualitative sense, the metabolism of TCE appears similar across species (11-13). The principal products of TCE metabolism measured in urine are trichloroacetaldehyde, trichloroethanol, trichloroacetic acid, and conjugated derivatives (glucuronides) of trichloroethanol (14). The metabolite trichloroethanol has been suggested as responsible for the long-term central nervous system (CNS) effects of TCE inhalation (15). In terms of reported carcinogenic and mutagenic effects of TCE, the metabolic pathway, rather than the final products of the pathway, is of paramount importance. The essential feature of the pathway is the formation of a reactive epoxide, trichloroethene oxide, which can alkylate nucleic acids and proteins (9, 16-19). Such covalent binding can be increased with epoxide hydrase inhibition (16). 2.3.3.4 Elimination
TCE and its metabolites are excreted in exhaled air, urine, sweat, faeces, and saliva (12, 13). TCE is lost from the body with a half-time of about 1.5 hours (20). Trichloroacetic acid, trichloroethanol, and the glucuronide of trichloroethanol are excreted more slowly. The biological half-life measured in urine in humans has ranged from 12 to 50 hours for trichloroethanol and 36 to 73 hours for trichloroacetic acid (15, 21).
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2.3.4 Health effects 2.3.4.1 Acute, subchronic, and chronic toxicity Classically, TCE is known as a central nervous system depressant. In fact, the compound has been used medically as a general anaesthetic (22). Although the predominant feature of the clinical picture is the direct CNS depression produced by high exposures to TCE, there is evidence of longer term CNS effects resulting from TCE exposure (23). Fatal hepatic failure has been observed following use of TCE as an anaesthetic. Such failure has generally involved patients with complicating diseases, such as malnutrition, toxaemias, or burns, or who had received transfusions (22). Liver failure in experimental animals is marked by generalized binding of TCE metabolites to proteins and nucleic acids (17). Renal failure has been an uncommon problem with TCE anaesthesia (22). Although depressed kidney function can be documented with TCE in experimental animals, it requires very high doses (24) and the effect is much less potent than that observed with chloroform or carbon tetrachloride. Renal damage has been reported in fatal cases involving TCE abuse (1). 2.3.4.2. Mutagenicity TCE has been reported to possess mutagenic activity in a number of bacterial strains. Greim et al. (25) demonstrated reverse mutations in E. coli K12 when coupled with phenobarbital-induced mouse-liver microsomes at a concentration of 3.3 mmoljlitre in the incubation media. In the presence of Aroclor-1254-induced rat-liver microsomes or B6C3-F1 mouse-liver microsomes, TCE increased the S. typhimurium revertant rate (26). Similar observations have been made in the yeast Saccharomyces cerevisiae (strain XV 185-14C) (27). There is some doubt as to the mutagenicity of TCE, however. On chemical analysis, technical grade TCE was found to contain epichlorohydrin and epoxibutane, two compounds that Henschler (28) observed to be more potent mutagens than TCE inS. typhimurium (TA 100). Pure TCE was weakly mutagenic. These investigators concluded that the mutagenic activity, formerly attributed to TCE, was probably due in part to mutagenic contaminants found in some samples of TCE. TCE has been uniformly negative in mutagenicity testing in the absence of metabolic activation (25-27), which would suggest that the two direct-acting compounds identified could not alone account for the activity. 2.3.4.3 Teratogenicity Exposure of mice and rats to TCE in air at a concentration of 1600mg/m 3 on days 6-15 of gestation for 7 hours a day did not produce teratogenic effects (29). Although not statistically significant,
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there was evidence of haemorrhages in the cerebral ventricles (2/12 litters). A few cases of undescended testicles (2/12 litters) were observed in the TCE-treated mice but the incidence was very low. This appears to be the only teratogenesis study conducted with TCE. 2.3.4.4 Carcinogenicity In the USA, the National Cancer Institute (30) observed increased incidence of hepatocellular carcinoma in mice (strain B6C3-F1) treated with TCE. The time-weighted doses administered for 5 days/week for 78 weeks were 1169 and 2339 mgjkg of body weight for males and 869 and 1739 mgjkg of body weight for females. Similar experiments with Osborne-Mendel rats failed to increase the incidence of tumours in this species. However, the rats also responded poorly in a control experiment with carbon tetrachloride, indicating that the B6C3-Fl mouse is a much more sensitive test animal than the rat to induction of carcinomas by chlorinated compounds. The data obtained from mice are summarized in Table I. Some evidence of metastasis of hepatocellular carcinomas to the lung was observed in male mice treated with both low and high doses of TCE (4/50 and 3j48, respectively). Table 1. Incidence of hepatocellular carcinoma in TCE-treated B6C3-F1 mice (30) Males Control Low dose High dose Females
1/20 26/50 31j48
0!20 4/50 11/47
TCE has been shown to induce transformation in a highly sensitive in vitro Fischer rat embryo cell system (Fl706), which is used for identifying carcinogens. At a concentration of I mol/litre, TCE-induced transformation of rat embryo cells was characterized by the appearance of progressively growing foci made up of cells lacking contact inhibition, and by the growth of macroscopic foci when inoculated in semisolid agar. The transformed cells grew as undifferentiated fibrosarcomas at the site of inoculation in 100% of newborn Fischer rats between 27 and 68 days after inoculation (31). It has been pointed out that the TCE used in the National Cancer Institute bioassay (30) contained traces of the monofunctional alkylating agents epichlorohydrin and epoxibutane as stabilizers (32). However, TCE was also shown to induce cell transformation in the Fischer rat embryo cell system (31). In these latter experiments the TCE used was stated to be 99.9% pure. TCE has also been shown to form covalent bindings with cellular macromolecules (16-19). In the majority of cases, covalent binding of TCE was disassociated from impurities by the use of radioactively labelled TCE (18). Such covalent binding properties are commonly associated with chemical carcinogens. In any event, it appears
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unlikely that epichlorohydrin or epoxibutane would account for liver tumours produced by TCE, since their instability in aqueous solution would militate against their responsibility for tumours remote from the site of application. The carcinogenicity of pure TCE was investigated by Renschler (33). Although it is neither desirable, nor the general practice, to estimate carcinogenic potency for chemicals whose carcinogenic activity has been demonstrated in only one animal species, the NCI bioassay is the only long-term study of TCE toxicity by the oral route of exposure. In particular, the development of liver tumours following exposure to chemicals that are hepatotoxic raises questions as to the mechanism by which such tumours are produced-whether it is one of initiation or promotion. Models used to estimate potency are based on the somatic mutation theory of chemical carcinogenesis, which postulates a genetic basis for tumour initiation. There are no extrapolation models appropriate for tumour promoters. Because there is a genuine need for guidance relating to acceptable levels of TCE, especially in contaminated groundwater, a conservative approach was taken and the indicated carcinogenic hazards associated with exposure to TCE were estimated from its demonstrated carcinogenicity in B6C3-Fl mice by means of a linear extrapolation of risk using a multistage model (30). This calculation, based on a consumption of 2 litres of water per day for a 70kg man and on an acceptable risk level of less than one additional case of cancer per 100 000 population for a lifetime of exposure, results in a tentative recommended value of 30 Jlg/litre for TCE. Epidemiological studies on workers exposed to TCE were inadequate for evaluation (3437).
REFERENCES I. HUFF, J. E. New evidence on the old problems of trichloroethylene. Industrial medicine, 40: 25 (1971 ). 2. PEARSON, C. R. & McCONNELL, G. Chlorinated C 1 and C 2 hydrocarbons in the marine environment. Proceedings of the Royal Society of London, Series B, 189: 305 (1975). 3. McCoNNELL, G. ET AL. Chlorinated hydrocarbons and the environment. Endeavour, 34: 13 (1975). 4. Statement of basis and purpose for an amendment to the national primary drinking water regulations on a treatment technique for synthetic organics. Washington, DC, US Environmental Protection Agency, 1978. 5. NATIONAL RESEARCH COUNCIL. Drinking water and health. Washington, DC, National Academy of Sciences., 1977. 6. BELLAR, T. A. ET AL. The occurrence of organohalides in chlorinated drinking waters. Journal of the Amerzcan Water Works Association, 66: 703 (1974). 7. FISHBEIN, L. Industrial mutagens and potential mutagens. I. Halogenated aliphatic derivatives. Mutation research, 32: 267 (1976). 8. GoLDSTEIN, A. ET AL. The absorption, distribution, and elimination of drugs. In: Principles of drug action: the basis of pharmacology. 2nd ed. New York, John Wiley and Sons, 1974, pp. 129-154. 9. DANIEL, J. W. The metabolism of C1-labelled trichloroethylene and tetrachloroethylene in the rat. Biochemical pharmacology, 12: 795 (1963).
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10. LAHAM, S. Studies of placental transfer of trichloroethylene. Industrial medicine, 39: 46 (1970). II. IKEDA, M. & OHTSUJI, H. A comparative study of the excretion of Fujiwara reactionpositive substances in urine of humans and rodents given trichloro- or tetrachloroderivatives of ethane and ethylene. British Journal of industridt medicine, 29: 99 (1972). 12. KIMMERLE, G. and EBEN, A. Metabolism, excretion and toxicology of trichloroethylene after inhalation. I. Experimental exposure on rats. Archives of toxicology, 30: 115 (1973). 13. KIMMERLE, G. & EBEN, A. Metabolism, excretion and toxicology of trichloroethylene after inhalation. 2. Experimental human exposure. Archives of toxicology, 30: 127 (1973). 14. IKEDA, M. ET AL. Urinary excretion of total trichloro-compounds, trichloroethanol, trichloroacetic acid, as a measure of exposure to trichloroethylene and tetrachloroethylene. British journal of industrial medicine, 29: 46 (1970). 15. ERTLE, T. ET AL. Metabolism of trichloroethylene in man. I. The significance of trichloroethanol in long-term exposure conditions. Archives of toxiciology, 29: 171 (1972). 16. VAN DUUREN, B. L. & BANERJEE, S. Covalent interaction of metabolites of the carcinogen trichloroethylene in rat hepatic microsomes. Cancer research, 36: 2419 (1976). 17. BOLT, H. M. & FILSER, J. G. Irreversible binding of chlorinated ethylenes to macromolecules. Environmental health perspectives, 21: 107 (1977). 18. UEHLEKE, H. & POPLAWSKI-TABARELLI, S. Irreversible binding of 14C-labelled trichloroethylene in mice liver constituents in vivo and in vitro. Archives of toxicology, 37: 289 (1977). 19. ALLEMAND, H. ET AL. Metabolic activation of trichloroethylene into a chemically reactive metabolite toxic to the liver. Journal of pharmacology and experimental therapeutics, 204: 714 (1978). 20. STEWART, R. D. ET AL. Observations on the concentrations of trichloroethylene in blood and expired air following exposure to humans. American Industrial Hygiene Association journal, 23: 167 (1962). 21. IKEDA, M. & IMAMURA, T. Biological half-life of trichloroethylene and tetrachloroethylene in human subjects. lnternationales Archiv fiir Arbeitsmedizin, 31: 209 (1973). 22. DEFALQUE, F. J. Pharmacology and toxicology of trichloroethylene. A critical review of the world literature. Clinical pharmacology and therapeutics, 2: 665 (1961). 23. GRANDJEAN, E. ET AL. Investigations into the effects of exposure to trichloroethylene in mechanical engineering. British journal of industrial medicine, 12: 131 (1955}. 24. KLAASEN, C. D. & PLAA, G. L. Relative effects of various chlorinated hydrocarbons on liver and kidney function in dogs. Toxicology and applied pharmacology, 10: 119 (1967). 25. GREIM, H. ET AL. Mutagenicity in vitro and potential carcinogenicity of chlorinated ethylenes as a function of metabolic oxirane formation. Biochemical pharmacology, 24: 2013 (1975). 26. SIMMON, V. F. ET AL. Mutagenic activity of chemicals identified in drinking-water. In: Scott, D. et al., ed., Progress in genetic toxicology. Amsterdam, Elsevier/North Holland Biomedical Press, 1977, pp. 249-258. 27. SHAHIN, M. & VON BARSTEL, R. Mutagenic and lethal effects of benzene hexachloride, dibutylphthalate and trichloroethylene in Saccharomyces cervisiae. Mutation research, 48: 173 (1977). 28. HENSCHLER, D. Metabolism of chlorinated alkenes and alkanes as related to toxicity. Journal of environmental pathology and toxicology, 1: 125 (1977). 29. SCHWETZ, B. A. ET AL. The effect of maternally inhaled trichloroethylene, perchloroethylene, methyl chloroform and methylene chloride on embryonal and fetal development in mice and rats. Toxicology and applied pharmacology, 32: 84 (1975). 30. NATIONAL CANCER INSTITUTE, Carcinogenesis bioassays of trichloroethylene. Washington, DC, US Department of Health, Education and Welfare, 1976 (CAS No. 79-01-6, NCI-CG-TR-2).
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31. PRICE, P. J. ET AL. Transforming activities of trichloroethylene and proposed industrial alternatives. In vitro, 14: 290 (1978). 32. RENSCHLER, D. ET AL. Carcinogenicity of trichloroethylene: fact or artifact? Archives of toxicology, 37: 233 (1977). 33. RENSCHLER, D. ET AL. Carcinogenicity study of trichloroethylene by long term inhalation in three animal species. Archives of toxicology, 43: 237 (1980). 34. Some halogenated hydrocarbons. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 20). 35. ToLA, S. ET AL. A short study on workers exposed to trichloroethylene. Journal of occupatwnal medicine, 22: 737 (1979). 36. BLAIR, A. Mortality among workers in the metal polishing and plating industry. Journal of occupational medicine, 22: 158 (1980). 37. BLAIR, A. & MASON, T. I. Cancer and mortality in United States counties with metal plating industries. Archives of environmental health, 35: 92 (1980).
2.4 Tetrachloroethene 2.4.1 General aspects Tetrachloroethene ( l, 1,2,2-tetrachloroethene, perchloroethylene, PCE)a is a colourless, nonflammable liquid used primarily as a solvent in the dry-cleaning industries. It is used to a lesser extent as a degreasing solvent in metal industries (1). Tetrachloroethene is widespread in the environment, and is found in trace amounts in water, aquatic organisms, air, foodstuffs, and human tissue (2). The highest environmental levels of PCE are found in the commercial dry-cleaning and metal-degreasing industries (3). Although PCE is released into water via aqueous effluents from production plants, consumer industries, and household sewage, its level in ambient water is reported to be minimal owing to its high volatility. PCE has been detected at levels below 1 JLg/litre in water surrounding chlorinated hydrocarbon production plants in England (4) and in surface-water in the USA (5). However, it is frequently found at high concentrations in contaminated groundwater.
2.4.2 Routes of exposure 2.4.2.1 Water The US National Organics Monitoring Survey (6) detected PCE in 9 of 105 samples of drinking-water between November 1976 and January 1977 (range, 0.2-3.1 JLg/litre). The mean concentration of the nine positive samples was 0.81 JLg/litre. In Switzerland, PCE concentrations as high as 954 JLg/litre have been found in contaminated groundwater ( 7). PCE was one of two halogenated compounds identified both in drinking-water and in the plasma of individuals living in New Orleans • Perchloroethylene-abbreviated to PCE-1s the old name for tetrachloroethene. Although the modem name has been adopted here, the abbreviatiOn PCE IS retained to avmd confusion with the abbreviation for tnchloroethene.
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(8). In the United Kingdom, municipal waters have been found to contain up to 0.38/lg of PCE per litre (4).
2.4.2.2 Food PCE concentrations in seafood collected from the Liverpool Bay area in England ranged from 0.5 to 30/lg/kg (4, 9). PCE concentrations in foodstuffs ranged from non-detectable amounts ( < 0.0 l11g/kg) in orange juice to 13/lg/kg in English butter (2). 2.4.2.3 Air General environmental PCE concentrations tend to be low. Pearson & McConnell (4) observed concentrations in city atmospheres in Great Britain to range from less than 0.68 to 6811g/m 3 . In a suburb of Munich, Loechner (10) found concentrations of 4/lg/m 3 whereas air in the centre of Munich contained 6!lg/m 3 . Surveys at 8 locations in the USA indicated concentrations up to 6.7 !lg/m 3 in urban areas but less than 0.013!lg/m 3 in rural areas (11). As with a number of related chlorinated hydrocarbon solvents of low relative molecular mass, by far the most significant exposure to PCE is in industrial environments (12). The major uses of PCE are in the textile and dry-cleaning industries (69 %), metal cleaning (16 %), and as a chemical intermediate (12 %). 2.4.3 Metabolism 2.4.3.1 Absorption Using inhalation exposure, Stewart et al. (13) found that PCE reached near steady-state levels in the blood of human volunteers after 2 hours of continuous exposure. Such results suggest a rapid attainment of steady-state levels of PCE within the body. Absorption of PCE by the oral route has not been specifically studied. However, there is every reason, based on analogy with other chemicals with similar properties (e.g., trichloroethene, chloroform), to believe that it would be completely absorbed from the gastrointestinal tract. 2.4.3.2 Biotransformation The metabolism of PCE has been studied extensively in humans and experimental animals. In a qualitative sense, metabolic products appear to be similar in humans (14, 15) and experimental animals (16- 18). Basically it is believed that PCE is metabolized through an epoxide (tetrachloroethene oxide) and an acid chloride intermediate (trichloroacetyl chloride) to form the end product of its metabolism, trichloroacetic acid. Ogata et al. (19) report that 1.8% of PCE retained by humans was converted to trichloroacetic acid and 1.0% to an unknown metabolite, in 67 hours.
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2.4.3.3 Elimination PCE itself is primarily eliminated from the body via the lungs (13, 20, 21). The half-time for respiratory elimination of PCE has been estimated at 65 hours (20, 21). Trichloroacetic acid, as a metabolite of PCE, is eliminated with a half-time of 144 hours via the urine (21).
2.4.4 Health effects 2.4.4.1 Acute, subchronic, and chronic toxicity As with all other members of the chloroethene family, acute effects of PCE are very much dominated by central nervous system depression. The only indications of long-term effects on the central nervous system are changes in EEG patterns associated with increased electrical impedance of the cerebral cortex at exposures as low as 100 mg of PCE per m 3 of air for 4 hours a day for 15-30 days (22, 23). These effects were reported to be associated with sporadic swollen and vacuolized protoplasm in some cells (23). Although only limited information is available from experimental animals, it generally supports findings of acute central nervous system depression (24). As in the case of human clinical studies, little information is available concerning long-term exposures. That more serious central nervous system problems may be associated with chronic PCE exposure in humans is suggested by a few sporadic case reports (25, 26) and small-scale epidemiological and clinical studies (27) involving a group of men occupationally exposed to concentrations of 1890-2600mg of PCE per m 3 . However, these studies have often been complicated by exposures to other solvents (28). Short-term PCE exposures at higher concentrations, and longer exposures at lower concentrations, can produce damage to the kidneys and liver in dogs (29). Kylin et al. (30) noted moderate fatty degeneration of the liver in mice following a single 240-minute exposure to 1340mg of PCE per m 3 . Exposure to this same concentration 4 hours daily, 6 days a week for up to 8 weeks was found to increase the severity of the lesions caused by PCE (31). Epidemiological studies on workers exposed to PCE are inadequate for evaluation (32-34). 2.4.4.2 Mutagenicity Cerna & Kypenova (35) found that treatment with PCE resulted in elevated mutagenic activity in Salmonella strains sensitive to both base substitution and frameshift mutation. However, PCE had no effect on the spontaneous mutation rate in E. coli Kl2 in the presence of liver microsomes (36). 2.4.4.3 Carcinogenicity PCE has been demonstrated to be a liver carcinogen in B6C3-Fl mice (37). Results in Osborne-Mendel rats were negative, but a high rate of
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early mortality precluded the use of rat data in evaluating the carcinogenicity of PCE. Furthermore, recent studies in which carbon tetrachloride was used as a positive control revealed that OsborneMendel rats have a low sensitivity to induction of hepatocellular carcinoma by chlorinated hydrocarbons (38). Carcinogenic hazards associated with exposure to PCE may be estimated on the basis of its production of hepatocellular carcinoma in B6C3-F1 mice (37) using a linear multistage extrapolation model. Such a model is based on the somatic mutation theory of chemical carcinogenesis most appropriate for tumour initiators. However, chemicals that produce necrotic damage to tissues such as the liver are known to enhance tumorigenesis by epigenetic mechanisms (39). Extrapolation models for such effects do not exist. There was some reluctance to base an estimate of carcinogenic hazards on data from experiments utilizing hepatotoxic doses in the absence of other evidence of carcinogenicity. However, the lack of other suitable data and the genuine need for guidance for PCE-contaminated groundwater have prompted a conservative approach and PCE was treated as if it were a tumour initiator in order to derive a tentative recommended value. The results of the calculation indicate that a concentration in drinking-water of 10 Jlg of tetrachloroethene per litre would be predicted to increase the disease risk by less than one additional cancer per 100 000 population assuming a daily consumption of 2 litres of water.
REFERENCES I. WINDHOLZ, M. ET AL The Merck index. 9th ed. Rahway, NJ, Merck and Co., 1976. 2. McCoNNELL, G. ET AL. Chlorinated hydrocarbons and the environment. Endeavour, 34: 13-18 (1975). 3. NATIONAL INSTITUTE OF OCCUPATIONAL SAFETY AND HEALTH. Criteria for a recommended standard-occupational exposure to tetrachloroethylene (perch/oroethylene). Washington, DC, US Department of Health, Education & Welfare, 1976. 4. PEARSON, c. R. & McCONNELL, G. Chlorinated c, and c2 hydrocarbons in the marine environment. Proceedings of the Royal Society, Series B, 189: 305 (1975). 5. EWING, B & CHIAN, E. Monitoring to detect previously unrecognized pollutants in surface waters. Washington, DC, US Environmental Protection Agency, 1977. 6. Statement of basis and purpose for an amendment to the national primary drinking water regulations on a treatment cnteria for synthetic organics. Washington, DC, US Environmental Protection Agency, 1978. 7. GIGER, W. & MoLNAR-KUBIEA, E. Tetrachloroethylene in contaminated ground and drinking waters. Bulletin of environmental contamination and toxicology, 19(4): 475 (1978). 8. DowTY, B. ET AL. Halogenated hydrocarbons in New Orleans drinkmg water and blood plasma. Science, 187: 75 (1975). 9. DICKSON, A. G. & RILEY, J. P. The distribution of short-chained halogenated aliphatic hydrocarbons in some marine organisms. Marine pollution bulletin, 7: 167 (1976). 10. LOECHNER, F. Perchloriithylen eine Bestandsaufnahme. Umwelt, 6: 434 (1976). II. LILLIAN, D. ET AL Atmospheric fates of halogenated compounds. Environmental science and technology, 9: I 042 (1975).
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12. FISHBEIN, L. Industrial mutagens and potential mutagens I. Halogenated aliphatic hydrocarbons. Mutation research, 32: 267 (1976). 13. STEWART. R. D. ET AL. Human exposure to tetrachloroethylene vapor. Archwes of environmental health, 2: 516 ( 1961 ). 14. IKEDA, M. ET AL Urinary excretion of total trichloro-compounds, tnchloroethanol and trichloroacetic acid as a measure of exposure to trichloroethylene and tetrachloroethylene. British journal of industrial medicine, 29: 328 (1972). 15. IKEDA, M. Metabolism of trichloroethylene and tetrachloroethylene m human subjects. Environmental health perspectit'es, 21: 239 (1977). 16. YLLNER, S. Urinary metabolites of 14 C-tetrachloroethylene in mice. Nature, 191: 820 (1961). 17. DANIEL, J. W. The metabolism of 36 CI-Iabelled trichloroethylene and tetrachloroethylene in the rat. Bwchemical pharmacology, 12: 795 (1963). 18. KEDA, M. & 0HTSUJI, H. A comparative study of the excretion of Fujiwara- reactionpositive substances in urine of humans and rodents given trichloro- or tetrachloroderivatives of ethane and ethylene. British Journal of industrial medicine, 29: 99 (1972). 19. OGATA, M. ET AL Excretion of organic chlorine compounds in the urine of persons exposed to vapours of trichloroethylene and tetrachloroethylene. British journal of mdustrial medtcine, 28: 386 (1971). 20. STEWART, R. D. ET AL Experimental human exposure to tetrachloroethylene. Archit'es of environmental health, 20: 225 (1970). 21. IKEDA, M. & IMAMURA, T. Biological half-life of trichloroethylene and tetrachloroethylene in human subjects. Internationales Archiv fiir Arbeitsmedizin, 31: 209 (1973). 22. DMITRIEVA, N. V. Maximum permissible concentratiOns of tetrachloroethylene in factory air. Hygiene and samtation, 31: 387 (1966). 23. DMITRIEVA, N. V. & KuLESHOV, E. V. Changes m the bioelectric activity and electric conductivity of the brain in rats chronically poisoned with certain chlorinated hydrocarbons. Hygtene and sanitation, 36: 23 (1971). 24. Some halogenated hydrocarbons. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of the carcmogenic risk of chemicals to humans, vol. 20.) 25. GoLD, J. H. Chronic perchloroethylene poisoning. Canadian Psychwtric Associatwn Journal, 14: 627 (1969). 26. MCMULLEN, J. K. Perchloroethylene intoxication. British medtcal journal, 2: 1563 (1976). 27. CoLER, H. R. & RossMILLER, H. R. Tetrachloroethylene exposure m a small industry. Archives of industrial hygtene & occupational medicine, 8: 227 (1953). 28. TuTTLE, T. C. ET AL A behavioral and neurological evaluation of dry cleaners exposed to perchloroethylene. Washington, DC, Department of Health, Education and Welfare, 1977 ((NIOSH) No. 77-214). 29. KLAASEN, C. D. & PLAA, G. L. Relative effects of chlorinated hydrocarbons on liver and kidney function in dogs. Toxicology and applied pharmacology, 10: 119 (1967). 30. K YLIN, B. ET AL Hepatotoxicity of inhaled trichloroethylene, tetrachloroethylene and chloroform. Smgle exposure. Acta pharmacologica et toxicologica, 20: 16 (1963). 31. K YLIN, B. ET AL. Hepatotoxicity of mhaled trichloroethylene. Long-term exposure. Acta pharmacologica et toxicologica, 22: 379 (1965). 32. BLAIR, A. ET AL Causes of death among laundry and dry cleanmg workers. American journal of publtc health, 69: 508 (1979). 33. BLAIR, A. Mortality among workers in the metal pohshing and platmg industry. Journal of occupational medicme, 22: 158 (1980). 34. BLAIR, A. & MASON, T. J. Cancer and mortality in the United States counties with metal platmg industries. Archives of environmental health, 35: 92 (1980). 35. CERNA, M. & K YPENOVA, H. Mutagenic activity of chloroethylenes analysed by screening system tests. Mutation research, 46: 214 (1977). 36. GREIM, H. ET AL Mutagenicity in vitro and potential carcmogenicity of chlorinated ethylenes as a function of metabolic ox1rane formation. Biochemical pharmacology, 24: 2013 (1975).
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37. NATIONAL CANCER INSTITUTE. Bioassay of tetrachloroethylene for possible carcinogenicity. Washington, DC, Department of Health, Education and Welfare, 1977 (CAS No. 127-18-4, NCI-CG-TR-13 (NIH) 77-813). 38. NATIONAL CANCER INSTITUTE. Carcinogenesis bioassay of trichloroethylene. Washington, DC, Department of Health, Education and Welfare, 1976 (CAS No. 7901-6, NCI-C6-TR-2 (NIH) 76-802). 39. ScHUMANN, A. ET AL. The pharmacokinetics and macromolecular interactions of perchloroethylene in mice and rats as related to oncogenicity. Toxicology and applied pharmacology, 55: 207 (1980).
3. POLYNUCLEAR AROMATIC HYDROCARBONS 3.1 General description Polynuclear aromatic hydrocarbons (PAH) are a large group of organic compounds consisting of two or more benzene rings with nonaromatic rings present in some instances. Adjacent rings share two carbon atoms.
Naphthalene
Fluoranthene
Benzo [a ]pyrene
PAH are formed as a result of incomplete combustion of organic compounds but may also be synthesized by some bacteria, algae, and higher plants (1, 2). They are of relatively low solubility in water and are strongly adsorbed on to particulates and sediment clays (3), leading to higher concentrations in the water environment than would otherwise be possible on the basis of solubility considerations alone (1). PAH dissolved in water and adsorbed on particulate matter undergo photodecomposition provided that exposure to ultraviolet light from solar radiation of sufficient energy occurs (4, 5). Some microorganisms in soil can degrade PAH (6) and it is probable that at least some degradation of PAH in sediments also occurs (7). 3.2 Routes of exposure PAH are present in the environment from both natural and anthropogenic sources. As a group, they are widely distributed in the environment, having been detected in animal and plant tissue:>, sediments, soils, air, and various water sources (8, 9). 3.2.1 Air PAH concentrations in air vary with location and season. Industrial urban environments have higher PAH concentrations and these tend to 182
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be at a maximum in the colder months when more heating is used (10). Concentrations of PAH vary from about 0.001 J.lg to 2J.1g per 1000m 3 of air in rural areas to about 0.1 J.lg to 60 J.lg per 1000 m 3 in urban areas (I 1). In some polluted atmospheres, levels of benzo [a ]pyrene varying from 20 to 400 J.lg per 1000 m 3 have been reported (I 2).
3.2.2 Food PAH are found in substantial quantities in some foods, depending on the method of cooking, preservation, and storage, and are detected in a wide range of meat, fish, vegetables, and fruits (13). American sources indicate a total intake of PAH from food of the order of 1.6-l6J.1g/day (14).
3.2.3 Water Levels of PAH in surface-waters are influenced by industrial discharges and in various German rivers have been shown to range from 0.12 to 3.1 J.lg/litre (15). Examination of a number of groundwater and drinking-water supplies for six PAH, namely fluoranthene, benzo[b]fluoranthene, benzo [k ]fluoranthene, benzo [a ]pyrene, benzo [ghi]perylene and indeno [1 ,2,3-cd]pyrene, indicated that the collective concentrations generally did not exceed 0.05 J.lg/litre in groundwater and 0.1 J.lg/litre in drinking-water (16). Later work by BornefT (17) with data from a large number of analyses showed concentrations of these six PAH to be greater than 0.11 J.lg/litre in only 1 % of samples, while for 90% of samples levels were between 0.001 and 0.01 J.lg/litre. Concentrations of benzo[a]pyrene were shown to range from 0.6 to 114ng/litre in surfacewaters and from 0.1 to 23.4 ng/litre in drinking-water (18, 16). According to BornefT (19), two-thirds of the PAH in surface-waters are bound to particulate matter which is removed by sedimentation, flocculation, and filtration processes. Much of the remaining one-third of dissolved PAH may be removed by oxidation, the efficiency varying with the system used. Chlorination of water can remove 50-60% of benzo [a ]pyrene, while filtration with activated carbon removed up to 99% of benzo[a]pyrene in field tests (19). Contact with coal-tar-based pipe linings during distribution is known in some instances to lead to increases in PAH concentrations in the water. In such cases, an increase in the level of fluoranthene is particularly marked (20). Estimates of annual intake of PAH are between 1 mg and lOmg, with benzo[a]pyrene contributing 0.1-1.5 mg per person in developed countries (21, 22). However, it has been estimated that food accounts for about 99% of the daily oral intake and the typical intake from drinkingwater is only 0.1 % of the total (23).
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3.3 Metabolism PAH are highly lipid-soluble and are readily absorbed from the gut and lungs of mammals (24-26). They are initially cleared from the blood and liver very rapidly (27) and are distributed to a wide variety of tissues with a particular tendency to localize in body fat (26-29). Metabolism of PAH is via the mixed-function oxidase system mediated by the cytochrome P450, with oxidation or hydroxylation .1s the first step. The resultant epoxides or phenols may then go through a detoxification reaction to produce glucuronides, sulfates, qr glutathione conjugates. Some of the epoxides, however, may be metabolized to dihydrodiols, which in turn may undergo conjugation to form soluble detoxification products or oxidation to diol-epoxides (30, 31). These latter compounds are thought to be the ultimate carcinogens in cases where carcinogenicity has been demonstrated (32). The metabolites of PAH are primarily eliminated in the urine and faeces as water-soluble compounds (33, 34). PAH can induce the synthesis of the enzymes responsible for its own metabolism (35). This in turn means that the efficiency of metabolism of PAH will increase with continued exposure. However, the effects of various exogenous and endogenous inducers, genetic factors, age, sex, and nutritional status, which have been shown to influence these enzyme systems, make it difficult to predict whether the carcinogenicity of PAH will be increased or decreased with time. Numerous studies show that despite their high lipid solubility, PAH show little tendency to undergo bioconcentration in the fatty tissues of animals or man, primarily because PAH are rapidly and extensivdy metabolized (34, 36, 37).
3.4 Health effects Little information is available on the acute, subacute, and chronic toxicity of PAH after ingestion. They have, however, been shown to cause hyperkeratosis, hyperplasia, and loss of sebaceous glands in the skin (38). Some, such as dimethylbenzanthracene, have been shown to exert marked effects on the bone marrow and lymphoid tissue in rats (39). However, overt signs of toxicity are not usually produced by carcinogenic PAH until the dose is sufficient to produce tumours. Only very limited data are available on the teratogenic effects of PAH, but the data on benzo[a]pyrene indicate that effects occur only at comparatively high doses (40, 41). A number of PAH have been shown to be mutagenic in bacterial systems, in vitro cell lines, and in vivo by sister chromatid exchange (4 245). The significance or impact of such changes on human populatiom is not clear, however, though there is a very good correlation between in vitro tests for mutagenesis and carcinogenic potential. Many PAH are capable of producing tumours in skin and other
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epithelial tissues in numerous test species (13, 46). Malignant tumours can often be induced by very small quantities of the test compound and with very short latency periods. The carcinogenesis of PAH by the oral route of administration has been less well studied but positive results have been obtained, particularly as regards gastrointestinal and oesophageal tumours (13, 47). PAH are rarely encountered singly in the environment and many interactions can occur with mixtures of P AH whereby the potency of known carcinogenic PAH may be enhanced (48-50). These systems are not well understood, however, and their significance as regards environmental exposure to PAH is not yet clear. Epidemiological studies of occupation-associated skin cancers in man provide strong presumptive evidence of the positive role of PAH in certain kinds of human cancer (51-53). However, this provides no quantitative data regarding environmental as opposed to occupational exposure or regarding the importance of mixtures of PAH. 3.4.1 Guideline value
Based on the premise that drinking-water should be comparable in quality with unpolluted groundwater, the drinking-water standards published by WHO in 1970 and 1971 set a limit of 200 ngjlitre for the sum of six indicator PAH in drinking-water. These were ftuoranthene, benzo [a ]pyrene, benzo [ghi]perylene, benzo [b ]fluoranthene, benzo [k ]fluoranthene, and indeno[l ,2,3-cd]pyrene. Concentrations of these indicator PAH were found to be 10-50 ngjlitre in groundwater and 50-250 ng/litre in relatively unpolluted river-water, with higher levels in polluted rivers and effluents. Subsequent studies showed that the levels of these PAH in drinkingwater were considerably lower than the standard and that the concentration was markedly influenced by fluoranthene leached from coal-tar-lined distribution systems. In addition, the representative PAH chosen and the limit set were not based on any toxicological considerations. In 1976, a limit of 5 ng/litre for benzo[a]pyrene in surface-water in the USSR was proposed and approved by the Ministry of Health of the USSR (~4). Though the relative contribution of PAH in drinking-water to the total intake of PAH is small, these are potentially hazardous substances and exposure should be minimized. Insufficient information is currently available, however, to set values for mixtures of PAH or individual compounds, except in the case of benzo[a]pyrene. It is probable that a guideline value for benzo[a]pyrene will affect the levels of other PAH in drinking-water, since methods tu reduce benzo[a]pyrene to acceptable levels will result in the reduction of all PAH. The proposed guideline value for benzo[a]pyrene is based on toxicological considerations and experimental data from Neal &
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Rigdon (47) who fed benzo[a]pyrene at doses ranging between 1 and 250 mg/kg in the diet to strain CFW mice for about 110 days. There was a statistically significant, dose-related increase in the incidence of stomach tumours (papillomas and carcinomas) in this experiment. A linearized, multistage model for risk assessment, which defines acceptable risk as 1 in 100 000, was applied to these data (55). The following is a summary of the recommendations made: 1. Based on the application of the multistage model to the available toxicological data for benzo[a]pyrene and taking account of the fact that this substance is associated in water with other PAH of known carcinogenicity, a guideline value of 0.01 J.Lg/litre is proposed for benzo [a ]pyrene in drinking-water. 2. Because of the close association of PAH with suspended solids, the application of treatment, when necessary, to achieve an acceptable levd of turbidity 0 will ensure that minimum PAH levels are achieved. 3. No PAH should be added to water during water treatment and distribution; therefore, the use of coal-tar-based pipe linings should be discontinued. It is recognized that it may be impractical to remove coaltar linings from existing pipes. However, research should be undertake·n to develop methods for minimizing leaching of PAH. 4. Monitoring of PAH levels should continue with the objective of determining background levels against which any changes can be assessed and if necessary remedial action taken. For the purpose of monitoring PAH levels, the use of several specific compounds as indicators for the group as a whole is considered to be advantageous. The choice of indicator compounds will vary with the individual situation. Some authorities consider the compound benzo [a ]pyrene can serve in some instances as an index of pollution by the whole of the PAH group of compounds. 5. The control of PAH in drinking-water should continue to be bast:d on the concept that unpolluted groundwater represents a baseline contamination that should not be exceeded. This concept was embodit:d in the standards published by WHO in 1970 and 1971. This was a useful approach, but it is not applicable in all situations and is not based on toxicological considerations.
REFERENCES I. ANDELMAN, J. G. & SNODGRASS, J. E. Incidence and significance of polynuclear aromatic hydrocarbons in the water environment. CRC crlllcal reviews in environmental control, 4: 69 (1974). 2. HARRISON, R. M. ET AL. Polynuclear aromatic hydrocarbons in raw, potable and wa>te waters. Water research, 9: 331 (1975). 3. STROSHER, M. T. & HoDGSON, G. W. Polycyclic aromatic hydrocarbons in lake waters and associated sediments. Analytzcal determination by gas-chromatography m<lSS • The level of acceptability IS
defined in Part V, sectiOn 16.4, p. 310.
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4. 5. 6. 7. 8. 9. 10. II. 12. 13. 14. 15. 16. 17. 18. 19. 20. 21. 22. 23. 24. 25. 26.
spectrometry. Philadelphia, American Society for Testing and Materials, 1975 (Water Quality Parameters ASTM STP 573, pp 259-270). McGINNES, P. R. & SNOEYINK, U. L. Determination of the fate of PAH m natural water systems. Urbana, IL, Illinois University, 1974 (Research Report No. 80, NTIS PB-232, 168). Water related environmental fate of 129 priority pollutants. Washington, DC, EnvironmentaL Protection Agency 1979 (EPA-440/4-79-0296). SHABAD, L. M. On the so-called MAC (Maximum Allowable Concentrations) for carcinogenic hydrocarbons. Neoplasma, 22: 459 (1975). NEFF, J. M. Polycyclic aromatic hydrocarbons in the aquatic environment. London, Applied Science Publishers, 1979. RADDING, S. B. ET AL. The environmental fate of selected polynuclear aromatic hydrocarbons. Menlo Park, CA, Stanford Research Institute, 1976 (EDPA 560/5-75009. NTIS PB-250 948). SHACKELFORD, W. M. & KEITH, L. H. Frequency of organic compounds identified in water. Washington, DC, Environmental Protection Agency, 1976 (EPA-600/4-76-062). HANGEBRAUK, R. P. ET AL. Sources of PAH in the atmosphere. Washington, DC, Department of Health, Education and Welfare, 1967 (999-AP-33). SAWICKI, E. ET AL. Benzo[a]pyrene content of the air of American communities. American Industrial Hygiene Association journal, 21: 443 (1960). EGAN, H. & SAWYER, R. Contaminants in food: analytical problems and achievements. Toxicology, 4: 245 (1975). Certain polycyclic aromatic hydrocarbons and heterocyclic compounds. Lyon, International Agency for Research on Cancer, 1973 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 3.) SANTODONATO, J. ET AL. In: Bjorseth A. & Dennis A. J., ed. Polynuclear aromatic hydrocarbons: Chemistry and biological effects. Columbus, Ohio, Batelle Press, 1980. ANDELMAN, J. B. & SUESS, M. J. Polynuclear aromatic hydrocarbons in the water environment. Bulletin of the World Health Organization, 43: 479 (1970). BoRNEFF, J. & KuNTE, H. Kanzerogene Substanzen in Wasser und Boden. XXVI: Routinemethode zur Bestimmung von polyzyklischen Aromaten im Wasser. A.rchiv fur Hygiene und Bakteriologie, 153: 220 (1969). BORNEFF, J. Drinking-water production from surface water. Example of the lake of Zurich with regard to polycyclics, metals and chlorinated hydrocarbons. Gas, Wasser, A.bwasser, 55: 467 (1975). BORNEFF, J. & KUNTE, H. Kanzerogene Substanzen in Wasser und Boden. XVI. Nachweis von polyzyklischen Aromaten in Wasserproben durch direkte Extraktion. Archiv fur Hygiene und Bakteriologie, 148: 585 (1964). Ambient water quality cnteria for polynuclear aromatic hydrocarbons. Washington, DC, Environmental Protection Agency, 1980. CRANE, R. I. ET AL Survey of polycyclic aromatic hydrocarbon levels in British waters. Medmenham, England, Water Research Centre, 1981 (Technical Report TR 158). BoRNEFF, J. Trinkwassergewinnung aus Cber-fliichenwasser am Beispiel des Ziirichsees unter Beriicksichtigung vor Polyzyklen, Metallen und ChlorkohlenwasserstofTen. Gas. Wasser, Abwasser, 55: 467 (1975). FRITZ, W. Unfang und Quellen der Kontamination unserer Lebensmittel mit Krebserzeugenden KohlenwasserstofTen. Erniihrungsforschung, 16: 547 (1972). SHABAD, L. M. & IL'NITSKII, A. P. [The perspectives of elaborating the problem of water bodies contamination with cancerogenic substances.] G1giena i sanitan;a, 8: 84 (1970) (in Russian). REES, E. 0. ET AL. A study of the mechanism of intestinal absorption of benzo[a]pyrene. Biochimica et biophysica acta, 225: 96 (1971). VAINII, H. The fate of intratracheally instilled benzo[a]pyrene in the isolated perfused rat lung of both control and 20-methylcholanthrene pretreated rats. Research communications in chemical pathology and pharmacology, 13: 259 (1976). KOTIN, P. Distribution, retention, and elimination of 14C-3,4-benzopyrene after administration to mice and rats. Journal of the National Cancer Institute, 23: 541 (1969}.
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27. SCHLEDE, E. ET AL. Effects of enzyme induction on the metabolism and tissue distribution of benzo(alpha)pyrene. Cancer research, 30: 2893 (1970). 28. BOCK, F. G. & DAO, T. L. Factors effecting the polynuclear hydrocarbon level in the rat mammary glands. Cancer research, 21: 1024 (1961). 29. SCHLEDE, E. ET AL. Stimulatory effect of benzo(alpha)pyrene and phenobarbitol pretreatment on the biliary excretion of benzo(alpha)pyrene metabolites in the rat. Cancer research, 30: 2898 ( 1970). 30. CONNEY, A. H. ET AL. Use of drugs in the evaluation of carcinogen metabolism in man. In: Montesano R. et al., ed. Screening tests in chemical carcinogenesis. Lyon, International Agency for Research on Cancer, 1976 (IARC Scientific Publications No. 12). 31. YANG, S. K. ET AL. In: Gelbom, H. V. & Ts'o, P.O.P., ed. Polycyclic hydrocarbons and cancer. Vol. 1, Chemistry and metabolism. New York, Academic Press, 1978. 32. LEHR, R. E. ET AL. In: Freudenthal, R. I. & Jones, P. W., ed. Polynuclear aromatic hydrocarbons: 2nd International Symposium on Analysis, Chemistry and Biology. New York, Raven Press, 1978. 33. DANIEL, P. M. ET AL. Metabolism of labelled carcinogemc hydrocarbons on rats. Nature, 215: 1142 (1967). 34. WHITTLE, K. J. ET AL. Occurrence and fate of organic and inorganic contaminants in marine animals. Annals of the New York Academy of Sciences, 298: 47 (1978). 35. MARQUARDT, H. In: Mohr, U. et al., ed. A1r pollution and cancer in man. Lyon, International Agency for Research on Cancer, 1977 (IARC Scientific Publications No. 16). 36. STICH, H. F. ET AL. In: Rosenfeld, C. & Davis, W., ed. Environmental pollution and carcinogenic risks. Lyon, International Agency for Research on Cancer, 1976 (IARC Scientific Publications No. 13). 37. LEE, R. F. ET AL. Petroleum hydrocarbons: uptake and discharge by the marine mussel Mytilus edulis. Science, 177: 344 (1972). 38. BOCK, F. G. Early effects of hydrocarbons on mammalian skin. Progress in expenmental tumor research, 4: 126 (1964). 39. CARWEIN, M. J. & SNYDER, K. L. Suppression of cellular activity in the reticuloendothelial system of the rat by 7,12-dimethylbenzo[a]anthracene. Cancer research, 28: 320 (1968). 40. BULAY, 0. M. The study of development of lung and skin tumours in mice exposed in vitro to polycyclic hydrocarbons. Acta medica Tumca, 7: 3 (1970) (cited in Registry of toxic effects of chemical substances, 1978). 41. BULAY, 0. M. & WATTENBERG, L. W. Carcinogenic effect of subcutaneous admimstration of benzo[a]pyrene dunng pregnancy on the progeny. Proceedings of the Society for Experimental Biology and Medicine, 135: 84 (1970) (cited in Registry of toxic effects of chemical substances, 1978). 42. LAVoiE, E. J. ET AL. In: Jones, P. W. & Leber, P., ed. Polynuclear aromatic hydrocarbons. Ann Arbor, MI, Ann Arbor Science Publishers, 1979. 43. McCANN, J. ET AL. Detection of carcinogens as mutagens in the Salmonella/microsome test: assay of 300 chemicals. Proceedings of the National Academy of Sciences, 72: 5135 (1975). 44. HUBERMAN, E. & SACHS, L. Mutability of different genetic loci in mammalian cells by metabolically activated carcinogenic polycyclic hydrocarbons. Proceedings of the National Academy of Sciences, 73: 188 (1976). 45. BAYER, U. In: Freudenthal, R. I. & Jones, P. W., ed. Polynuclear aromatic hydrocarbons: Second International Symposium on Analysis, Chemistry and Biology. New York, Raven Press, 1978. 46. IBALL, J. Relative potency of carcinogenic compounds. American journal of cancer, 35: 188 (1939). 47. NEAL, J. & RIGDON, R. H. Gastric tumours in mice fed benzo[a]pyrene: a quantitative study. Texas reports on biology and medicine, 25: 553 (1967). 48. PFEIFFER, E. H. Investigations on the carcinogenic burden by air pollution in man. VII. Studies on the oncogenic interaction of polycyclic aromatic hydrocarbons.
3. B), 158: 69 (1973).
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Zentralblatt fiir Bakteriologie, Parasitenkunde, Infektwnskrankheiten und Hyg1ene (Orig.
49. PFEIFFER, E. H. Oncogenic interaction of carcinogenic and non-carcinogenic polycyclic aromatic hydrocarbons in mice. In: Mohr, U. et al., ed. Air pollution and cancer in man. Lyon, International Agency for Research on Cancer, 1977 (IARC Scientific Publications No. 16). 50. SCHMAHL, D. ET AL. Syncarcinogenic action of polycyclic hydrocarbons in automobile exhaust gas condensates. In: Mohr, U. et al., ed. Air pollutwn and cancer in man. Lyon, International Agency for Research on Cancer, 1977 (IARC Scientific Publications No. 16). 51. Porr, P. In: Chirurgical observations. London, Hawes, Clark and Collins, 1975. 52. ECKHARDT, R. D. Industrial carcinogens. New York, Grune and Stratton, 1959. 53. SELIKOFF; I. J. Air pollution and asbestos carcmogenesis: investigation of possible synergism. In: Mohr, U. et al., ed. Air pollution and cancer in man. Lyon, International Agency for Research on Cancer, 1977 (IARC Scientific Publications No. 16). 54. [Rules of surface waters protection from pollution by sewage and industrial wastes.] Moscow, Ministry of Health, 1976. 55. Ambient water quality crueria for polynuclear aromatic hydrocarbons. Washington, DC, US Environmental Protection Agency, 1980 (EPA 440(5-70-069).
4. PESTICIDES Pesticides that may be of importance to water quality include chlorinated hydrocarbons and their derivatives, persistent herbicides, soil insecticides, pesticides that are easily leached out from the soil, and pesticides systematically added to water supplies for disease vector control or other purposes. Of these compounds, only the chlorinated hydrocarbon insecticides occur frequently. Chlorinated hydrocarbon pesticides persist in the environment and have become ubiquitous. For example, traces of DDT have been recovered from dust known to have drifted over thousands of kilometres and from water melted from antarctic snow. Traces of chlorinated hydrocarbon pesticides in water may progressively accumulate in different steps of a food chain; for example, DDT can bioaccumulate in fish at levels more than 10 000 times the concentrations present in the water of their habitat. Several of the pesticides in this group, including ones that have been used extensively in the past in agriculture and some that continue to be used for purposes such as disease control, have been shown to possess tumorigenic properties in animals. Guideline values are recommended for several chlorinated hydrocarbon pesticides because of their known occurrence as adventitious residues in water. These values are derived from the acceptable daily intake (ADI) values set by the FAOjWHO Joint Expert Committee on Pesticide Residues; I % of the ADI has been adopted as the basis for values in drinking-water. Application of the linear multistage extrapolation model for estimating potential carcinogenic risks to the recommended values suggests that water containing pesticide residues at these levels is unlikely to result in an incremental cancer risk per lifetime exceeding one per 100 000 of the population. Short-term excursions above these values resulting, for example, from vector control operations, may be permissible but require monitoring and assessment of health implications. It is recognized that the pesticides for which values have been specified do not represent all of those that have been identified in water. Local circumstances may require the extensive use of a pesticide for which guidelines have not been developed. Monitoring for the presence of such substances in drinking-water may, therefore, be deemed desirable. The recommended values are designed to protect human health. They may not be adequate for the protection of aquatic life.
190
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4.1 DDT (total isomers) 4.1.1 General aspects DDT is an acronym for dichlorodiphenyltrichloroethane. Its full chemical name is I, I'-(2,2,2-trichloroethylidene)bis- [4-chlorobenzene]. This is also known as p,p'-DDT, since the structure of DDT permits several different isomeric forms. The term DDT is also applied to commercial products consisting predominantly of p,p'-DDT together with some o,p'-DDT and smaller amounts of p,p'-TDE, o,p'-TDE, p,p'-DDE, o,p'-DDE, and other unidentified compounds. 0 DDT is highly insoluble in water but soluble in organic solvents. Its vapour pressure is 2.53 x w-s Pa (1.9 x I0- 7 mmHg) at 20 ac. DDT was first used for the protection of military areas and personnel, mainly against malaria, typhus, and certain other vector-borne diseases during the Second World War. Widespread agricultural use dates from I946 in the USA and slightly later in most other countries. DDT was also applied extensively to forests. Its use has been restricted or even prohibited in some countries for ecological reasons and because of the increasing resistance of pests to the insecticide. A decrease in production has been observed particularly since I970. DDT is still used extensively, both in agriculture and for vector control, in some tropical countries. DDT is a persistent insecticide. It is stable under most environmental conditions and is resistant to complete breakdown by the enzymes present in soil microorganisms and higher organisms. Some of its metabolites, notably I, I'-(2,2-dichloroethenylidene)bis [4-chlorobenzene] (ODE), have a stability equal to, or greater than that of the parent compound. WHO has issued an environmental health criteria document on DDT and its derivatives (1).
4.1.2 Routes of exposure 4.1.2.I Air Evaporation from fields treated with DDT can be detected for more than six months after application. Most of the DDT returns to the soil in the area of application with an almost straight-line inverse relationship to the logarithm of the distance from the source. However, a small proportion may undergo worldwide transportation and traces of DDT have been recovered from dust known to have drifted over 1000 km. Normally, the concentration in air in non-agricultural areas is in the range I-2.36 x I0- 6 mgjm 3 . In communities with anti-mosquito fogging programmes, concentrations of DDT may be much higher. Knowledge of the circulation and fate of DDT and its analogues in the environment is generally lacking. Recently it has been demonstrated, 0 p,p'-TDE (also known as DDD) is 1,1'-(2,2-dichloroethyhdene)b•s[4..:hlorobenzene]; p,p'-DDE I, I' -(2,2-<llchloroethenylidene)bis[4-<:hlorobenzene].
IS
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under laboratory conditions similar to those found in the upper atmosphere, that DDT breaks down to carbon dioxide and hydrochloric acid. 4.1.2.2 Water Concentrations of DDT in rainwater have usually been of the same order of magnitude (1.8 x I0- 5 to 6.6 x Io-s mg/litre) in both agricultural and remote non-agricultural areas. Concentrations of DDT in surface-water depend on levels in rainwater and in the soil and the nature of the soil. Concentrations in the USA are said to have reached a peak in 1966 and then dropped sharply in 1967 and 1968. The highest surface-water value for a DDT-related compound in those years was 0.84J.tg/litre. By 1971, the average concentration in the Federal Republic of Germany was O.OlJ.tg/litre, never reaching as high as lJ.tg/litre. In recent years, concentrations in potable water have been much less than lJ.tg/litre and average concentrations have been similar to those for rainwater. 4.1.2.3 Food The daily intake of DDT from food has been measured in several countries. In the USA during 1953 and 1954, average daily intakes of DDT and of total DDT (DDT+ ODE+ TOE) were 0.184 and 0.286 mg per person, respectively, most of which originated from foods of animal origin. Ten years later, following restrictions with regard to the application of DDT to livestock, intake had been reduced by more than 75%. The US Market Basket Survey showed a gradual decrease in daily intake of DDT to 0.015 mg per person in 1970. Intake in Canada and the United Kingdom was slightly less for comparable periods. In many countries of Europe and in other countries with similar diets the intake of DDT has been judged to be about the same. Worldwide measurements of storage of DDT and its metabolites in human body fat indicate that the extremes of total exposure have varied by a factor of about 10, but that total exposure for most populations has varied by a factor of no more than 3. Food is the major source of DDT for the general population and over 90% of the DDT stored in man is derived from food. 4.1.3 Metabolism DDT is absorbed via both inhalation and ingestion. Absorption of small doses is virtually complete and is facilitated by the presence of fat in food. DDT is poorly absorbed through the skin. Levels of storage in adipose tissue increase rapidly at first and then more gradually until a steady state is reached. In man, the time necessary to reach storage equilibrium is at least one year. There is a gradual reduction in the
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amount of DDT stored in the tissue if exposure to the compound is discontinued. Like most species, man converts some DDT to DOE. A small amount of TOE (DOD), an intermediate in the formation of the main excretory product 2,2-bis(4-ch1orophenyl)acetic acid (DDA), may also be found in tissues. Concentrations of total DDT in the blood of the general population of different countries lie within the range 0.01-0.07 mg/litre. In the blood and other tissues of the fetus or newborn infants, DDT concentrations are lower than in the corresponding tissues of the mother. Levels of DDT in human milk have usually been reported to be in the range 0.01-0.10 mgjlitre, with the concentration of DDT plus its metabolites, especially DOE, being about twice as high. Among the general public, the average concentration of DDA in the urine is 0.014 mgjlitre. Animal studies indicate that the concentration in serum most accurately reflects the concentration in the brain, the critical tissue.
4.1.4 Health effects The acute oral toxicity of DDT is affected by the solvent vehicle; administered in oil to the rat, a typical median lethal dosage is 250 mg/kg of body weight. DDT is poorly absorbed through the skin. The main effect of DDT is on the nervous system. All parts, both central and peripheral are affected to some degree. It appears that its toxic action is associated with its effects on the membranes of the nervous system. The liver is the only other organ significantly affected by DDT. Potentially fatal doses cause focal necrosis of liver cells in several species. DDT produced liver tumours in several strains of mice, nonmetastasizing liver tumours in one study in rats, and no carcinogenic effect in hamsters (2-4). DDT induces microsomal enzymes in all species tested, but only in some rodents does the endoplasmic reticulum increase so much that the entire liver enlarges and granules are displaced to the margin of the cell. These changes are accompanied by a moderate increase in fat droplets, some of which form so-called lipospheres. In long-term feeding tests in mice and rats, the changes in the liver progress from hypertrophy margination and lipospheres to the formation of nodules of affected cells. The same series of changes can be produced in rodents by other inducers of microsomal enzymes. The entire continuum of changes from the prompt response in isolated cells to the eventual formation of tumours is peculiar to some rodents, and other species do not respond morphologically in the same way. Levels as high as 200 mgjkg of feed do not produce adverse effects on the reproduction of rats. In dogs receiving a dosage of 10 mg/kg of body weight per day, reproduction was also normal. No teratogenic effects have been observed in the several animal species
I94
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studied. DDT has not been found to be mutagenic in bacterial test systems. The evidence from mammalian test systems, in vitro and in vivo, is inconclusive. Investigation of workers who have been exposed for as long as 25 years to higher levels of DDT than the general population has not revealed any evidence that DDT causes cancer in man. Although a large number of epidemiological studies were carried out over the years I960 to 1981, involving workers exposed to DDT, the results were inadequate to permit an evaluation of its carcinogenicity to humans. The toxicity data have been evaluated and in I969 a conditional acceptable daily intake for man was estimated as 0-0.005 mgjkg of body weight (5).
REFERENCES I. DDT and its derivatives. Geneva, World Health Organization, 1979 (Environmental Health Criteria 9). 2. Some organochlorine pesticides. Lyon, International Agency for Research on Cancer, 1974 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 5). 3. ROSSI, L. ET AL. Long term administration of DDT or phenobarbitai-Na in Wistar rats. International journal of cancer, 19: 179 (1977). 4. National Cancer Institute Bioassay, Bethesda, MD, Department of Health, Education, and Welfare, 1978 (Technical Report No. 131). 5. Evaluations of some pesticide restdues in food. Geneva, World Health Organization, 1970 (FAO/PL: l969/Mfl7jl; WHO/Food Add./70. 38).
4.2 Aldrin and dieldrin 4.2.1 General aspects The chemical names of these two related pesticides are as follows: Aldrin (HHDN): I ,2,3,4, IO, IO-hexachloro-I ,4,4a,5,8,8a-hexahydroendo-I ,4-exo-5,8-dimethanonaphthalene. Dieldrin (HEOD): I ,2,3,4, I 0, I 0-hexachloro-6, 7-epoxy-1 ,4,4a,5,6, 7,8,8aoctahydro-endo-I, 4-ex o- 5, 8-dimethanonaphthalene. Aldrin and dieldrin are persistent insecticides and accumulate in the food chain. Dieldrin is formed from aldrin by metabolic oxidation in animals and by chemical oxidation in soils. Both insecticides were used for soil treatment against various soil insects, for seed treatment, and for foliar application on various agricultural crops. Over the years their use as foliar treatments has almost disappeared. Their use for the other purposes has also been gradually limited or prohibited; currently the largest use is for termite control. In some countries, it is necessary to treat the soil around fruit trees to form a barrier against termites attacking roots and trunks of trees.
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The risk attached to the use of aldrin and dieldrin has been evaluated and reviewed by many groups since 1966 (J-5). 4.2.2 Routes of exposure As aldrin is readily converted to dieldrin in plants and animals it is rarely found as such in soil, food, and water. 4.2.2.1 Air Dieldrin has been detected in ambient air at very low concentrations. In the USA a maximum level of 29.7 ng/m 3 has been reported; in London and its suburbs, levels up to 1.9 x 10- 4 ngjm 3 of air have been found associated with airborne dust. In the USA, a threshold limit value of 0.25 mgjm 3 has been established for an 8-hour, time-weighted average occupational exposure. 4.2.2.2 Water Concentrations found in rainwater range from 5 to 42 ng/litre and in surface-water (mostly rivers and lakes in the USA) from 0 to 0.1 Jlg/litre. Concentrations of 1-2 ngjlitre were found in drinking-water. 4.2.2.3 Food Dieldrin is stored in adipose tissues, liver, brain, and muscle of mammals, fish, and birds, and in other parts of the food chain. Fish can build up concentrations of dieldrin amounting to several mg/kg of body weight from concentrations of a few ngjlitre in water. In the USA, the total dietary intake was found to range between 0.05 and 0.08 Jlg/kg of body weight per day during the period 1965-1970. In the United Kingdom, a total diet study from 1966 to 1967 showed the intake of dieldrin to be 0.09 Jlgfkg of body weight per day. This intake dropped to 0.03 Jlg/kg of body weight per day in 1970-71. An intake of 0.07 Jlg/kg of body weight per day has been estimated in Japan. Levels of dieldrin in the body fat of the general population in the years 1961-68 in the USA and the United Kingdom and from 1964 to 1966 in six other countries ranged from 0.03 to 0.45 mgjkg of tissue. On the basis of a linear relationship between storage in body fat and exposure, the average intake of dieldrin can be estimated to have ranged from 0.01 to 0.35 Jlgfkg of body weight per day for a 70-kg man (5).
4.2.3 Metabolism In all species examined (mouse, rat, rabbit, rhesus monkey, chimpanzee) 12-hydroxydieldrin and 4,5-aldrin-trans-dihydrodiol were the major metabolites. Regarding the ratios of these two metabolites, the rats and primates seemed to metabolize dieldrin mainly by direct oxidatton
196
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resulting in 12-hydroxydieldrin, whereas in mice and rabbits the main metabolic process seemed to be the opening of the epoxide ring to the diol. This metabolic pathway, together with a high rate of metabolism compared with other animal species examined, could result in relatively high concentrations of 4,5-aldrin-trans-dihydrodiol in the mouse liver. In humans, neither unchanged dieldrin nor the urinary metabolites found in rats were identified in workers occupationally exposed to dieldrin or in monkeys fed the insecticide. Dieldrin is a powerful inducer of microsomal enzymes. Mice react rather anomalously and, therefore, may not be an appropriate model for humans. The results of prolonged enzyme induction are proliferation of the endoplasmic reticulum in the cells and hypertrophy and hyperplasia of the liver. No measurable microsomal enzyme induction occurs in man after long exposure to approximately 0.01 mgjkg of body weight per day. The calculated half-life of dieldrin in workers who were removed from exposure was seven months. The mean dieldrin concentration in blood during the last half-year of exposure was 0.1 mgjlitre, which corresponded to an average daily oral intake of about 0.17 mg/kg of body weight (2, 4).
4.2.4 Health effects The primary site of action of dieldrin is the central nervous system. CNS stimulation is the cause of death in acute poisoning (J). Dieldrin was not proved to be mutagenic in any of the in vitro or in vivo mutagenicity studies. Teratogenicity studies in different animal species showed dieldrin not to be a teratogen. The administration of a single high dose of 15 mg of dieldrin per kg of body weight to mice or 30 mgjkg of body weight to hamsters resulted in minor malformations, which were attributed to maternal toxicity. Results of various carcinogenicity tests in mice and in other mammalian species indicate that there is a species-specific effect of aldrin and dieldrin on the mouse liver resulting in an increased frequency of liver tumours that is peculiar to this animal species (4). A dose-response effect has been demonstrated in both sexes with an increased tumour incidence in females at the lowest dose tested (about 0.015 mgjkg of body weight per day). The available data in rats have not provided evidence of carcinogenicity at levels of up to 2.5 mgjkg of body weight per day. A carcinogenicity study on dieldrin in hamsters was negative (6).
An epidemiological study carried out on occupationally exposed workers did not allow any conclusions to be made concerning the existence of an excess risk of developing cancer (5). Epidemiological studies were inadequate for evaluation (7, 8).
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Toxicological data available in 1977 supported the view that dieldrin and aldrin are not carcinogens; therefore, the previously estimated acceptable daily intake of 0.0001 mgfkg of body weight for aldrin and dieldrin residues, separately or for combined exposure, was reaffirmed (4).
REFERENCES I. Evaluations of some pesticide residues in food. Geneva, World Health Organization, 1967
(FAOjPL: CP/15; WHO/Food Add./67.32). 2. Evaluations of some pesticide residues in food. Geneva, World Health Organization, 1971 (FAO/AGP/1970/M/12/1; WHO/Food Add./71.42). 3. WHO Technical Report Series, No. 612, 1976 (Pesticide residues in food: report of the 1976 Joint FAOjWHO Meeting). 4. Evaluations of some pesticide residues in food. Report 1977. Rome, FAO, 1978 (Plant Production and Protection Paper, No. 10). 5. Some organochlorine pesticides. Lyon, International Agency for Research on Cancer 1974. (IARC Monographs on the evaluation of carcinogenic risk of chemicals to humans vol. 5) 6. CABRAL, J. R. ET AL. A carcinogenicity study of pesticide dieldrin in hamsters. Cancer letter, 6: 241 (1979). 7. VAN RAALTE, H. G. Human experience with dieldrin in perspective. Ecotoxicology and environmental safety, 1: 201 (1977). 8. DEICHMANN, W. B. & MACDoNALD, W. E. Organochlorine pesticides and liver cancer deaths in the United States, 1932-1972. Ecotoxicology and environmental safety, 1: 89 (1977).
4.3 Chlordane 4.3.1 General aspects Pure chlordane is a pale yellow liquid having a molecular formula C 10H 6 Cl 8 and a relative molecular mass of 409.8 {1, 2). The chemical name of chlordane is 1,2,4,5,6,7,8,8-octachloro-2,3,3a,4,7,7a-hexahydro-4,7methanoindene (1). Pure chlordane is composed of a mixture of stereoisomers, with the cis- and trans- forms predominating and referred to as alpha- and gamma-isomers, respectively (3). It is soluble in water at concentrations that have been shown to be toxic to aquatic organisms. Brooks (3) reported the solubility of chlordane in water to be approximately 9 JLgflitre at 25 °C. Chlordane is a broad-spectrum insecticide of the group of polycyclic chlorinated hydrocarbons called cyclodiene insecticides. Chlordane has been used extensively over the past 30 years for termite control, as an insecticide for homes and gardens, and as a control for soil insects during the production of crops, such as corn. Both the uses and the production volume of chlordane have decreased extensively in recent years.
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4.3.2 Routes of exposure 4.3.2.1 Water A study of the persistence of technical chlordane in river-water showed 85 '/0 remaining after 8 weeks (4). Of the major components of technical chlordane, cis- and trans-chlordane were completely stable over the 8-week period. All but two of the remaining components were at least partially changed. Schaefer et al. (5) examined over 500 grab samples from water supplies of the Mississippi and Missouri Rivers. Chlordane was detected in over 20 'l'o of the finished water samples, with the maximum concentration being 8 Jig/litre. Chlordane has also been detected in rainwater (6, 7). Although reports are occasionally received of individual household wells becoming contaminated after a house is treated with chlordane for termite control, only one report has been published of the contamination of a municipal water system (8). On 24 March 1976, a section of the public water system supplying I 05 persons in Chattanooga, Tennessee, USA, became contaminated. Back siphonage apparently occurred while a chlordane concentrate was being diluted with a hose during a period of negative pressure. Of the 71 residents affected, 13 had mild symptoms of acute chlordane toxicity. None of the residents has had prolonged sequelae from the exposure. 4.3.2.2 Food Th~: Food and Drug Administration (FDA) of the USA has been systematically monitoring chlordane in the food supply of the USA since 1965. Chlordane was found infrequently during the first II years of the survey (9). The only quantifiable sample collected was found to contain 0.059 mg/kg, measured in a sample of grain in 1972 (10). In the most recent published results (for 1975) chlordane was not detected. The National Academy of Sciences of the USA (11) in reviewing the results of Moorea reported that of 200 samples of milk collected in Illinois during the period 1971-73, 87 '/0 were positive for chlordane. The average concentration was 50 Jig/litre. Cyclodienes, such as chlordane, are apparently ingested with forage and tend to concentrate in lipids. Oxychlordane, a major mammalian metabolite of chlordane and heptachlor, was found in 46 'l'o of 57 human milk samples collected during 1973-74 in Arkansas and Mississippi. The mean concentration was 5 Jig/litre and the maximum was 20 Jig/litre (12).
4.3.2.3 Air In a survey of the extent of atmospheric contamination by pesticides in the USA, air was sampled at nine localities representative of both a MOORE,
S III Proc 27th Ilhn01s Custom Spray Operators Trammg School Urbana, IL. 1975.
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urban and agricultural areas. At least one chlorinated pesticide was found at all locations, but chlordane was not found in any samples (13). In a larger survey summarized by Nisbet (9), 2479 samples were collected at 45 cities in 16 states of the USA. Chlordane was detected in only two samples, with concentrations of 84 and 204 ng/m 3 . 4.3.2.4 Other routes of exposure Chlordane can be absorbed through the skin and produce toxic effects (14). Dermal exposure would be expected to occur only with occupational manufacture or use of the pesticide. Absorption can range from negligible to that producing acute effects, depending on the degree of exposure. For the general population, dermal exposure would be negligible. In persons using chlordane, the pesticide may persist on the skin for long periods. In one study, hexane rinsings of the hands of a former pest-control operator contained chlordane two years after his last known exposure (15). 4.3.2.5 Relative significance of routes of exposure For the non-occupationally exposed individual, fish and shellfish probably represent the most significant exposure to chlordane. The relative significance of seafood is due to the high bioconcentration factor whereby trace amounts of chlordane in the water may be concentrated to much higher levels in the seafood. Drinking-water and ambient air are relatively insignificant sources of chlordane. Occupational exposure, especially for persons who apply chlordane for pest-control purposes, undoubtedly represents the single greatest exposure to this chemical.
4.3.3 Metabolism A single oral dose of chlordane administered to rats resulted in approximately 6% absorption (16). Small daily doses resulted in absorption of I 0-15 '/'~. The levels of residues in the fat of the rats, after being fed diets containing 1, 5, and 25 mg/kg for 56 days, were approximately three times the concentration in the diet. Concentrations in the liver, kidney, brain, and muscle were 12, 10, 4 and 2%, respectively, of the concentration in the feed. Once the chlordane was removed from the diet, all the residues declined steadily for 4 weeks, by which time the concentrations were reduced by about 60 %. During the following 4 weeks, the residues declined only slightly. In rats, most chlordane is excreted in the faeces. Only about 6% of the total intake is voided in the urine. Rabbits, however, show a different pattern. Urinary elimination of chlordane in rabbits is greater than excretion in the faeces. Human half-life data were obtained when chlordane was accidentally ingested by a young boy (17). A whole-body half-life of 21 days was calculated, which is long compared with drugs used in therapy, but quite
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short when compared with other chlorinated insecticides. This half-life value compares with a half-life of about 23 days obtained by Barnett & Dorough (16) in studies with rats fed chlordane for 56 days. After the levels reached 60 %, further reduction was slight. The serum half-life of chlordane in a young girl was found by Aldrich & Holmes (18) to be 88 days. Polen et al. (19) and Street & Blau (20) found oxychlordane to be a mammalian metabolite of chlordane, and to persist in adipose tissue. Street & Blau (20) observed that the toxicity of oxychlordane was greater than that of the parent compound. Barnett & Dorough (16) tentatively identified several hydroxylated metabolites of chlordane in rat excreta, in addition to oxychlordane, and concluded that the metabolism of chlordane takes place via a series of oxidative enzyme reactions.
4.3.4 Health effects 4.3.4.1 Toxicity Human toxicity data for chlordane are usually obtained after accidental exposure to the compound. Curley & Garrettson (17) reported that shortly after a 20-month-old boy accidentally drank an unknown amount of chlordane he vomited and began a series of convulsions lasting 3-5 minutes each. After being given phenobarbital in a dose of 14 mg/kg of body weight, the seizures stopped. Body temperature rose to 38.9 oc and then gradually decreased to normal. At no time was there evidence of pulmonary disease. Neurological examination at the time seizures were occurring revealed brisk deep tendon reflexes in all extremities. Cranial nerve function was intact and nystagmus was absent. An EEG taken 48 hours after exposure was normal. Three months after exposure, all tests appeared normal. Similar cases were reported by Dadey & Krammer (21) and Aldrich & Holmes (18). A review of the literature by the National Institute for Occupational Safety and Health of the USA (22) indicated chlordane LD 50 values ranging from 100 mg/kg of body weight for rabbits with oral administration to 700 mgfkg of body weight for rats with dermal administration. Chlordane fed to rats in a concentration of 2.5 mg/kg of feed caused slight liver damage (11). 4.3.4.2 Teratogenicity Chlordane was found not to be teratogenic in rats when fed at concentrations of 150-300 mgfkg of feed during pregnancy (23). 4.3.4.3 Mutagenicity Arnold et al. (24) administered chlordane to Charles River CO-l male mice in a single dose of 50 or 100 mg/kg of body weight. The males were subsequently mated with untreated female mice. No dominant lethal changes were produced. Studies by Ahmed et al. (25) with SV-40 transformed human fibroblast cell line V A-4 showed chlordane-induced
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unscheduled DNA synthesis, indicating that chlordane is a potential mutagen. Metabolic activation eliminated the induction of unscheduled DNA synthesis. Simmon et al. (26) found that neither pure cis-chlordane nor trans-chlordane were mutagenic in the Ames Salmonella microsome assay. Technical grade chlordane, however, was found to be mutagenic in Salmonella typhimurium strains TA 1535, TA 98, and TA 100. An S-9 liver activation mix did not enhance the mutagenic activity. 4.3.4.4 Carcinogenicity A bioassay of chlordane for possible carcinogenicity was conducted in the USA, by the Gulf South Research Institute under contract to the National Cancer Institute (27). Groups of 50 mice of each sex and 35 days of age were administered the test material at two concentrations for 80 weeks and then observed for 10 weeks. Hepatocellular carcinoma showed a highly significant dose-related trend for the mice. Similar studies were conducted by the Gulf South Research Institute with chlordane using Osborne-Mendel strain rats. In contrast to the findings in mice, hepatocellular carcinomas failed to appear at a significant rate of incidence in rats administered chlordane. Several epidemiological studies have been published on occupationally exposed persons. At Marshall, Illinois, where chlordane had been manufactured since 1946, no discernible hazard was uncovered among the workers engaged in the production of chlordane, and there was no evidence that chlordane was carcinogenic (28). Studies of personnel at three large pest control companies in the USA, including termite control operators, uncovered no evidence of increased mortality from cancer. There were no deaths attributable to cancer of the liver (28). Epidemiological studies on pesticide applicators exposed to chlordane were inadequate for evaluation (29). In one report, 5 out of 14 children with neuroblastoma had had prenatal and/or postnatal exposure to chlordane, and in one epidemiological study, three persons with acute leukaemia were found to have been exposed to chlordane (which contained 3-7% heptachlor) (30).
As carcinogenicity has been conclusively demonstrated in only one animal, the mouse, a limit for chlordane should be based on toxicity. An acceptable daily intake for man has been estimated as 0.001 mgjkg of body weight (31). For a 70-kg man, this gives an acceptable daily intake of 0.07 mg. Allocating 1 % of this figure to water, and assuming an average daily water intake of 2 litres, the value for chlordane in drinking water can be calculated as 0.35 J.Lg/litre, rounded off to 0.3 J.Lg/litre.
REFERENCES I. WINDHOLZ, M. The Merck index. Rahway, NJ, Merck and Co., 1976. 2. WHETSTONE, R. R. Kirk-Othmer encyclopedia of chemzca/ technology. New York, John Wiley and Sons, 1972.
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3. BROOKES, G. T. Chlorinated insecticides. Cleveland, OH, Chemical Rubber Company Press, 197 4. 4. EICHELBERGER, J. W. & LICHTENBERG, J. J. Persistence of pesticides in river water. Environmental science and technology, 5: 541 (1971). 5. ScHAEFER, M. L. ET AL. Pesticides in drinking water. Environmental science and technology, 3: 1261 (1969). 6. BEVENUE, A. ET AL. Organochlorine pesticides in rainwater, Oahu, Hawaii, 1971-72. Bulletin of environmental contammation and toxicology, 8: 238 (1972). 7. ENVIRONMENTAL PROTECTION AGENCY. Consolidated heptachlor/chlordane hearing. Federal register, 41: 7552 (1976). 8. HARRINGTON, J. M. ET AL. Chlordane contaminatiOn of a municipal water system. Environmental research, 15: 155 (1978). 9. NISBET, I. C. T. Human exposure to chlordane, heptachlor, and their metabolites. Washington, DC, US Environmental Protection Agency, 1976. 10. MANSKE, D. D. & JOHNSON, R. D. Pesticide residues in total diet samples (VIII). Pesticides monitoring journal, 9: 94 (1975). II. NATIONAL RESEARCH COUNCIL Drinkmg water and health. Washington, DC, National Academy of Sciences, 1977. 12. STRASSMAN, S. C. & KuTZ, F. W. Insecticide residues in human milk from Arkansas and Mississippi, 1973-74. Pesticides monitoring journal, 10: 130 (1977). 13. STANLEY, C. W. ET AL. Measurement at atmospheric levels of pesticides, Environmental science and technology, 5: 430 (1971 ). 14. GossELIN, R. E. ET AL. Clinical toxicology of commercial products, 4th ed. Baltimore, MD, Williams and Wilkins Co., 1976. 15. KAZEN, C. ET AL. Persistence of pesticides on the hands of some occupationally exposed people. Archwes of env~ronmental health, 29: 315 (1974). 16. BARNETT, J. R. & DoROUGH, H. W. Metabolism of chlordane in rats. Journal of agricultural and food chemistry, 22: 612 (1974). 17. CuRLEY, A. & GARRETTSON, L. K. Acute chlordane poisoning. Archives of environmental health, 18: 211 (1969). 18. ALDRICH, F. D. & HOLMES, J. H. Acute chlordane intoxication in a child. Archives of environmental health, 19: 129 (1969). 19. POLEN, P. N. ET AL. Characterization of oxychlordane, animal metabolite of chlordane. Bulletin of environmental contamination and toxicology, 5: 521 (1971). 20. STREET, J. E. & BLAU, S. E. Oxychlordane: accumulation in rat adipose tissue on feeding chlordane isomers or technical chlordane. Journal of agricultural and food chemistry, 20: 395 (1972). 21. DADEY, J. L. & KRAMMER, A. G. Chlordane intoxication. Journal of the American Medical Association, 153: 723 (1953). 22. NATIONAL INSTITUTE FOR OCCUPATIONAL SAFETY AND HEALTH. Suspected carcinogens-subfile of the NIOSH registry of tox1c effects of chemical substances. 2nd ed. Washington, DC, NIOSH, 1976 (NISH-77-149). 23. INGLE, L. Chrome oral toxicity of chlordane to rats. Archives of industrial hygiene and occupational medicine, 6: 357 (1952). 24. ARNOLD, D. W. ET AL. Dominant lethal studies with technical chlordane, HCS-3260, and heptachlor: heptachlor epox1de. Journal of toxicology and environmental health, 2: 547 (1977). 25. AHMED, F. E. ET AL. Pesticide induced DNA damage and its repair in cultured human cells. Mutatation research, 42: 161 (1977). 26. SIMMON, V. F. ET AL. Mutagenic activity of chemicals identified in drinking water. Developments in toxicology and environmental science, 2: 249 (1977). 27. NATIONAL CANCER INSTITUTE, Bioassay of chlordane for possible carcinogenicity. Bethesda, MD, Department of Health, Education and Welfare, 1977 (NCI-CG-TR-8). 28. ENVIRONMENTAL PROTECTION AGENCY. Consolidated heptachlor /chlordane hearing. Federal register, 41: 7552-7572, 7584-7585 (1976). 29. WANG, H. H. & MAcMAHON, B. Mortality of pesticide applicators. Journal of occupational medicine, 21: 741 (1979).
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30. Some halogenated hydrocarbons. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of the carcinogenic risk of chem1cals to humans, vol. 20). 31. Evaluations of some pesticide residues in food. (FAO/PL 1967/M/11/1; WHO Food Add./68.30.)
4.4 Hexachlorobenzene 4.4.1 General aspects Hexachlorobenzene (HCB) is a white solid with a melting point of 230 oc. It is of low solubility in water (6 Jlg/kg) (1) but is soluble in organic solvents. The pure material is produced commercially, principally for use as a fungicide. The largest input to the environment, however, is its generation as a by-product in the manufacture of chlorine and other chlorinated chemicals, particularly solvents.
4.4.2 Routes of exposure 4.4.2.1 Air People are exposed to HCB in air, water, and food. It is disseminated in the air as dust particles and as a result of volatilization from sites having a high HCB-concentration. Airborne HCB-laden dust particles appear to have been a major cause of increased blood concentrations of HCB in the general public living near an industrial site in Louisiana, USA (2). 4.4.2.2 Water HCB has been found in 4 river-water samples, 8 finished drinkingwater samples, 1 sample from a sewage-treatment plant, and in effluent water from 7 chemical plants in various locations in Europe and the USA (3). It has also been detected in urban rainwater runoff in the USA, at levels of 0-339 ng/litre (4); in the Rhine River (5); in 108 samples of surface-water in Italy, at average levels of 2.5 ng/litre (6); and in most river-water residues in an industrialized region of the USA, generally. at levels of less than 2 jlgjlitre, but as high as 90 jlgjlitre in one sample (7). 4.4.2.3 Food HCB occurs in a wide variety of foods, in particular, terrestrial animal products, including dairy products and eggs (8). The average intake of HCB from food in the USA was estimated as 0.4 Jlgjday in 1973 and 0.07 Jlg/day in 1974 (9). The dietary intake in Japan has been estimated to be 0.5 Jlg/day (10) and in Australia 35 Jlg/day (11). Breast-fed infants In Australia and Norway may consume up to 40 Jlg/day (11, 12). Average HCB levels in human adipose tissue have ranged from 0.02 to
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8.2 mgjkg (13) and in human blood samples from 0.004 to 0.06 mgjkg (13). The levels of HCB in the body fat of swine and sheep were, respectively, sixfold and eightfold greater than the dietary level (14). If these comparisons are valid when applied to man, it would appear that some adults have been exposed to levels of HCB of several mgjkg of body weight per day. A similar conclusion is reached by extrapolating the values for human blood. The HCB levels in the blood of rats are about one-tenth the dietary level (15). Current evidence would indicate that food intake may be the primary source of the body burden of HCB for the general population, although inhalation and dermal exposure may be more important in selected groups, e.g., industrial workers.
4.4.3 Metabolism Hexachlorobenzene (16) administered orally to rats was absorbed slowly from the gut, mainly via the lymphatic system, and was stored extensively in the fat after 48 hours (17). The quantitative recovery of intraperitoneally and orally administered [l 4 C]-HCB in rats was dosedependent, but more 14 C was recovered from the faeces than from the urine. The major urinary metabolites were pentachlorophenol, tetrachlorohydroquinone, and pentachlorothiophenol. The other urinary metabolites were tetrachlorobenzene, pentachlorobenzene, 2,4,5- and 2,4,6-trichlorophenols, and 2,3,4,6- and 2,3,5,6-tetrachlorophenols; 2,3,4trichlorophenol and other tetrachlorophenols were present in traces. These metabolites were excreted as conjugates or in free form in the urine. Unchanged HCB was found in the faeces and in fat (18-21). When [l 4 C]-HCB was given orally in a dosage of 110 J.Lgfday to Macaca mulatta monkeys for 11-15 months, 50% of the radioactivity found in the urine was in pentachlorophenol and 25% in pentachlorobenzene, the remainder being in unidentified metabolites and unchanged HCB. In the faeces, 99% of the radioactivity was in unchanged HCB. During the last 10 days of the experiment, males excreted 7.2% of the administered dose in the urine and 52% in the faeces; females excreted 4.6% and 42.2 %, respectively (22).
4.4.4 Health effects 4.4.4.1 Toxicity Although HCB has a low acute toxicity for most species (> 1000 mg/kg of body weight), it has a wide range of biological effects at prolonged moderate exposure. Subacute toxic effects were examined in rats fed with HCB for 15 weeks. Histopathological changes were confined to the liver and spleen. In the liver, there was an increase in the severity of centrilobular liver lesions with as little as 2 mg of HCB per kg of food per day. It would
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appear that 0.5 mg of HCB per kg of body weight per day is the noeffect level in the rat (15). In a similar study, it was found that 0.05 mg of HCB per kg of feed per day is the no-effect level for swine (23). In rats given 50 mg of HCB per kg of body weight every other day for 53 weeks, an equilibrium between intake and elimination was achieved after nine weeks. In general, the changes observed in the long-term studies resembled those described for short-term studies. When the administration of HCB was discontinued, elimination of the xenobiotic continued slowly for many months (24). An epidemic of HCB-induced porphyria cutanea tarda occurred in Turkey during the period 1955 to 1959 (25). More than 600 patients were observed during a 5-year period, and it was estimated that a total of 3000 people were affected. The outbreak was traced to the consumption of wheat as food after it had been prepared for planting by treatment with HCB. The syndrome involves blistering and epidermolysis of the exposed parts of the body, particularly the face and hands. It was estimated that the subjects ingested 50-200 mg of HCB per day for a relatively long period before the skin manifestations became apparent. The symptoms were seen mostly during the summer months, having been exacerbated by intense sunlight. The disease subsided and symptoms disappeared 20-30 days after discontinuing the intake of HCBcontaminated bread. Relapses were often seen, either because the subjects were eating HCB-containing wheat again, or because of redistribution of HCB stored in body fat. A disorder called pembe yara was described in infants of Turkish mothers who either had HCB-induced porphyria or had eaten HCBcontaminated bread (26). The maternal milk contained HCB. At least 95% of these infants died within a year. On the basis of toxicological considerations, FAO/WHO (27) suggested a human ADI of 0.6 J,l.g per kg of body weight, but this has now been withdrawn. 4.4.4.2 Teratogenicity Placental transfer of HCB has been reported in mice and rats (28, 29). A minimal teratogenic effect of HCB observed in Wistar rats could not be reproduced in the same laboratory with doses up to 120 mg/kg of body weight administered during organogenesis (30). In other studies with hexachlorobenzene and with pentachloronitrobenzene (PCNB) contaminated by hexachlorobenzene at a dose level of 100 mg/kg of body weight, cleft palate and some kidney malformations were found in mice (31).
In a 4-generation test, groups of 10 male and 20 female SpragueDawley rats were treated with HCB in dietary concentrations of 0, 10, 20, 40, 80, 160, 320 or 640 mg/kg from weaning. Suckling pups in the Fl generation were particularly sensitive, and many died prior to weaning when the mothers were fed dietary concentrations of 320 or 640 mg/kg of feed. No gross abnormalities were found (32).
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4.4.4.3 Carcinogenicity
Two studies have been conducted that indicate that HCB is a carcinogen. The carcinogenic activity of HCB in hamsters fed 4, 8, or 16 mgfkg of body weight per day for life was assessed (33). HCB appears to have multipotential carcinogenic activity; the incidence of hepatomas, haemangioendotheliomas, and thyroid adenomas was significantly increased. Whereas 10% of the unexposed hamsters developed tumours, 92% of the hamsters fed 16 mg of HCB per kg of body weight per day developed tumours. The incidence of tumour-bearing animals was doserelated: 56% for hamsters fed 4 mg of HCB per kg of body weight per day and 75% for 8 mg per kg of body weight per day. Thyroid tumours, hepatomas, or liver haemangioendotheliomas were not detected in the unexposed group. An intake of 4-16 mg of HCB per kg of body weight per day in hamsters is near the exposure range estimated for Turkish people who accidentally consumed HCB-contaminated grain (33). The carcinogenic activity of HCB in mice fed 6.5, 13, or 26 mgfkg of body weight per day for life was assessed. The incidence of hepatomas was increased significantly in mice fed 13 or 26 mg of HCB per kg of body weight per day. None of the hepatomas occurred or metastasized in the untreated control groups. The results presented by Cabral et al. (34, 35) confirm their earlier conclusion that HCB is carcinogenic. However, the incidence of lung tumours in strain A mice treated three times a week for a total of 24 injections of 40 mg(kg of body weight each was not significantly greater than the incidence in control mice (36). HCB is also carcinogenic in rats (37). On the basis of the 10- 5 risk level, the guideline value for HCB in drinking-water is recommended as 0.01 J.Lg/litre. REFERENCES I. Lu, P. Y. & METCALF, R. L. Environmental fate and biodegradability of benzene derivatives as studied in a model aquatic ecosystem. Enmronmental health perspectives, 10: 269 (1975). 2 BURNS, J. E. & MILLER, F. M. Hexachlorobenzene contamination: its effects in a Louisiana population. Archives of environmental health., 30: 44 (1975). 3. SHACKELFORD, W. M. & KEITH, L. H. Frequency of organic compounds zdentified in water. Athens, GA, US Environmental Protection Agency, 1976 (EPA-600/4-76-062, p. 69). 4. DAPPEN, G. Pestzczde analysis from urban storm runoff. Spnngfield, VA, Department of the Interior, 1974 (Report No. PB-238 593). 5. GREVE, P. A. Potentially hazardous substances· in surface water. I. Pesticides in the river Rhine. Science of the total envzronment, I: 173 (1972). 6. LEONI, V. & D'ARCA, S. U. Experimental data and critical review of the occurrence of hexachlorobenzene in the Italian environment. Science of the total environment, 5: 253 (1976). 7. LASKA, A. L. ET AL. Distribution of hexachlorobenzene and hexachlorobutadiene in water, soil, and selected aquatic organisms along the lower Mississippi River, Louisiana. Bulletin of environmental contamination and toxicology, 15: 535 (1976). 8. Environmental contamination from hexach/orobenzene. Washington, DC, US
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Environmental Protection Agency, 1976 (EPA 560/6-70-014). 9. US FOOD & DRUG ADMINISTRATION. Compliance program evaluation total diet studtes (7320.08). Washington, DC, US Government Printing Office, 1977. 10. UsHIO, F. & DoGUCHI, M. Dietary intakes of some chlorinated hydrocarbons and heavy metals estimated on the experimentally prepared diets. Bulletin of environmental contamination and toxicology, 17: 707 (1977). II. MILLER, G. J. & Fox, J. A. Chlorinated hydrocarbon pesticide residues in Queensland human milks. Medical journal of Australia, 2: 261 (1973). 12. BAKKEN, A. F. & SEIP, M. Insecticides in human breast milk. Acta paediatrica scandinavica, 65: 535 (1976). 13. Ambient water quality criteria for ch/ormated benzenes. Washington, DC, US Environmental Protection Agency, !980 (Document No. 440/5-80-028). 14. HANSEN, L. G. ET AL. Effects and residues of dietary hexachlorobenzene in growing swine. Journal of toxicology and environmental health, 2: 557 (1977). 15. KUIPER-GooDMAN, T. ET AL. Subacute toxicity of hexachlorobenzene in the rat. Toxicology and applied pharmacology, 40: 529 (1977). 16. Some halogenated hydrocarbons. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 20.). 17. IATROPOULOS, M. J. ET AL. Absorption, transport and organotropism of dichlorobiphenyl (DCB), dieldrin, and hexachlorobenzene (HCB) in rats. Environmental research, 10: 384 (1975). 18. ENGST, R. ET AL. The metabolism of hexachlorobenzene (HCB) in rats. Bulletin of environmental contamination and toxicology, 16: 248 (1976). 19. Koss, G. ET AL. Studies on the toxicology of hexachlorobenzene. II. Identification and determination of metabolites. Archives of toxicology, 35: 107 (1976). 20. MEHENDALE, H. M. ET AL. Metabolism and effects of hexachlorobenzene on hepatic microsomal enzymes in the rat. Journal of agricultural and food chemistry, 23: 261 (1975). 21. RENNER, G. & ScHUSTER, K. P. 2,4,5-Trichlorophenol, a new urinary metabolite of hexachlorobenzene. Toxicology and applied pharmacology, 39: 355 (1977). 22. RoZMAN, K. ET AL. Long-term feeding study of hexachlorobenzene in rhesus monkeys. Chemosphere, 6: 81 (1977). 23. DEN TONKELAAR, E. M. ET Al. Hexachlorobenzene toxicity in pigs. Toxicology and applted pharmacology, 43: 137 (1978). 24. Koss, G. ET Al. Studies on the toxicology of hexachlorobenzene. III Observations in a long-term experiment. Archives of toxicology, 4: 285 (1978). 25. CAM, C. & NIGOGOSYAN, G. Acquired toxic porphyria cutanea tarda due to hexachlorobenzene. Journal of the American Medical Association, 183: 88 (1963). 26. CAM, C. Une nouvelle dermatose epidemique des enfants. Annates de dermato/ogie et de syphiligraphie, 87: 393 (1960). 27. 1973 Evaluations of some pesticide residues in food. Geneva, World Health Organization, 1974 (FAO/AGP/1973/M/9/1; WHO Pesticide Residue Series, 3 p. 291). 28. ANDREWS, J. E. & CoURTNEY, K. D. Inter-and intralitter variation of hexachlorobenzene (HCB) deposition in fetuses. Toxicology and applied pharmacology, 37: 128 (1976). 29. VILLENEUVE, D. C. & HIERLIHY, S. L. Placental transfer of hexachlorobenzene in the rat. Bulletin of environmental contamination toxicology, 13: 489 (1975). 30. KHERA, K. S. Teratogenicity and dominant lethal studies on hexachlorobenzene in rats. Food and cosmetics toxicology, 12: 471 (1974). 31. CoURTNEY, K. D. ET AL. The effects of pentachloronitrobenzene, hexachlorobenzene, and related compounds on fetal development. Toxicology and applied pharmacology, 35: 239 (1976). 32. GRANT, D. L. ET Al. Effect of hexachlorobenzene on reproduction in the rat. Archives of environmental contamination and toxicology, 5: 207 (1977). 33. CABRAL, J. R. P. ET AL. Carcinogenic activity of hexachlorobenzene in hamsters. Nature, 269: 510 (1977).
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34. CABRAL, J. R. P. ET AL. Carcinogenesis study in mice with hexachlorobenzene. Toxicology and applied pharmacology, 45: 323 (1978). 35. CABRAL, J. R. P. ET AL Carcinogenesis of hexach1orobenzene in mice. International journal of cancer, 23: 47 (1979). 36. THEISS, J. C. ET AL. Test for carcmogemcity of organic contammants of United States drinking waters by pulmonary tumor response in strain A mice. Cancer research, 37: 2717 (1977). 37. SMITH, A. G. & CABRAL, J. R. P. Liver cell tumours in rats fed hexach1orobenzene. Cancer letter, 11: 169 (1980).
4.5 Heptachlor and heptachlor epoxide 4.5.1 General aspects Pure heptachlor (C 10 H 5 Cl 7 ; relative molecular mass 373.35) is a white crystalline solid with a camphor-like odour. The chemical name for heptachlor is I ,4,5,6, 7,8,8-heptachloro-3a,4, 7, 7a-tetrahydro-4, 7-methanoindene. It has a vapour pressure of 4 x 10- 2 Pa (3 x 10- 4 mmHg) at 25 °C, a solubility in water of 0.056 mg/litre at 25-29 °C, and is readily soluble in relatively nonpolar solvents (1). Heptachlor is a broad-spectrum insecticide of the group of polycyclic chlorinated hydrocarbons called cyclodiene insecticides. From 1971 to 1975 the most important use of heptachlor was to control agricultural soil insects (1). Since 1975, both the uses and the production volume have declined extensively because of a voluntary restriction by the sole producer and because of the subsequent issuance (2 August 1976) of a registration suspension notice by the EPA for all food crop and home use of heptachlor. However, significant commercial use of heptachlor for termite control and non-field crops continues. Heptachlor persists for prolonged periods in the environment. It is converted to the more toxic metabolite, heptachlor epoxide, in the soil (2-5), in plants (6), and in mammals (7). Heptachlor, in solution or thin films, undergoes photodecomposition to photoheptachlor (8) which is more toxic than the parent compound to insects (9) and aquatic invertebrates (10, 11). Heptachlor and its epoxide undergo bioconcentration in numerous species and will accumulate in the food chain (1).
4.5.2 Routes of exposure 4.5.2.1 Water Various investigators have detected heptachlor and/or heptachlor epoxide in the major river basins of the USA at a mean concentration for both of 0.0063JLgflitre (12). Levels of heptachlor ranged from 0.001 to 0.035 JLg/litre (13).
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4.5.2.2 Food In their Market Basket Study (1974-75) for 20 different Cities, the FDA showed that 3 of 12 food classes contained residues of heptachlor epoxide ranging from 0.0006 to 0.003 mgjkg (14). Heptachlor epoxide residues greater than 0.03 mgjkg have been found in 14-19% of red meat, poultry, and dairy products sampled from 1964 to 1974.a Heptachlor andjor heptachlor epoxide were found in 32 % of 590 fish samples obtained nationally, with whole fish residues from 0.01 to 8.33 mgjkg (15). Human milk can be contaminated with heptachlor epoxide. A nationwide survey indicated that 63.1% of 1936 mothers' milk samples contained heptachlor epoxide residues (16). 4.5.2.3 Air Heptachlor volatilizes from treated surfaces, plants, and soil. Heptachlor and to a lesser extent hepta<:hlor epoxide are widespread in ambient air, with typical mean concentrations of approximately 0.5 ngjm 3 • On the basis of these data, typical human exposure was calculated to be 0.01 JJ.g per person per day.a Thus, it appears that inhalation is not a major route for human exposure to heptachlor.
4.5.3 Metabolism Heptachlor is readily metabolized by mammals to heptachlor epoxide. This metabolite is stored mainly in adipose tissue, but also in liver, kidney, and muscle (17). Both the rat and the dog rapidly metabolize ingested heptachlor to heptachlor epoxide and accumulate heptachlor epoxide primarily in adipose tissue. A positive relationship is found between the amount of heptachlor in the diet and the amount of heptachlor stored in the fatty tissue, female rats accumulating approximately six times as much heptachlor epoxide in their fat as males (7, 18). Although there is no direct evidence showing the conversion of heptachlor to its epoxide in humans, there is little doubt that the epoxide that has been found in human tissues is derived from heptachlor. Various levels of heptachlor epoxide have been found in the blood, fat, and milk of humans. Because of its high lipid content, milk is one of the major excretion routes for organohalogenated compounds, including heptachlor epoxide. An extensive survey carried out in the USA indicated that women who had lactated following several births had lower pesticide levels in milk than primiparae. Heptachlor epoxide, dieldrin, and oxychlordane were the most common a NISBETT, I. C. T. Human exposure to chlordane, hepatoch/or and the.r metabolites (unpublished review).
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pesticides found in human milk. Only 2% of human milk samples showed heptachlor residues, but 63.1% of the samples showed heptachlor epoxide residues, the levels ranging from 15 to 2050 J.Lg/litre on a fat-adjusted basis, with a mean concentration of 91 J.Lg/litre. Of the high-residue group of women, II % were either occupationally exposed or lived in households where a household member was occupationally exposed (16).
4.5.4 Health effects 4.5.4.1 Toxicity The mammalian LD 50 of heptachlor and its metabolites has been reported in a variety of species as ranging from 6 mg/kg to 531 mg/kg of body weight (1). Little information on chronic effects is available. When administered to rats in small daily doses over a prolonged period of time, heptachlor induced alterations in glucose homoeostasis, which were thought to be related to an initial stimulation of the cyclic AMP-adenylate cyclase system in liver and kidney cortex (19-21). A joint F AO /WHO committee (22) has established a maximum acceptable daily intake for heptachlor plus heptachlor epoxide of 0.5 J.Lg per kg of body weight. 4.5.4.2 Teratogenicity In long-term feeding studies with heptachlor, cataracts developed in parent rats and in the offspring shortly after their eyes opened (23). 4.5.4.3 Other reproductive effects In long-term feeding studies in rats, heptachlor caused a marked decrease in litter size and a decreased lifespan in suckling rats (23). 4.5.4.4 Mutagenicity Heptachlor has been reported to be mutagenic in mammalian assays but not in bacterial assays. Heptachlor caused dominant lethal changes in male rats as demonstrated by the number of resorbed fetuses in intact pregnant rats (24). Bone marrow cells of the treated animals showed increases in the incidences of abnormal mitoses, chromatid abnormalities, pulverization, and translocation Both heptachlor and heptachlor epoxide induced unscheduled DNA synthesis in SV -40 transformed human cells (VA-4) in culture with metabolic activation (25). Neither heptachlor nor heptachlor epoxide was mutagenic for Salmonella typhimurium in the Ames test (26).
4. 4.5.4.5 Carcinogenicity
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Heptachlor and/or heptachlor epoxide have induced hepatocellular carcinomas in mice during three chronic feeding studies. Heptachlor epoxide has produced the same response in rats in one study (27).a 4.5.4.6 Guideline value On the basis of the maximum acceptable daily intake of 0.5 p.g per kg of body weight recommended by a joint FAO/WHO committee (22), the guideline value in drinking-water for a 70-kg man consuming 2 litres of water per day can be calculated as 0.1 p.g/litre. As a carcinogenic response has been well documented in only one species of animal, it seems justified to recommend the application of this guideline value.
REFERENCES I. Heptachlor: ambient water quality criteria. Washington, DC, US Environmental
Protection Agency, 1980 (Document No. 440/5-80-052). 2. LICHTENSTEIN, E. P. Insecticidal residues in various crops grown in soils treated with abnormal rates of aldrin and heptachlor. Journal of agricultural and food chemistry, 8: 448 (1960). 3. LICHTENSTEIN, E. P. ET AL. Degradation of aldrin and heptachlor in field soils. Journal of agricultural and food chemistry, 18: 100 (1970). 4. LICHTENSTEIN, E. P. ET AL. Effects of a cover crop versus soil cultivation on the fate of vertical distribution of insecticide residues in soil 7 to II years after soil treatment. Pesticides monitoring journal, S: 218 (1971). 5. NASH, R. G. & HARRIS, W. G. Chlorinated hydrocarbon insecticide residues in crops and soil. Journal of environmental quality, 2: 269 (1973). 6. GANNON, N. & DECKER, G. C. The conversion of aldrin to dieldrin in plants. Journal of economic entomology, S1: 8 (1958). 7. DAVIDOW, B. & RADOMSKI, J. L. Isolation of an epoxide metabolite from fat tissues of dogs fed heptachlor. Journal of pharmacology and experimental therapeutics, 107: 259 (1953). 8. BENSON, W. R. ET AL. Photolysis of solid and dissolved dieldrin. Journal of agricultural and food chemistry, 19: 66 (1971). 9. KHAN, M. H. ET AL. Insect metabolism of photoaldrin and photodieldrin. Science, 164: 318 (1969). 10. GEORGACKAKIS, E. & KHAN, M. A. Q. Toxicity of the photoisomers of cyclodiene insecticides to freshwater animals. Nature, 233: 120 (1971). II. KHAN, M. A. Q. ET AL. Toxicity-metabolism relationship of the photoisomers of certain chlorinated cyclodiene insecticide chemicals. Archives of environmental contamination and toxicology, 1: 159 (1973). 12. Chlordane and heptachlor in relation to man and the environment. Washington, DC, US Environmental Protection Agency, 1976 (EPA 540/476005). 13. BREIDENBACH, A. W. ET AL. Chlorinated hydrocarbon pesticides in major river basins, 1957-65. Public health reports, 82: 139 (1967). 14. JOHNSON, R. D. & MANSKE, D. D. Pesticide and other chemical residues in total diet samples (XI). Pesticides monitoring journal, 11: 116 (1977). 15. HENDERSON, C. ET AL. Organochlorine insecticide residues in fish (National Pesticide Monitoring Program). Pestzcides monitoring journal, 3: 145 (1969). • US Environmental Protection Agency RISk assessment of chlordane and heptachlor. Carcinogen Assessment Group, Washington, DC, 1977 (unpublished report).
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16. SAVAGE, E. P. National study to determine levels of chlorinated hydrocarbon insecticides in human mrlk. Washingdon, DC, US Environmental Protection Agency, 1976 (EPA/540/9-78/005). 17. Evaluations of some pesticide residues m food. Geneva, World Health Organization, 1967. 18. RADONSKI, J. L. & DAVIDOW, B. The metabolic of heptachlor, its estimation, storage and toxicity. Journal of pharmacology and experimental therapeutics, 107: 266 (1953). 19. KAcEw, S. & SINGHAL, R. L. The influence of p,p-DDT, and chlordane, heptachlor and endrin on hepatic and renal carbohydrate metabohsm and cyclic AMP-adenyl cyclase system. Life sciences, 13: 1363 (1973). 20. KAcEw, S. & SINGHAL, R. L Effect of certain halogenated hydrocarbon insecticides on cyclic adenosine 3',5'-monophosphate- 3 H formation by rat kidney cortex. Journal of pharmacology and expenmental therapeutrcs, 188: 265 (1974). 21. SINGHAL, R. L. & KAcEw, S. The role of cyclic AMP in chlorinated hydrocarboninduced toxicity. Federation proceedings, 35: 2618 (1976). 22. 1971 evaluations of some pe~trcide resrdues in food. (AGP: 1971/M/9/1; WHO Pesticide Residues Senes, No. I, p. 314). 23. MESTITZOVA, M. On reproduction studies on the occurrence of cataracts in rats after long-term feeding of the insecticide heptachlor. Expenentia, 23: 42 (1967). 24. CEREY, K. ET AL. Effect of heptachlor on dominant lethality and bone marrow in rats. Mutation research, 21: 26 (1973). 25. AHMED, F. E. ET AL. Pesticide-induced DNA damage and its repair in cultured human cells. Mutation research, 42: 161 (1977). 26. MARSHALL, T. C. ET AL. Screening of pesticides for mutagenic potential using Salmonella typhrmurium mutants. Journal of agricultural and food chemistry, 24: 560 (1976). 27. EPSTEIN, S. S. Carcinogenicity of heptachlor and chlordane. Science of the total environment, 6: 103 (1976).
4.6 Lindane 4.6.1 General aspects Lindane (gamma-hexachlorocyclohexane), also known as y-HCH or }•-BHC, is a white solid with a melting point of 112.5 oC. It is fairly soluble in water (10 mg/litre) (1), but more soluble in organic solvents. Lindane is a broad-spectrum insecticide of the group of cyclic chlorinated hydrocarbons called organochlorine insecticides and is used in a wide range of applications, including treatment of animals, buildings, man (for ectoparasites), clothes, water (for mosquitos), plants, seeds, and soils (2). It is slowly degraded by soil microorganisms (3) and can be isomerized to the alpha and/or delta isomers by microorganisms and plants (2). 4.6.2 Routes of exposure 4.6.2.1 Water Contamination of water has occurred from direct application of technical hexachlorocyclohexane (HCH) or lindane to water for control of mosquitos, from the use of HCH in agriculture and forestry, and, to a
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lesser extent, from occasional contamination of wastewater from manufacturing plants (2). In a survey of finished drinking-water in the USA, the highest reported level of lindane was 0.1 Jlgflitre (4). It is a ubiquitous contaminant of surface waters at levels up to 100 ngjlitre (5, 6), presumably owing to its volatilization and precipitation in rain (7). In the Federal Republic of Germany, lindane was present in all surfacewater samples taken, at levels ranging from 0.005 to 7.1 Jlg/litre. 4.6.2.2 Food The daily intake of lindane has been reported to be 1-5 Jlgfkg of body weight and the daily intake of all other HCH isomers to be 1-3 Jlg/kg of body weight (8). The chief sources of HCH residues in the human diet are milk, eggs, and other dairy products (2). In the USA, the EPA (2) has estimated the weighted average bioconcentration factor for lindane at 780. This estimate is based on the measured steady-state bioconcentration in the bluegill.• 4.6.2.3 Air Traces of HCH have been detected in the air of central and suburban London (2). The uptake of lindane by inhalation is estimated at 0.002 Jlg per kg of body weight per day (9).
4.6.3 Metabolism The rapidity of lindane absorption is enhanced by lipid-mediated carriers. Compared with other organochlorine insecticides, HCH and lindane are fairly soluble in water, which contributes to rapid absorption and excretion (2, 10). Lindane is absorbed after oral or dermal exposure (2). After administration to experimental animals, lindane was detected in the brain at higher concentrations than in other organs (11-13). At least 75% of an intraperitoneal dose of 14 C-labelled lindane was consistently found in the skin, muscle, and fatty tissue (14). Lindane enters the human fetus through the placenta; higher concentrations were found in the skin than in the brain and never exceeded the corresponding values for adult organs (15, 16). Lindane is metabolized to an intermediate hexachlorocyclohexene; further degradation yields 2,3,4,5,6-pentachloro-2-cyclohexene-1-ol, two tetrachlorophenols, and three trichorophenols (17). These are commonly found in the urine as conjugates (18). Both free and conjugated chlorophenols are far less toxic than the parent compounds (19).
a A species of freshwater sunfish common in North America.
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4.6.4 Health effects The oral LD 50 of lindane for rats is 125-230 mgjkg of body weight (20).
In chronic studies with rats given lindane in oil, liver cell hypertrophy (fat degeneration and necrosis) and nephritic changes were noted at high doses (21-23). Rats inhaling lindane (0.78 mgjm 3 of air) for 7 hours, 5 days a week, for 180 days showed liver cell enlargement, but showed no toxic symptoms or other abnormalities (24). The addition of lindane to the diet of rats in a concentration of 10 mgjkg for 1 or 2 years decreased body weight after 5 months of treatment and altered ascorbic acid levels in urine, blood, and tissues (25). Dogs chronically exposed to lindane in the diet had slightly enlarged livers (26). Irritation of the central nervous system with other toxic side-effects (nausea, vomiting, spasms, weak respiration with cyanosis, and blood dyscrasia) were reported after prolonged or improper use of a preparation containing 10 g of lindane per kg for the treatment of scabies on humans (27). Production workers exposed to technical HCH exhibited symptoms including headache, vertigo, irritation of the skin, eyes, and respiratory tract mucosa. In some instances, there were apparent disturbances of carbohydrate and lipid metabolism and dysfunction of the hypothalamo-pituitary-adrenal system (28, 29). A study of persons occupationally exposed to HCH for 11-23 years revealed biochemical manifestations of toxic hepatitis (30). A maximum acceptable daily intake for lindane of 10 J.Lg/kg of body weight has been estimated by a joint FAO/WHO meeting (31). 4.6.4.1 Teratogenicity
Lindane given in the diet during pregnancy at levels of 12 or 25 mg per kg of body weight per day did not produce teratogenic effects in rats (32).
4.6.4.2 Other reproductive effects
Chronic lindane feeding in a study of four generations of rats increased the average duration of pregnancy, decreased the number of births, increased the proportion of stillbirths, and delayed sexual maturation in F 2 and F 3 females. In addition, some of the F 1 and F 2 animals exhibited spastic paraplegia (25). In rats and rabbits, lindane given in the diet during pregnancy increased postimplantation death of embryos (32, 33). 4.6.4.3 Mutagenicity
Evidence for the mutagenicity of lindane is equivocal. Some alterations in mitotic activity and the karyotype of human lymphocytes cultured with lindane at 0.1-10 gjlitre have been reported (34). Lindane was not
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mutagenic in a dominant-lethal assay0 or a host-mediated assay (35). It was found to be mutagenic in microbial assays using Salmonella typhimurium with metabolic activation, the host-mediated assay, and the dominant lethal test in rats. Other reports indicate that it does not have significant mutagenic activity (2). 4.6.4.4 Carcinogenicity An increased incidence of liver tumours was reported in male and female mice of various strains fed lindane (gamma-HCH) (36-40). Epidemiological data were inadequate for evaluation (40, 41). 4.6.4.5 Guideline value Based on the maximum acceptable daily intake of lindane for a 70-kg man recommended by an FAO/WHO meeting (31) and assigning 1% of the ADI to water, a guideline value of 3 ,ug/litre can be calculated for a water consumption of 2 litres per person per day.
REFERENCES I. ULMANN, E., ed. Lindane: monograph of an insecticide. Freiburg, Verlag K. Schillinger, 1972. 2. Hexachlorocyclohexane: ambient water quality criteria. Washington, DC, US Environmental Protection Agency, 1979. 3. MATHUR, S.P. & SAHA, J. G. Microbial degradation of lindane-C-14 in a flooded sandy loam soiL Sot/ science, 120: 301 (1975). 4. US Environmental Protection Agency. Preliminary assessment of suspected carcinogens in drinking-water. Report to Congress. Washington, DC, 1975, p. 11-4. 5. CROLL B. T. Organo-chlorine insecticides in water-Part I. Water treatment examination, 18: 255-274 (1969). 6. GREVE, P. A. Potentially hazardous substances in surface waters. I. Pesticides in the R. Rhine. Science of the total environment. 1: 173-180 (1972). 7. TARRANT, K. R. & TATION, J. O'G. Organo-chlorine pesticides in rainwater in the British Isles. Nature, 219: 725-727 (1968). 8. DUGGAN, R. E. & DUGGAN, M. B. Residues of pesticides in milk, meat and foods. In: Edwards, L. A., ed., Environmental pollution from pesticides. London, 1973, p. 334. 9. BARNEY, J. E. Pesticide pollution of the air studied. Chemical and engineering news, 47: 42 (1969). 10. HERBST, M. & BoDENSTEIN, G. Toxicology of lindane, In: Ulmann, E., ed. Lindane: monograph of an insecticide. Freiburg, Verlag K. Schillinger, 1972, p. 23. II. LANG, E. P. Tissue distnbution of a toxicant following oral ingestion of the gammaisomer of benzene hexachloride by rats. Journal of pharmacology and experimental therapeutics, 93: 277 (1948). 12. DAVIDOW, B. & FRAWLEY, J. P. Tissue distribution, accumulation and elimination of isomers of benzene hexachloride. Proceedings of the Society for Experimental Biology and Medicine, 76: 780 (1951). 13. HUNTINGDON RESEARCH CENTRE. In: Ulmann, E., ed. Lindane: monograph of an insecttcide. Freiburg, Verlag K. Schillinger, 1972, p. 97. a US Environmental Protection Agency. BHC-Lindane. Washmgton, DC, Criteria and Evaluation DIVIsiOn (unpublished report).
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14. KORANSKY, S. ET AL. Absorption, distnbution and elimination of alpha- and betabenzene hexachloride. Archiv fiir experimentelle Pathologie und Pharmakologie, 244: 564 (1963). 15. PoRADOVSKY, R. ET AL. Transplacental permeation of pesticides during normal pregnancy. Ceskoslovenska gynekologie, 42: 405 (1977). 16. NICHIMURA, H. ET AL. Levels of polychlorinated biphenyls and organochlorine insecticides in human embryos and fetuses. Pediatrician, 6: 45 (1977). 17. CHADWICK, R. W. ET AL. Dehydrogenation, a previously unreported pathway of lindane metabolism in mammals. Pesticide biochemistry and physiology, 6: 575 (1975). 18. CHADWICK, R. W. & FREAL, J. J. The Identification of five unreported lindane metabolites recovered from rat urine. Bulletin of" environmental contamination and toxicology, 7: 137 (1972). 19. NATIONAL RESEARCH CoUNCIL. Drinking water and health. Washington, DC, National Academy of Sciences, 1977, p. 939. 20. 1966 Evaluations of some pesticide res1dues in food. Geneva, WHO, 1967 (WHO/Food Add.f67.32) pp. 126-147. 21. FITZHUGH, 0. G. ET AL Chronic toxicities of benzene hexachloride, and its alpha, beta, and gamma isomer. Journal of pharmacology and experimental therapeutics, 100: 59 (1950). 22. LEHMAN, A. J. Chemicals in food: A report to the Association of Food and Drug Officials. US Association of Food and Drug Ofjictals quarterly bulletin, 16: 85 (1952). 23. LEHMAN, A. J. Chemicals in food: A report to the Association of Food and Drug Officials on current development. Part II, Pesticides. Section V: Pathology. US Association of Food and Drug Officials quarterly bulletin, 16: 126 (1952). 24. HEYROTH, F. F. In: Leland, S. J., Chem1cal Specialities Manufacturers Association. Proceedings of the annual meeting, 6: 110 (1952). 25. PETRESCU, S. ET AL. Studies on the effects of long term admmistratiOn of chlorinated organic pesticides (lindane, DDT) on laboratory white rats. Revue medico-chirurgicale (Jassy), 78: 831 (1974). 26. RIVETT, K. F. ET AL. Effects of feedmg lindane to dogs for periods of up to 2 years. Toxicology, 9: 237 (1978). 27. LEE, B. ET AL Suspected reactions to gamma benzene hexachloride. Journal of the Amencan Medical Association, 236: 2846 (1976). 28. KAZAHEVICH, R. L. State of the nervous system in persons with a prolonged professional contact with hexachlorocyclohexane and products of its synthesis. Vrachebnoe delo, 2: 129 (1974). 29. BESUGLYI, V. P. ET AL. State of health of persons having prolonged occupational contact with hexachlorocyclohexane. Zdravookhranenie Belorussu, 19: 49 (1973). 30. SASINOVICH, L. M. ET AL. Toxic hepatitis due to prolonged exposure to BHC. Vrachebnoe delo, 10: 133 (1974). 31. Pesticide residues in food. Report 1977. FAO Plant production and protection paper, Rome, 1978. 32. MAMETKULIEV, C. H. Study of embryotoxic and teratogenic properties of the gamma isomer of HCH in experiments with rats. Zdravookhranenie Turkmenistana, 20: 28 (1978). 33. PALMER, A. K. ET AL. Effect of linda!le on pregnancy in the rabbit and rat. Toxicology, 9: 239 (1978). 34. TSONEVA-MANEVA, M. T. ET AL. Influence of diazinon and lindane on the mitotic activity and the karyotype of human lymphocytes cultivated in vitro. Biblwtheca haematologia, 38: 344 (1971). 35. BusELMAIR, W. ET AL. Comparative investigation on the mutagenicity of pesticides in mammalian test systems. Mutation research, 21: 25 (1973). 36. GoTo, M. ET AL. Ecological chemistry. Toxizitat von a-HCH in mausen. Chemosphere, 1: 153 (1972). 37. HANADA, M. ET AL. Induction of hepatoma in mice by benzene hexachloride. Japanese journal of cancer research, 64: 511 (1973). 38. NATIONAL CANCER INSTITUTE. Bioassay of lindane for possible carcinogenicity. Washington, DC, 1977 (Technical Report Series, No. 14; Department of Health, Education, and Welfare Publication No. (NIH) 77-814). Federal register, 42: (1977).
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39. THORPE, E. & WALKER, A. I. The toxicology of dieldrin (HEOD). II. In mice with dieldrin, DDT, phenobarbitone. beta-BCH, and gamma-BCH. Food and cosmetics toxicology, 11: 433 (1973). 40. Some halogenated hydrocarbons. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 20). 41. ERIKSSON, M. ET AL Soft-tissue sarcomas and exposure to chemical substances: a casereferent study. British journal of mdustrial medicine, 38: 27 (1981).
4. 7 Methoxychlor 4.7.1 (;eneral aspects Methoxychlor refers to 1, I' (2,2,2-trichloroeth)ll.idene)bis( 4-methoxybenzene). It is practically insoluble in water, but readily soluble in most aromatic organic solvents. Technical methoxychlor consists of about 88% of the p,p'-isomer, the remainder being the o,p'-isomer. Methoxychlor is an insecticide used for the treatment of agricultural crops and livestock. Reviews of this compound are available (1-4) and relevant data are summarized below.
4. 7.2 Routes of exposure 4.7.2.1 Water The half-life of methoxychlor in water is about 46 days. Residues of methoxychlor have occurred in river-water at levels of 2.9-89.1 f.J.g I litre, in lake water at levels up to 0.1 f.J.g/litre, and in effluent from a biological sewage-treatment plant at levels of up to 106 f.J.g/litre. It has been found in tributary streams of Lake Michigan at levels of 2.9-89.1 ng/litre. In finished drinking-water from the Mississippi and Missouri rivers no residues of methoxychlor were detected in 500 samples analysed. 4.7.2.2 Food On the basis of measured residues in food, an average daily intake of 0.5 f.J.g/day was calculated in the USA for the period 1965-70. Methoxychlor showed very little tendency to be stored in tissues or to be excreted in milk (3, 4).
4. 7.3 Metabolism Methoxychlor is rapidly metabolized in rats by the liver, its metabolic products being excreted mainly in the faeces and to a lesser extent in the urine. In mice, 98% of labelled methoxychlor given orally was eliminated within 24 hours. Methoxychlor is mainly degraded by hydrolysis of the methylether group leading to a polar phenol which is rapidly excreted. Dose-dependent storage of methoxychlor occurred in fatty tissues of rats. At a dose of 500 mg/kg of diet, an equilibrium was reached within
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4 weeks; methoxychlor disappeared from the fatty tissue within 2 weeks after the end of exposure. From experiments in animals it may be concluded that the high rate and completeness of methoxychlor metabolism accounts for its low storage and accumulation (1, 2, 4).
4. 7.4 Health effects Methoxychlor is a compound of relatively low acute toxicity. Its oral LD 50 in rats is 3460 mg/kg of body weight. Methoxychlor did not exhibit teratogenicity in rats. It was not mutagenic in bacteria, yeast, or Drosophila melanogaster. Cytogenic and dominant lethal tests in mice were also negative. Methoxychlor was tested for carcinogenicity in mice and in several experiments in rats by oral exposure. The study in mice gave negative results. A suggestion that methoxychlor was hepatocarcinogenic was not confirmed in three studies in rats. The available data did not provide evidence that methoxychlor is carcinogenic in experimental animals (4). An acceptable daily intake for humans of 0.1 mg/kg of body weight, established in 1965, was reaffirmed in 1977 (1, 2). On this basis, the maximum daily intake for a 70-kg man would be 7 mg. Assigning 1% of this value to water and assuming a water consumption of 2 litres per person per day, a guideline value of 30 Jlg/litre can be recommended.
REFERENCES I. Evaluation of the toxicity of pesticide residues in food. (FAOjPL/1965:10/1; WHO/Food
Add./27.65). 2. Pesticide residues m food. Report 1977. FAO Plant Production and Protection Paper. Rome, 1978. 3. Some organochlorine pesticides. Lyon, International Agency for Research on Cancer, 1974 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 5). 4. Some halogenated hydrocarbons. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 20).
4.8. 2,4-Dichlorophenoxyacetic acid 4.8.1 General aspects 2,4-D (2,4-dichlorophenoxyacetic acid) is used as a herbicide for the control of broad-leafed plants and as a plant growth regulator. It is produced commercially by the chlorination of phenol to form 2,4dichlorophenol, which is reacted with monochloroacetic acid to form 2,4-0.
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Commercial formulations are generally composed of salts or esters of the acid. Analyses have shown that dioxins are generally absent. 2,4-D is chemically quite stable, but its esters are readily hydrolysed to the free acid. The herbicide may be rapidly broken down in water. Residues are infrequently found in the soil as the substance is broken down by soil microorganisms and there is reportedly no accumulation. 2,4-D has been detected in streams following aerial application on adjacent forestland, while smaller concentrations ( < 0.1 Jl.g/1) have been reported in potable water sources prior to treatment.
4.8.2 Routes of exposure 4.8.2.1 Drinking-water Information on the levels of 2,4-D in drinking-water is not available. Very low concentrations would be expected; microbial breakdown rapidly reduces the level of 2,4-D in contaminated surface-water. 4.8.2.2 Food Some foods have been shown to be contaminated with low levels of 2,4-D (0.021-0.16 mgjkg) (1). It has been reported that plants treated with 2,4-D produce increased amounts of nitrate. Although this is of concern with regard to the use of the herbicide on edible crops, there is no indication that foodstuff exposed to 2,4-D has become toxic in this manner. 4.8.2.3 Industrial exposure Adverse health effects as the result of industrial exposure to the chemical have been reported (2).
4.8.3 Metabolism 2,4-D is rapidly excreted virtually unchanged in the urine of man and animals. Where the chemical is absorbed, it is distributed in various tissues but not stored. In rats that received 1-10 mg of 2,4-D, there was almost complete excretion in the urine and faeces in 48 hours; at higher doses some accumulation occurred in the tissues. The effect of ingested phenoxyacid herbicides on muscular function may be related to interference with carbohydrate mtltabolism. 2,4-D is eliminated in the milk of cows maintained on pastures treated with 2,4-D or its esters.
4.8.4 Health effects Individuals exposed to 2,4-D through use or manufacture have complained of fatigue, headache, liver pains, loss of appetite, etc.
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However, these claims are subjective. Cases of arterial hypertension and liver dysfunction have also been encountered. Workers exposed to 2,4-D at 0.43-0.57 mgfkg of body weight per day over a period of 0.5-22 years showed no differences in comparison with an unexposed human population. Studies on the carcinogenic properties of this compound have proved inconclusive, because of either inadequate reporting or the small number of animals used. However. indications are that 2,4-D is not a potential carcinogen. In a case-control study on the association between malignant lymphomas and exposure to chlorophenols and phenoxyacids, 7 cases and I control were said to have been exposed only to 2,4-D (relative risk 14.6 with 95% confidence interval 2.9-29.9) (3). The LD 50 of 2,4-D has been estimated to be over 90 mgfkg of body weight.
4.8.4.1 Guideline value The acceptable daily intake of 2,4-D has been established by a joint FAO/WHO meeting (4, 5) at 0.3 mgfkg of body weight. The guideline value for 2,4-D in drinking-water is 0.1 mgflitre, based on toxicity data. However, some individuals may be able to detect 2,4-D by taste or odour at levels around 0.05 mgjlitre. REFERENCES I. NATIONAL RESEARCH CouNCIL. Drinking water and health. Washington, DC, National Academy Press, 1977. 2. Some fumigants, the herbicides 2,4-D and 2,4,5-T, chlorinated dibenzodioxins and miscellaneous industrial chemicals. Lyon, International Agency for Research on Cancer, 1977 (!ARC Monographs on the evaluation of carcinogenic risk of chemicals to humans, vol. 15). 3. BARDELL, L. ET AL. Malignant lymphoma and exposure to chemicals, especially organic solvents, chlorophenols and phenoxy acids: a case-control study. British journal of cancer, 43: 169 (1981). 4. 1974 Evaluations of some pesticide residues in food. Geneva, World Health Organization, 1975 (Pesticide Residues Series, No. 4). 5. 1975 Evaluations of some pesticide residues in food. Geneva, World Health Organization, 1976 (Pesticide Residues Series, No. 5).
5. CHLOROBENZENES Monochlorobenzene is widely used as a solvent and in the manufacture of several chemicals, such as insecticides and phenols. Dichlorobenzenes are important intermediates for dyestuffs. I ,2dichlorobenzene is used as a solvent and pesticide and I ,4-dichlorobenzene as a moth repellant and deodorant. I ,2,4-trichlorobenzene is used as a solvent, dielectric fluid, heat transfer medium and insecticide. Of the tetrachlorobenzenes, I ,2,4,5-tetrachlorobenzene is used as an intermediate in chemical syntheses, e.g., for the production of 2,4,5trichlorophenol (J). Pentachlorobenzenes are not widely used. Hexachlorobenzene has been considered in the section on pesticides and is therefore not included here. Some lower chlorobenzenes are formed as by-products of chlorination of water. Monochlorobenzene has been detected in ground water, surface water, and drinking-water at levels up to 10 !lg/litre (1, 2). Dichlorobenzenes are frequently found in raw water sources at levels of 1-10 !lg/litre or more (3). The main dichlorobenzenes in the aquatic environment are I ,2- and I ,4-dichlorobenzene. In drinking-water these compounds have been found at levels between 0.01 and i!lg/litre. Of the trichlorobenzenes, I ,2,4-trichlorobenzene is detected most frequently; levels in drinkingwater have varied between 0.01 and i!lg/litre (2). The absence of data on the occurrence of the other trichlorobenzenes and of the tetrachlorobenzenes indicates that their concentrations are probably below 0.1 /lg/litre, the limit of detection of the techniques used for their analysis. On the basis of the anticipated maximum levels of chlorobenzenes in drinking-water and an examination of the data on toxicity and odour threshold concentrations (OTC), a selection can be made of the compounds for which the development of a guideline value is required (4, 5). A summary of these data is given in Table 2, page 222. Considering the low levels detected in drinking-water compared with the organoleptic and toxicological properties of these compounds and the paucity of toxicological data, only the following three compounds were selected for further consideration: chi oro benzene; I ,2-dichlorobenzene; I ,4-dichlorobenzene. As Table 2 shows, the organoleptic and the toxicological limits for chlorobenzene are of the same order of magnitude. For the dichlorobenzenes, the odour thresholds are somewhat lower than the toxicologically derived values. 221
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Table 2. Criteria for selecting chlorinated benzenes of toxicological concern Indication of maximum levels Raw water ( JlQ/Iitre) 10 10 1 10 10 Drinking water (llgflitre) 10 1 0.1 1 Indication of OTC ( JlQ/Iitre) 20-100 2-10 20 0.3--30 10 5--30 50 20 400 130
Compound Chlorobenzene 1,2-dichlorobenzene 1,3-dichlorobenzene 1.4-dichlorobenzene 1,2,3-trichlorobenzene 1,2,4-trichlorobenzene 1,3,5-trichlorobenzene 1,2,3,4-tetrachlorobenzene 1,2,3,5-tetrachlorobenzene 1,2.4,5-tetrachlorobenzene -
Indication of toxicity based on no-adverseeffect level ( llQ/Iitre) 5--50 5--50 5--50 5--50
* * * * * 2-20
no dote avo1l1bla. wh1ch ind1cotes thot the levels are probobly below 0 1 I'Q/htre • no data ava111ble on chronic toxic1ty
5.1 Chlorobenzene (monochlorobenzene)0 5.1.1 General aspects Chlorobenzene (monochlorobenzene) is widely used as a solvent and as an intermediate in the manufacture of dyestuffs, pesticides, and other chemicals (1). It can also be formed upon chlorination of water.
5.1.2 Routes of exposure 5.1.2.1 Water Chlorobenzene has been reported in groundwater, surface water, and drinking-water at levels of 0.005-10 Jlg/litre (1, 2). Assuming a water consumption of 2litresfday and an absorption efficiency of 100 %. the daily intake can vary from 0.01 to 20 Jlg/day. 5.1.2.2 Air There are no reports of the compound being detected in ambient air. The only data concerning exposure to chlorobenzene via air are from the industrial working environment. Reported industrial exposures vary from 0.004 to 0.3 mg/litre (2). a The references for monochlorobenzene are listed together with those for dichlorobenzene at the end
of section 5.2.
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223
5.1.2.3 Food No data are available on the intake of chlorobenzene from food. 5.1.3 Metabolism
Chlorobenzene is metabolized in mammals to diphenolic derivatives, probably via epoxide formation. These chlorophenols are then excreted as conjugates. Chlorobenzene may also be metabolized to p-chlorophenylmercapturic acid (2). 5.1.4 Health effects
5.1.4.1 Observations in man Chlorobenzene is irritating to the respiratory system and is a central nervous system depressant (1). Several cases of chlorobenzene intoxication due to inhalation have been reported. There are few, if any, usable human exposure data for chlorobenzene alone. Data are available from studies of exposure to chlorobenzenes in combination with other materials. 5.1.4.2 Observations in other species There is enough evidence to suggest that chlorobenzene causes doserelated target organ toxicity, although data are lacking with respect to an acceptable chronic toxicity study. The no-observed-adverse-effect level in dogs after 3 months was 27.25 mgfk:g of body weight per day. Studies in rats showed no observed adverse effects after 3 and 6 months of doses of 12.5 and 14.5 mg/kg of body weight per day, respectively. Another study in rats showed no observed adverse effects after the administration of 0.001 mgjkg of body weight per day for 7 months (2). 5.1.4.3 Mutagenicity No data are available. 5.1.4.4 Teratogenicity No data are available. 5.1.4.5 Carcinogenicity There is no information in the literature to indicate that monochlorobenzene is carcinogenic. 5.1.4.6 Guideline value The dose levels producing no detectable adverse effects range from 0.001 mgjkg of body weight per day to 14.5 mg/kg of body weight per
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day for rats. Choosing the higher no-detectable-adverse-effect dose of 14.5 mgfkg of body weight per day derived from the short-term study of rats and applying a high safety factor of 1000-10 000, a tentative acceptable daily intake of 0.0015-0.015 mgfkg of body weight can be derived. For a 70-kg man, this value would represent an intake of 0.11 mg/day. Allocating 10% of this dose to water consumption results in a tentative toxicological value in drinking-water of 5-50 J.Lg/litre. As the odour threshold concentration for monochlorobenzene in water is 30 J.Lg/litre, a value that approaches the calculated limits based on health effect considerations, the recommended guideline value for monochlorobenzene in drinking-water is 10% of the threshold odour value, i.e., 3 J.Lg/litre. 5.2 Dichlorobenzenes 5.2.1 (;eneral aspects The dichlorobenzenes (DCBs) are a class of three isomeric halogenated aromatic compounds. 1,2-dichlorobenzene (I ,2-DCB) and 1,3-dichlorobenzene (1,3-DCB) are liquids at normal environmental temperatures, while 1,4-dichlorobenzene (1 ,4-DCB) is a solid; all are relatively volatile. The major uses of 1,2-DCB are as a process solvent in the manufacture of toluene diisocyanate and as an intermediate in the synthesis of dyestuffs, herbicides, and degreasers, but chiefly as an intermediate in the production of pesticides. The primary use of 1,4-DCB is as an air deodorant and insecticide and moth repellant {1). 1,3-DCB may occur as a contaminant of 1,2- or 1,4-DCB, but no information is available concerning its commercial production and use. Both 1,2- and 1,4-DCB are produced almost entirely as by-products during the production of monochlorobenzene. 5.2.2 Routes of exposure The production, use, transport, and disposal of dichlorobenzenes have resulted in widespread dispersal and in contamination of the environment. Dichlorobenzenes have been detected in rivers, groundwater, municipal and industrial discharges, drinking-water, air, and soil. 5.2.2.1 Water 1,2- and 1,4-dichlorobenzenes are frequently found in potable water sources prior to treatment, at levels of 1-10 J.Lg/litre. In drinking-water these compounds have been found at levels of 0.001-1 J.Lg/litre (5).a Assuming a human water consumption of 2 litresjday and an absorption efficiency of 100%, the daily intake from water can vary a CAMPBELL, I. I. Maxzmum acceptable limll in dr1nking water for dichlorobenzenes. US Environmental Protection Agency (Unpublished document, prepared for WHO, 1980).
5.
CHLOROBENZENES
225
from 6 X 10- 3 f,J.g (median )eve) in drinking-water of 3 ng/Jitre) to 6 f,J.g (maximum reported level of total DCB of 3 f,J.g/litre). 5.2.2.2 Air Data on air contamination by DCBs are very limited. Values ranging from 0.002 to 50 mg of 1,2-DCB/m 3 have been measured in outdoor air in California. 1,4-DCB was not detected. In domestic houses in Tokyo, concentrations of 1,4-DCB are much higher, ranging from 105 f,J.gjm 3 (bedroom) to 1700 f,J.g/m 3 (wardrobe). Assuming a daily inspired volume of 20m 3 of air (adult male) and an absorption efficiency by inhalation of 50%, the daily intake can vary from 0.02 mg (lowest suburban concentration reported) to 20 mg (reported in wardrobe air as a result of use of 1,4-DCB). 5.2.2.3 Food Food may also be contaminated by DCB. Pork has been tainted by the presence of 1,4-DCB in the air breathed by the animals; tainting of eggs has been reported when the hens were exposed to 20-38 mg of 1,4-DCB per m 3 of air. 1,4-DCB has also been detected in fish in Japanese coastal waters. No data are available from which to estimate specific exposure to DCBs resulting from the consumption of food.
5.2.3 Metabolism The dichlorobenzenes may be absorbed through the lungs, gastrointestinal tract, and intact skin. The relatively low water solubility and high lipid solubility of halobenzenes favour their penetration of most membranes by diffusion, including pulmonary and gastrointestinal epithelia, brain, hepatic parenchyma, renal tubules, and placenta. Studies of I ,2- and I ,4-DCB given in single doses to chinchilla rabbits by stomach tube showed that 1,2-DCB is mainly metabolized by oxidation to 3,4-dichlorophenol and excreted primarily in the urine as conjugates of glucuronic and sulfuric acids. 1,4-DCB was metabolized mainly by oxidation to 2,5-dichlorophenol and excreted only as glucuronides and ethereal sulfates. In humans, also, 2,5-dichlorophenol is indicated as the principal metabolite of I ,4-DCB.
5.2.4 Health effects 5.2.4.1 Observations in man Most reported cases (16 of 22) of human poisoning by DCBs since 1939 have resulted from long-term exposure, primarily by inhalation of vapours, but some have also resulted from exposure by ingestion (3 of 22) or skin absorption (3 of 22). Most toxic exposures have been
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occupational in nature but some have been due to the use or misuse of DCB products in the home. Most case reports (15 of 22) have involved exposure to agents containing primarily 1,4-DCB and the remainder primarily 1,2-DCB. In a few cases, mixtures including 1,3-DCB were involved. The target body systems or tissues have been one or more of the following: liver, blood (or reticuloendothelial system, including bone marrow and/or immune components), central nervous system, respiratory tract, and integument. The clinical findings in these reports imply a wide spectrum of target organs for the DCBs. For example, at least 17 of the 22 reported clinical cases have involved general toxic or irritative symptoms (e.g., fatigue or weakness, anorexia, weight loss, nausea, headache, irritation, and malaise) and the same number have displayed symptoms or signs indicating involvement of the circulatory system, including blood and/or bone marrow or other reticuloendothelial components (e.g., anaemia, leukaemia, leukopenia or leukocytosis, polynucleosis, bone marrow hyperplasia, leukoblastosis, haemorrhagic tendency, splenomegaly, and jaundice). 5.2.4.2 Observations in other species Intubation of 10 guinea-pigs with 1,2-DCB (50% in olive oil) in single oral doses of 800 mg/kg of body weight resulted in loss of body weight, but all the subjects survived. Doses of 2000 mg/kg of body weight were fatal in all cases. In a test of repeated doses of 1,2-DCB in olive oil emulsified with acacia, groups of white rats were dosed by stomach tube five days a week to give a total of 138 doses in 192 days at dose levels of 18.8, 188, and 276 mg/kg of body weight per day. Positive findings in the high-dose animals included increased liver and kidney weights, decreased spleen weight, and slight to moderate cloudy swelling on microscopic examination of the liver. In the intermediate-dose group, liver and kidney weights were slightly increased. No adverse effects were noted at the low dose level. Two drops of undiluted I ,2-DCB in rabbits' eyes caused pain and conjunctival irritation, which cleared completely within one week. In a chronic toxicity test, rats were given 1,2-DCB at daily doses of 0.001, 0.01 and 0.1 mgjkg of body weight. After 9 months, the high dose disturbed higher cortical function in the central nervous system and caused decreased haemoglobin, thrombocytosis, and neutropenia, and inhibited bone marrow mitotic activity. The dose level of 0.1 mg/kg of body weight was "liminal", and the low dose level (0.001 mgjkg) was "subliminal". White female rats were fed 1,4-DCB in oil (emulsified with acacia) by stomach tube five days a week to give a total of 138 doses in 192 days. At the high dosage level of 360 mgjkg of body weight per day, increased liver and kidney weights, and hepatic cirrhosis and focal necrosis were observed. At the intermediate dose level of 188 mg/kg of body weight per day, increased liver and kidney weights were observed. No adverse effects were noted at the low dose level (18.8 mgjkg per day).
5. 5.2.4.3 Mutagenicity
CHLOROBENZENES
227
Insufficient data are available to permit a decision mutagenicity. 5.2.4.4 Teratogenicity No data are available. 5.2.4.5 Carcinogenicity
regarding
No reports of specific carcinogenicity tests of DCBs in animals or of pertinent epidemiological studies in humans are available. Although no strong direct evidence of the carcinogenicity of DCBs is available, there are some data suggesting that they should be regarded as suspect carcinogens, pending the availability of better data. Available data in rats do not provide evidence for the carcinogencity of 1,4-DCB. No adequate data on 1,2-DCB are available. Epidemiological studies provide inadequate data for evaluating the carcinogenicity of DCBs in humans (6). 5.2.4.6 Guideline value Published values for non-adverse-effect levels range from 0.001 mgjkg of body weight to 13.4 mgjkg of body weight per day. Very different results can be derived depending on which figures are used. Using a no-detectable-adverse-effect dose of 13.4 mgjkg of body weight per day, derived from a short-term study with rats, and applying a safety factor of 1000-10 000, a tentative acceptable daily intake of 0.00134-0.0134 mgjkg of body weight can be calculated for both 1,2and 1,4-DCB. For an adult of 70 kg this represents an intake of 0.11 mgjday. Allocating 10% of this dose to water consump"tion, and assuming a consumption of 21itresjday, a tentative toxicological value for drinking-water of 5-50 J.tg/litre, can be derived. These values are in excess of the levels quoted as threshold odour values. For the 1,2-isomer the threshold odour value is given as approximately 3 J.tg/litre; I 0% of this value, i.e., 0.3 J,tgjlitre, is recommended as a reasonable value. Similarly, as the threshold odour value for the 1,4-isomer is only I J,tgjlitre, a recommended value is 0.1 J.lgflitre.
REFERENCES I. Toxicological appraisal of halogenated aromatic compounds following groundwater pollution. Copenhagen, WHO Regional Office for Europe, 1980.
2. Ambient water quality criteria for chlorinated benezenes. Washington, DC, US Environmental Protection Agency, 1979 (EPA 440/5-80-028). 3. VAN GEMERT, J. L. & NETTENBREYER, A. H. Compilation of odour threshold values in air and water. Leidschendam, The Netherlands. National Institute for Water Supply, 1977 (2260 AD).
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4. ZoETEMAN, B. C. J. Sensory assessment of water quality. Oxford. Pergamon Press, 1980. 5. Ambient water quality criteria for dichlorobenzenes. Washington, DC, US Environmental Protection Agency, 1979 (440/5-80-039). 6. Some industrial chemicals and dyestuff. Lyon, International Agency for Research on Cancer, 1982 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 29).
6. BENZENE AND LOWER ALKYLBENZENES Benzene and lower alkylbenzenes, such as toluene and ethylbenzene, are widely used in the chemical industry as intermediates in the production of various chemicals, e.g., phenol and cyclohexane. Lower alkylbenzenes are components of gasoline and are used as solvents for paints and coatings. Because of the suppression of the evaporation process, benzene and lower alkylbenzenes can be present in groundwater at higher levels than those usually found in surface-water. As a result of spills and dumping of chemical wastes, concentrations of the chemicals in groundwater up to levels of several mg per litre have recently been found. Levels in drinking-water do not generally exceed ltLg/litre. Except for benzene, there seems to be no potential health risk from the levels of alkylbenzenes found in drinking-water.
6.1 General aspects Benzene and toluene are produced mainly from petroleum or as byproducts in the manufacture of gas and coke. They are used in large quantities by the chemical industry, the three major uses being for the production of styrene, cumene (used to manufacture phenol and acetone), and cyclohexane (used in manufacturing nylon). Much of the toluene produced is used in the production of benzene. Relatively minor (but still significant) quantities are used in a variety of industries (e.g., plastics, paints, detergents, and as gasoline additives), either as intermediates for many syntheses, or as solvents. Other alkylbenzenes (ethylbenzene, xylenes) are also used extensively as solvents or as chemical intermediates.
6.2 Routes of exposure Benzene and the lower alkylbenzenes are volatile and comparatively unreactive in the environment. They become widely dispersed in the environment through movement of air masses and there is continuous recycling between air and water bodies through rain and volatilization from water surfaces. Ultimately, degradation is likely, as a result of biological and microbial oxidation.
229
230 6.2.1 Air
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Levels of benzene and toluene in urban air are commonly of the order of 100 JLg/m\ the major part of which originates from gasoline-related emissions (automotive emissions, gasoline handling). Solvent losses and emissions from industrial activjties also contribute. Average levels in the air of a number of cities worldwide have been compiled and found generally to be less than 90 JLg/m 3 for benzene (1) and 112.5-150 JLg/m 3 for toluene (2). A mean exposure level to benzene in the USA was calculated to be 9.5 Jlgjm 3 (range, < 3.5 Jlgjm 3 to 1 mgjm 3 ) (3).a 6.2.2 Food There are few data on levels of benzene or toluene in foods, although there are indications that benzene occurs naturally in some fruits, fish, vegetables, nuts, dairy products, beverages, and eggs (3), and toluene is found in fish from areas adjacent to petrochemical activities (4). 6.2.3 Water The major sources of benzene and toluene in water are atmospheric deposition (through rain and snow) and chemical plant effluents (with minor contributions from urban runoff and sewage plants) (5). Benzene levels of up to 179 Jig/litre have been reported in chemical plant effluents, although levels in finished drinking-water are generally much lower; levels of 0.1-0.3 Jig/litre have been found in four city drinking-water supplies in the USA (6), whilst levels below 0.01 Jig/litre have been reported in Canada (1). Levels of up to 19 Jlg of toluene per litre have been repor~d (2). Volatile hydrocarbons, such as benzene and toluene, would be expected to evaporate rapidly into the atmosphere from bodies of water (half-lives of 37.3 minutes and 30.6 minutes respectively for benzene and toluene at 25 oq (2), a fact that may explain why levels in groundwater are occasionally much higher than those in surface-water. Noteworthy is one observation in which benzene in finished water appeared to originate from anthracite filters used in water treatment (7). 6.2.4 Occupational exposure Levels of benzene in the air of industries in which benzene is produced or used are generally three orders of magnitude higher than those in the general atmospheric environment; levels are commonly in the range 0.39 mgjm 3 (time-weighted average) (1). Individuals exposed to a benzene concentration of 3.0 mgjm 3 (time-weighted average) receive a yearly dose about thirty times higher than non-occupationally exposed individuals (1). a In the original publications, the concentrations were given m parts per bilhon. Conversion factors of I ppb = 3.0 JJgfm' for benzene and I ppb = 3.75 JJg/m' for toluene have been used.
6.
BENZENE AND LOWER ALKYLBENZENES
231
6.2.5 Exposure estimates for man The total uptake of benzene by urban dwellers from background (i.e., non-occupational) exposures has been estimated to be about 125 mg/year, 90 mg of which comes from food (1). It should be noted that the information on benzene levels in food is very scanty, so this background level should be considered only as an approximate reference point. Furthermore, this estimate does not take into account possible sporadic uses of benzene-containing products by the consumer. Nevertheless, it does suggest that levels commonly found in drinkingwater are minimal (compared with intake from food and air) and are probably only of marginal importance. Similar conclusions apply to toluene.
6.3 Metabolism The metabolism and excretion of benzene in humans and animals appear to follow similar pathways (8). Regardless of the route of administration, predominantly unchanged benzene is eliminated in the expired air. Conjugated metabolic products, typically large quantities of phenol accompanied by smaller amounts of catechol, hydroquinol, and hydroxyhydroquinol are excreted in the urine. The liver is the major site of both oxidation and conjugation. Toluene is rapidly and extensively metabolized to hippuric acid, which is excreted in the urine; most of a dose of toluene is almost completely eliminated within 12 hours as unchanged toluene in expired air or as hippuric acid (2).
6.4 Health effects A number of reviews on benzene toxicity have appeared recently (1, 9-11). Acute exposure to benzene results in central nervous system depression. Although most studies of benzene toxicity have involved exposure by the inhalation route, limited animal studies suggest that exposure by other routes of administration leads to similar sequelae. Chronic exposure to benzene leads to haemopoietic tissue changes in the form of anaemia and leukopenia. Epidemiological studies and several case reports suggest a relationship between benzene exposure and leukaemia, evidence that has led to its categorization as a human carcinogen by a Working Group convened by the International Agency for Research on Cancer (1 2). In animals, benzene exposure has been shown to affect immunological defence mechanisms and in numerous studies it has been shown possible to induce chromosome damage, both in exposed animals and using cell culture techniques. Acute exposure to toluene by the inhalation route results in central nervous system depression (6). However, toluene appears to produce no
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irreversible tissue injury and its major metabolite, benzoic acid, is considered relatively nontoxic. Although most toxicity studies have involved inhalation exposure, studies on chronic oral administration have been conducted in rats, the highest dosage group being given toluene at 590 mgjkg of body weight five times weekly for 193 days, after which no adverse effects were observed (5). Toluene does not appear to be teratogenic, mutagenic, or carcinogenic (2), although one study in mice reported teratogenic effects when massive doses were used (13). There are numerous studies involving long-term industrial exposure to toluene without any detectable changes in blood characteristics or liver damage (6). The above data indicate that toluene is of relatively low toxicity and a maximum permissible concentration in water of 14.3 mg/litre has been calculated (2). In view of the fact that this figure is very much higher than the levels found in water it was decided not to recommend a guideline value for toluene. For an excess cancer risk of I in 10 5 per lifetime and rounding to the nearest decade, a guideline value of 10 J.Lg/litre is recommended for benzene in drinking-water.
REFERENCES 1. HOLLIDAY, M. ET AL. Benzene: Human health implications of benzene at levels found in the Canadian environment and workplace. Ottawa, Health and Welfare Canada, 1978 (Environmental Health Directorate Report No. 79-EHD-40). 2. Ambient water quality criteria for toluene. Washington, DC, US Environmental Protection Agency, 1980 (EPA-440/5-80-75). 3. MARA, S. J. & LEE, S. S. Assessment of human exposure to atmospheric benzene. Washington, DC, US Environmental Protection Agency, 1978 (EPA-450/3-78-031). 4. OGATA, M. & MIYAKE, Y. Identification of substances in petroleum causing objectionable odour in fish. Water research, 7: 1493 (1973). 5. WOLFE, M. A. ET AL. Toxicological studies of certain alkylated benzenes and benzene. Archives of industrial health, 14: 387 (1956). 6. NATIONAL RESEARCH CoUNCIL. Drinking water and health. Washington, DC, National Academy of Sciences, 1977. 7. SMILLIE, R. D. ET AL. Low molecular weight hydrocarbons in drinking water. Journal of environmental health, A13: 187 (1978). 8. RuscH, G. M. ET AL. Benzene metabolism in benzene toxicity: a critical evaluation. Washington, DC, American Petroleum Institute, 1977, pp 23-36. 9. LASKIN, A. & GoLDSTEIN, B. D., ed. Benzene toxicity: a crillca/ evaluation. Washington, DC, American Petroleum Institute, 1977, 147 pp. 10. NATIONAL REsEARCH CoUNCIL Health effects of benzene: a review. Washington, DC, National Academy of Sciences, 1976, 23 pp. 11. Some antithyroid and related substances, nitrofurans and industrial chemicals. Lyon, International Agency for Research on Cancer, 1974 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 7). 12. Chemicals and industrial processes associated with cancer in humans. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of carcinogenic risk of chemicals to man, Suppl. 1). 13. NAWROT, P. S. & STAPLES, R. E. Embryo-foetal toxicity and teratogenicity of benzene and toluene in the mouse. Teratology, 19: 41A (1979).
7. PHENOL AND CHLOROPHENOLS Chlorophenols are used as biocides and are found as a result of chlorination of water containing phenol. Chlorophenols are well known for their low taste and odour thresholds. Taste thresholds for the most odorous compounds (mono- and dichlorophenols) are as low as 1 JLg/litre. For aesthetic reasons, therefore, individual (chloro)phenols should not, as a general rule, be present in drinking-water above the 0.1 JLg/litre level; exceptions are phenol and pentachlorophenol, which have taste thresholds of 100 JLgjlitre. Provided that no chlorination is applied, phenol may therefore be accepted at levels up to 100 JLg/litre in drinking-water. For chlorophenols, in cases where a guideline value of 0.1 JLg/litre cannot be achieved, it must be borne in mind that some chlorophenols exert toxic effects at somewhat higher concentrations. Without extreme concentration procedures, spectrophotometric techniques will only detect chlorophenols at levels above 1 JLg/litre, which is higher than the taste threshold concentration of several chlorophenols, If chlorophenols are suspected of being involved in taste problems, direct taste assessment by taste panels, and analytical determination by chromatographic techniques will need to be used. The best approach to controlling pollution of drinking-water by chlorophenols is to prevent the contamination of the source water by phenol and chlorinated phenolic pesticides. When high phenol levels are present in the raw water these should be reduced, as far as possible, before chlorination is applied. Lower substituted chlorophenols, once present in the water, may be removed by oxidation processes, while the higher substituted ones can be effectively removed only by activated carbon adsorption. 7.1. Chlorinated phenols of toxicological significance 7.1.1 (;eneral aspects Chlorinated phenols are known to be present in drinking-water, resulting either from contamination of raw water sources or from chlorination of water containing phenolic compounds. Phenols may be present in raw water owing to the discharge of wastewaters from coke distillation plants, the petrochemical industry, and numerous other industries where phenols serve as intermediates. They are also present in municipal wastewaters. When water containing phenol itself is 233
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chlorinated, the main reaction products are 2- and 4-chlorophenol, 2,4-dichlorophenol, and 2,4,6-trichlorophenol. 2,4-dichlorophenol is produced commercially as an intermediate in the manufacture of the herbicide 2,4-D, related biocides, and pentachlorophenol. This last compound is used as a wood preservative, 2,4,5-trichlorophenol as a fungicide, and 2,4,6-trichlorophenol as an antiseptic. 2,4,6-trichlorophenol is also a major metabolite of the insecticide lindane. The major uses of 2,3,4,6-tetrachlorophenol are as an insecticide and wood preservative.
7.1.2 Occurrence Contaminated raw water, including groundwater, may contain 1-lOJ.lg of phenol and mono- and dichlorophenols per litre. Similar levels have been reported in drinking-water. Tri- and tetrachlorophenol have been detected in raw water at levels of 1-10 J.tg/litre and occasionally higher. The concentrations usually found in drinking-water are, however, 1 or 2 orders of magnitude lower. In view of the apparently low frequency of their occurrence in water, chlorophenols other than those shown in Table 3 can, for the time being, be excluded from further consideration with respect to effects on public health (1 -4).
7 .1.3 Preliminary screening of health effects From a consideration of the anticipated maximum levels in drinkingwater and the existing literature on the toxicity and organoleptic data, it is evident that for several chlorinated phenols, toxicity-based limits will be much higher than taste- and odour-based limits. To be able to select the compounds of primary toxicological concern, a preliminary compilation of the relevant data is presented in Table 3 (5-7). As this table shows, the limits for some chlorinated phenols calculated from consideration of their toxicity are much higher than organoleptic considerations allow. 2,4,6-trichlorophenol and pentachlorophenol are exceptions and toxicological limits for these compounds have been calculated. Limits for the other chlorophenols, as well as phenol itself, are proposed on the basis of organoleptic considerations. 7 .1.4 Complementary organoleptic considerations It is a matter of concern whether the low organoleptic thresholds of mono- and dichlorophenols can in practice guarantee the safety of a tasteless drinking-water, when little is known about the toxicity of several of the compounds. Consideration of the toxicology of those chlorophenols that are well documented indicates that if water is free
Table 3. Criteria for setting limits for chlorinated phenols of toxicological concern Indication of maximum levels Raw water ( llg/1) 100 10 10 10 10 1 1 <0.1 10 Drinking water ( llQ/1) 1 1 1 10 1 < 0.1 I < 0.1 1 Odour threshold concentratio.,. ( llg/1) 1000 1 1 1 10 100 100 1000 1000 Taste threshold concentration• ( llgfl) 100 1 1 1 1 1 1 1 100
Typical criterion levels based on Toxicity ( llg/1) 3000 Carcinogenicity ( llg/1)
;-.J
., :z: 0
Compound Phenol 2-Chlorophenol 4-Chlorophenol 2.4- Dichlorophenol 2,6-Dichlorophenol 2.4.5-Trichlorophenol 2.4,6-Trichlorophenol 2,3.4,6-Tetrachlorophenol Pentachlorophenol lnauff1c1ent.dlta ava1lable
z z
m
t""
_b
> 0 (j
3000 2600
, 0 0
:z: t""
21
12
-
0 ., :z:
m
z
"'
t""
• Thresholds reported 1n the literature vary considerably Values h1gher and lower than those 1nd1cated have been reported
b 2,3,7,8-tetrachlorod•benzo-p-d•oxm (TCDD) ,san 1mpunty of techmcal2.4,5-tnchlorophenol In the environment the compounds behave differently and they should therefore be treated separately
w
N
VI
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from taste and odour, it is unlikely to present any direct health risks due to chlorinated phenols. However, it must be realized that in general the absence of any adverse taste or smell does not guarantee the safety of drinking-water. Water with a noticeable taste and odour should be investigated for the possible presence of chlorophenolic compounds.
7.2 2,4,6-Trichlorophenol 7.2.1 General aspects 2,4,6-trichlorophenol is a yellow solid with a melting point of 69.5 oc and a boiling point of 246 o C. It is slightly soluble in water ( < 0.1 g per 100 ml) (8) and soluble in organic solvents. It is manufactured for use as a wood preservative, bactericide, and fungicide. 2,4,5-trichlorophenol is one of the products formed when drinking-water containing low levels of phenol is disinfected using chlorine. 7 .2.2 Routes of exposure 7 .2.2.1 Water 2,4,6-trichlorophenol has been detected in river-water at levels up to 1 Jig/litre. In 1978, the River Rhine in the Netherlands contained levels ranging from 0.04 to 0.63 Jig/litre. Similar concentrations of 2,4,5trichlorophenol were also found by Wegman.a Levels in drinking-water are generally lower but data are scarce. When water containing phenol is chlorinated, concentrations of 2,4,6-trichlorophenol up to a level of a few micrograms per litre may be formed. It is estimated that the maximum exposure of a 70-kg adult from a daily intake of 2 litres of water containing 1 Jig/litre of 2,4,6-trichlorophenol would be 0.00003 mg/kg of body weight per day. 7.2.2.2 Food Another possible route of ingestion is from dairy products since chlorophenolic antiseptics are widely used in the dairy industry (9). 1,3,5-trichlorobenzene has been shown to be metabolized to 2,4,6trichlorophenol (10). Pentachlorocyclohexene is also converted to 2,4,6trichlorophenol in corn and pea plants (1 1). Ingestion can take place via the consumption of fish and shellfish. No quantitative data are available on inhalation studies. In view of the paucity of quantitative data on other routes of exposure, it has been assumed that water is the major route of ingestion. a WEGMAN, R C. C. Unpublished data of the NatiOnal lnstttute for Pubhc Health, Btlthoven, The Netherlands, 1980.
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7 .2.3 Metabolism 2,4,6-trichlorophenol is rapidly cleared from the body, predominantly in the urine (1 2). Very little further is known of the metabolism of this compound.
7 .2.4 Health effects The intraperitoneal LD 50 for rats, using olive oil as the solvent for 2,4,6-trichlorophenol, is 276 mgjkg of body weight (1 3). As with other chlorinated phenols, 2,4,6-trichlorophenol is capable of increasing body temperature and produces convulsions at high doses. 7.2.4.1 Mutagenicity 2,4,6-trichlorophenol increased the mutation rate in a strain of Saccharomyces cerevisiae (14) but was not found to be mutagenic in the Salmonella mammalian microsome Ames test, with or without metabolic activation (15). 7.2.4.2 Carcinogenicity In studies with F344 rats and B6C3-Fl mice at the National Cancer Institute in the USA (16), the compound was found to increase tumour incidence. In male rats, dose-related increases in lymphoma and leukaemia were observed. Leukocytosis and monocytosis of the peripheral blood and hyperplasia of the bone marrow occurred in both male and female rats. In both the male and female mice, the incidence of hepatocellular carcinomas or adenomas was increased significantly in a dose-related manner. Epidemiological data are inadequate for evaluating the carcinogenicity of 2,4,5- and 2,4,6-trichlorophenols (8, 17). The guideline value, calculated via the linear multistage extrapolation model, assuming a lifetime cancer risk of I per I 00 000, is 12 J,tg/litre whereas the quoted threshold taste level is only 1 J,tgjlitre. To provide a drinking-water of acceptable potability a guideline value of 0.1 J.tg/litre is recommended.
7.3 Pentachlorophenol 7.3.1 (;eneral aspects Pentachlorophenol is a widely used fungicide and wood preservative with an estimated world production of 40 000 tonnes per annum. It has a melting point of 190 °C, a boiling point of 310 °C, and a water solubility of 14 mgjlitre at 20 °C.
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7.3.2 Routes of exposure Contaminated surface-water can contain up to 10 11g of pentachlorophenol per litre. In the River Rhine in the Netherlands in 1978, levels ranged from 0.15 to 1.5J1.g/litre. Levels in drinking-water are generally well below I 11g/litre. Levels in worms of 0.2 mgjkg of body weight and in mammals of 0.1-1.3 mg/kg of body weight have been reported.
7.3.3 Metabolism Pentachlorophenol is well absorbed from the gastrointestinal tract and may also be absorbed through the skin. Most of the absorbed pentachlorophenol in humans is excreted unchanged in the urine. Approximately 20 'lo of the systemic dose is dechlorinated to form tetrachlorohydroquinone and trichloro-1-hydroquinone in rats and mice.
7.3.4 Health effects The acute oral LD 50 of pentachlorophenol for rats is 27 mg/kg of body weight. Clinical signs of poisoning include profuse sweating, thirst, elevated temperature, rapid pulse and respiration, and ultimately cardiac arrest. Some cases of Hodgkin's disease and leukaemia have been reported in woodworkers using pentachlorophenol (18), but epidemiological studies are inadequate for evaluation (19, 20). At least 30 fatal cases of pentachlorophenol poisoning have been reported. Pentachlorophenol has been shown to cause chloracne in rabbits. Furthermore, it produces damage to kidneys and liver of experimental animals while similar indications have been obtained for occupationally exposed populations. Pentachlorophenol has been shown to be embryotoxic and fetotoxic in experimental animals. It also increases the frequency of mutation of yeast but showed no mutagenicity in the Ames test. No carcinogenicity has been demonstrated in experimental animals. The available studies on carginogenicity in experimental animals are inadequate (8). In the USA, the National Research Council (21) has calculated an ADI of 3 11g/kg of body weight per day for pentachlorophenol. Assuming a 70-kg man drinks 2 litres of water per day, and assigning 10% of the ADI to water, a guideline value of lOJ1.g/litre may be calculated for drinking-water.
REFERENCES I. Ambient water quality criteria for phenol. Washington, DC, US Environmental Protection Agency, 1980 (EPA 440(5-80-066). 2. Ambient water quality criteria for chlorinated phenols. Washington, DC, US Environmental Protection Agency, 1980 (EPA 440/5-80-032).
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239
3. Ambient water quality criteria for 2.4-dichloropheno/. Washington, DC, US Environmental Protection Agency, !980 (EPA 440/5-80-042). 4. Ambient water quality criteria for pentachlorophenol. Washington, DC, US Environmental Protection Agency, 1980 (EPA 440/5-80-065). 5. VAN GEMERT, L. J. & NETTENBREIJER, A. H. Compilation of odour threshold values in air and water. Leidschendam, The Netherlands, National Institute for Water Supply, 1977. 6. ZoETEMAN, B. C. J. Sensory assessment of water quality. Oxford, Pergamon Press, 1980. 7. DIETZ, F. and TRAUD, J. Geruchs- und Geschmacks-Schwellenkonzentrationen von Phenolkorpem. GWF-Wasserjabwasser, 119:H6 318 (1978). 8. Some halogenated hydrocarbons. Lyon, International Agency for Research on Cancer, 1979 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans. vol. 20). 9. STANNARD, D. J. & Sc011ER, A. The determination of phenol residues in dairy products. New Zealand journal of dairy science and technology, 12: 140 (1977). 10. KoHLI, J. ET AL. The metabolism of higher chlorinated benzene isomers. Canadian journal of biochemistry, 54: 203 (1976). II. MozA, P. ET AL. Beitrage zur okologischen chemie LXXXIX Orientierende versuche zum metabolisms Von-pentachlorocyklohex-1-en in hoheren pflanzen in hydrokultur. Chemosphere, 6: 255 (1974). 12. KORTE, F. ET AL. Ecotoxicologic profile analysis, a concept for establishing ecotoxtcologic priority list for chemicals. Chemosphere, 7: 79 (1978). 13. FARQUHARSON, M. E. ET AL. The biologtcal action of chlorophenols. British journal of pharmacology, 13: 20 (1958). 14. FAHRIG, R. ET AL. Genetic activity of chlorophenols and chlorophenol impurities. In: Rao, K. R., ed., Pentachlorophenol: chemistry, pharmacology and environmental toxicology. New York, Plenum Press, 1978. 15. RASANEN, L. ET AL. The mutagenicity of MCPA and its soil metabolites, chlorinated phenols, catechols and some widely used slimicides in Finland. Bulletin of environmental contamination and toxicology, 18: 565 (1977). 16. NATIONAL CANCER INSTITUTE. 1979 Bioassay of 2,4,6-trich/oropheno/ for possible carcinogenicity. Washington, DC, US Department of Health, Education and Welfare, 1979 (Technical Service Report Series, No. 155). 17. THEISS, J. C. ET AL. In: Long-term hazards of polychlorinated dibenzodioxins and polychlorinated dibenzofurans. Lyon, International Agency for Research on Cancer 1978 (IARC Internal Technical Report No. 78j001). 18. GREEN, M. H. Familial and sporadic Hodgkin's disease associated with occupational wood exposure. Lancet, 2: 626 (1978). 19. HARDELL, L. & SANDSTROM, A. Case-control study: Soft tissue sarcomas and exposure to phenoxy acetic acids or chlorophenols. British journal of cancer, 39: 711 ( 1979). 20. ERIKSSON, M. ET AL. Soft-tissue sarcomas and exposure to chemical substances: a case reference study. British journal of industrial medicine, 38: 27 (1981). 21. NATIONAL RESEARCH CoUNCIL. Drinking water and health. Washington, DC, National Academy of Sciences, 1977.
8. TRIHALOMETHANES Trihalomethanes in drinking-water occur principally as products of the reaction of chemicals used in oxidative treatment reacting with the naturally occurring organic materials present in the water. Their formation is particularly associated with the use of chlorine. The four most frequently occurring trihalomethanes are chloroform, bromodichloromethane, dibromochloromethane, and bromoform. The total concentration of these four trihalomethanes in drinking-water may vary up to 1000 flg/litre, but it is frequently less than 100 flg/litre. Chloroform has been shown to cause cancer in two species of laboratory animal. The three bromine-containing trihalomethanes are only now being subjected to lifetime cancer bioassay tests similar to those in which chloroform was shown to be a carcinogen. These other trihalomethanes are, however, known to be more active than chloroform in the Ames Salmonella test for mutagenesis. It is important to recognize that chlorine is an effective water disinfectant and the hazards of disease arising from microbiological contaminants resulting from incomplete disinfection are substantial. This is particularly true in developing countries where it is estimated that waterborne disease causes thousands of deaths per day. Chlorine is the most convenient and easily controlled disinfectant and is widely used.
8.1 General aspects Trihalomethanes are halogen-substituted, single-carbon compounds having the general formula CHX 3 , where X may be fluorine, chlorine, bromine, or iodine, or combinations thereof. With respect to drinkingwater contamination, discussion may appropriately be confined to four members of the group: chloroform (CHC1 3 ), bromodichloromethane (CHBrC1 2 ), dibromochloromethane (CHBr 2 Cl), and bromoform (CHBr 3 ). Chloroform is the most commonly encountered of these and available information pertains almost exclusively to this substance. The most important use fot chloroform is as the starting material for the manufacture of chlorodifluoromethane, which is used as a refrigerant, aerosol propellant, and in the synthesis of polytetraftuoroethene. Chloroform is an important solvent and degreasing agent. It has also been used in small quantities as an anaesthetic, in liniments, permanent wave lotions, dentifrices, and fumigants, and as the active ingredient and preservative in antitussive formulations (1, 2). The 240
8.
TRIHALOMETHANES
241
estimated world production of chloroform in 1973 was 245 x 10 6 kg (3). Bromoform is used in industry as a gauge fluid, as a heavy liquid in solids separation, and as an intermediate in the synthesis of other chemicals (1). There are no known commercial applications for the mixed halogen derivatives. 8.2 Routes of exposure 8.2.1 Water Trihalomethanes occur in water primarily from reactions between chlorine (and adventitiously present bromide ion) and naturally occurring organic compounds. Data gathered in many countries (4-6) have shown that the levels of trihalomethanes in finished water that has been chlorinated are generally very much higher than the levels in raw water, in which they are often undetectable. Table 4 shows the levels found in a survey of potable waters in the USA. Other surveys have given similar results. Table 4. Trihalomethane content of finished waters in the USA Compound No of locations Range of concentrations (mgjlitre) Med1an concentration (mgjlitre)
Chloroform Bromod1chloromethane Chlorodibromomethane Bromoform
80 78 72
26
0 0001-0.311 0.0003--0.116 0.0004-0 110 0.0008--0.092
0.021 0.006 0.0012 <0.005
Studies have indicated that, for a given chlorine dose, the rate (and hence degree) of trihalomethane formation is increased at higher humic acid concentrations, higher temperatures, and higher pH (6). Trihalomethane formulation has been shown to proceed within the distribution system provided that a free chlorine residual exists (7). The literature contains no information on the occurrence of bromodichloromethane, chlorodibromomethane, or bromoform. 8.2.2 Air Chloroform has been detected in rural atmospheres at levels between 100 and 180 ng/m 3 (8). Maximum levels of chloroform found in the atmosphere at any one location range from <0.05 to 73.5 p.g/m 3 ; mean levels are between 0.045 and 5.0 p.gjm 3 (9). Atmospheric concentrations of chloroform at ground level at sites in the northern and southern hemispheres were 130 ngjm 3 and <15 ng/m 3 , respectively (10).
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8.2.3 Food Two specimens of barley treated with a gaseous fumigant mixture containing chloroform, after airing at 17 ac and 30 °C, respectively, were found initially to contain chloroform residues of 123 and 132 mgjkg. After 60 days it was found that residues had disappeared from the specimen aired at 30 ac but were still present at a level of 16 mg/kg in that aired at 17 °C. Corn and sorghum behaved in a similar manner. The following concentrations of chloroform were found in foodstuffs in the United Kingdom in 1973: dairy produce, 1.4-33 mg/kg; meat, 1-4 mg/kg; oils and fats, 2-10 mg/kg; beverages, 0.4-18 mgjkg; fruits and vegetables, 2-18 mg/kg (JJ).
8.3 Metabolism Exposure to air containing chloroform in a concentration of 13.231.8 gjm 3 for 3-10 minutes resulted in 73 % absorption (12). Inhaled chloroform rapidly enters the bloodstream and is transported to the tissues. In mice, body fat was found to be the important storage site for chloroform; lesser amounts were found in the brain, lungs, kidneys, muscles, and blood. Metabolism of chloroform takes place in the liver, and whole-body autoradiography studies have shown the gradual transfer of radioactivity from the fat storage depots to the liver (13). Chloroform appears to be capable of crossing the human placental barrier since concentrations of chloroform in cord blood were found to be higher than those in the maternal blood (14). In humans, up to 50.6% of an oral dose (7 mgjkg of body weight) was metabolized to C0 2 but there was considerable variation between individuals and up to 68.3% of the ingested chloroform was expired unchanged (15). Much of the absorbed chloroform is eliminated during its first passage through the liver and lungs and only 50-65% of a 500-mg dose was found to be available for general circulation to the rest of the body (16). Of a single 500-mg dose ingested by human volunteers, 18-67% was exhaled unchanged within 8 hours (15). The major metabolites of chloroform are excreted through the lungs (as C0 2 ) or through the kidney (as inorganic chloride) (17).
8.4 Health effects 8.4.1 Toxicity This discussion of the health effects of trihalomethanes will deal mainly with chloroform, which, because of its former use as an inhalation anaesthetic and because it is the predominant trihalomethane in drinking-water, has received the most intensive study. The toxic effects of the other trihalomethanes are likely to be similar to those of chloroform.
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243
Chloroform is a central nervous system depressant. It also affects liver and kidney function. The immediate effect of chloroform intoxication is loss of consciousness, which may be followed by coma and death (18, 19). Renal damage is noted 24-48 hours after exposure and hepatic injury is seen after 2-5 days; thus, symptoms of poisoning may occur several days after recovery from chloroform anaesthesia (18). The mean lethal dose in man is considered to be about 44 g or 630 mg/kg of body weight for a 70-kg man; however, ingestion of more than 250 g of chloroform has been survived (20). The lowest published lethal dose of chloroform in man is 210mgjkg of body weight (21). Ingestion of as little as 440 mg causes gastric irritation and increased peristalsis, as well as some local narcosis in the intestinal tract (19). Except for studies on the carcinogenicity of chloroform, few investigations of the chronic toxicity of chloroform have been carried out. The long-term oral administration of chloroform at a dose of 0.4 mg/kg of body weight did not produce any changes in the investigated indices in albino rats, and the only effect of this dose on guinea-pigs was an increase of vitamin C in the adrenals (22). The safety of adding chloroform to toothpaste and mouth-rinse was assessed in two long-term studies, involving 229 human subjects (23). Daily consumption of chloroform was estimated to be 0.34-0.96 mg/kg of body weight over a 1-5-year period. Results from this study showed no hepatotoxicity based on liver function tests. Reversible hepatotoxicity was the only effect observed in a 47-year-old man who, each day for 10 years, consumed between 12 and 20 ounces (336-560 g) of a chloroformcontaining cough suppressant; his daily dose of chloroform was estimated to be between 23 and 37 mg/kg of body weight (24). Bromoform is considered to produce toxic symptoms that are similar to those of chloroform. An LD 50 of 1820 mg/kg of body weight was reported when bromoform was administered to mice by the subcutaneous route (25). There appears to be little information on the toxic effects of the other trihalomethanes. A maximum tolerated dose of bromodichloromethane was determined in strain A/st male mice to be 100 mgjkg of body weight; the dose was injected intra peritoneally 6 times over a 2-week period. The maximum tolerated dose of bromoform, obtained using the same protocol, was also 100 mg/kg of body weight (26).
8.4.2 Carcinogenicity A study was conducted in the USA by the National Cancer Institute (27). Chloroform dissolved in corn oil was administered by gavage to
Osborne-Mendel rats and B6C3-Fl mice at two dose levels five times per week. Dose levels of 90 or 180 mg/kg of body weight were given to the male rats for 78 weeks; the female rats received doses of 125 or 250 mg/kg for the first 22 weeks and the same dose as the males thereafter. After Ill weeks the rats were killed and a statistically significant incidence of kidney epithelial tumours was found in the males
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(24% in the high-dose and 8% in the low-dose groups) but not in the females. There was an increase in thyroid tumours in the female rats but it was not considered significant. The male mice first received doses of 100 or 200 mg/kg of body weight, and the females were initially dosed with 200 or 400 mgjkg. After 18 weeks, the doses were changed to 150 and 300 mg/kg for males and to 250 and 500 mg/kg for females. Highly significant increases in hepatocellular carcinoma were found in both sexes: 98% of the males and 95% of the females at the high dose and 36% of the males and 80% of the females at the low dose. Nodular hyperplasia was a frequent finding in low-dose male mice that had not developed hepatocellular carcinoma. It should be emphasized, however, that the doses used in the NCI studies were extremely high and, as a greater than 10% weight loss was observed in the animals, can be considered higher than a true maximum tolerated dose.
8.4.3 Epidemiology Cantor et al. (28) looked at the association between the rates for 16 different cancers and the levels of trihalomethanes in the drinking-water; the chloroform and non-chloroform components were studied separately. Exposure information came from the National Organics Reconnaissance Survey and the EPA Region V Survey of 1975. Seventy-six counties in which more than 50% of the population was served by the measured water supply were included in the study. The most consistent finding was an associatiOn between bladder cancer mortality rates and trihalomethane levels. The association was observed in both sexes and was proportional in strength to the percentage of the population served by the studied water supply. The correlations noted were stronger with the brominated trihalomethanes than with chloroform. However, in its review of 13 epidemiology studies, the National Academy of Sciences Safe Drinking Water Committee concluded that a causal relationship had not been established by the studies that had been reported (29). Hogan et al. (30) used much the same data base and applied various statistical procedures in order to determine the appropriateness of the statistical model. When a weighted regression analysis was applied, the results were similar to those of previous studies showing positive correlations between rectal-intestinal and bladder cancer mortality rates and chloroform levels in drinking-water.
8.4.4 Guideline value Chloroform has several adverse effects on the health of humans. Safe levels of consumption for most of the effects observed directly in man would be difficult to estimate because adequate quantitative studies have not been conducted. Of the hazards that could potentially arise at concentrations approximating to those observed in drinking-water, the most serious are the carcinogenic effects observed in experimental
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245
animals and the suggestion of similar effects in humans exposed to elevated trihalomethane levels in drinking-water. Estimation of safe levels of chloroform is made by applying a multistage linear extrapolation model to the data obtained in rats in the NCI bioassay of chloroform. The rat data were used in preference to the data obtained in the B6C3-Fl mouse because there is serious doubt about the mechanism responsible for producing liver tumours in the latter model when exposed to hepatotoxic agents such as chloroform. Data obtained subsequent to the NCI bioassay clearly indicate that this animal is much more sensitive than the rat to liver damage due to chloroform (31). Such damage occurs at doses lower than those used in the NCI study, suggesting an epigenetic mechanism. Extrapolation models for such effects do not exist. Additionally, rates of chloroform metabolism in the rat approximate more closely to those observed in man than do the rates in the mouse. Given the above considerations, a guideline value of 30 .ug/litre is recommended for chloroform in drinking-water. At an average consumption of 2 litres per day, such a concentration would lead to less than I additional case of cancer in a population of 100000 in a lifetime. Additionally, it should be pointed out that the risk associated with inadequate disinfection would be much higher than that resulting from concentrations of chloroform considerably greater than the recommended value.
REFERENCES I. HARDIE, D. W. F. Chloroform. In: Kirk, R. E. & Othmer, D. T., ed. Encyclopedia of chemical technology. 2nd ed. Vol. 5. New York, lnterscience Publishers, 1964. p. 119. 2. NATIONAL INSTITUTE FOR OcCUPATIONAL SAFETY AND HEALTH. Critena for a recommended standard. Occupational exposure to chloroform. Washington, DC, US Department of Health, Education and Welfare, 1974, p. 75-114. 3. PEARSON, C. R. & McCONNELL, G. Chlorinated C 1 and C 2 hydrocarbons in the marine environment. Proceedings of the Royal Society, Senes B, 189: 305-332 (1975). 4. RooK, J. J. Formation of haloforms during chlorination of natural waters. Water treament and exammatwn, 23: 234 (I 974). 5. RooK, J. J. Haloforms in drinking water. Journal of the American Water Works Assoczatwn, 68: 186 (I 976). 6. STEVENS, A. A. ET AL. ChlorinatiOn of organics in drinking water. In: Jolley, R. L., ed., Water chlorination. Environmental impact and health effects, Ann Arbor, MI, Ann Arbor Science, 1975. 7. HEALTH & WELFARE, CANADA, National survey for halomethanes in drinking water. Ottawa, Health & Welfare, Canada, 1977 (7-EHD-9). 8. RussELL, J. W. & SHADOFF, L. A. The sampling and determination of halocarbons in ambient air using concentration on porous polymer. Journal of chromatology, 134: 375 (I 977). 9. LILLIAN, D. ET AL. Atmospheric fates of halogenated compounds. Envtronmental science and technology, 9: 1042 (1975). 10. Cox, R. A. ET AL. Photochemical oxidation of halocarbons in the troposphere. Atmospheric environment, 10: 305 (1976).
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II. McCoNNELL, G. ET AL. Chlorinated hydrocarbons and the environment. Endeavour, 34: 13 (1975). 12. LEHMANN, K. B. & HASEGAWA, D. Studies of the absorption of chlorinated hydrocarbons in animals and humans. Archiv flir Hygiene und Bakteriologie, 72: 327 (1910). 13. COHEN, E. N. & Hooo, N. Application of low-temperature autoradiOgraphy to studies of the uptake and metabolism of volatile anesthetics in the mouse. I. Chloroform. Anestheswlogy. 30: 306 (1969). 14. DowTY, B. J. ET AL. The transplacental migration and accumulatwn in blood of volatile organic constituents. Pediatric research, 10: 696 (1976). 15. FRY, R. J. ET AL. Pulmonary elimination of chloroform and its metabolite in man. Archives internationales de pharmacodynamie et de therapie, 196: 98 (1972). 16. CHIOU, W. L. Quantitatwn of hepatic and pulmonary first-pass effects and its implications in pharmacokinetic study. I. Pharmacokinetics of chloroform in man. Journal of pharmacokinetics and biopharmaceutics, 3: 193 (1975). 17. VAN DYKE, R. A. ET AL. A metabolism of volatile anesthestics. I. Conversion in vivo of several anesthetics to 14C0 2 and chloride. Biochemical pharmacology, 13: 1239 (1964). 18. WHIPPLE, G. H. & SPERRY, J. A. Chloroform poisoning-liver necrosis and repair. Bulletin of the Johns Hopkins University, 20: 278 (1909). 19. SECHER, 0. Physical and chemical data on anaesthetics. Acta anaesthesiologica scandinavica, 42 (Suppl.): I (1971). 20. GossELIN, R. E. ET AL. Clinical toxicology of commercial products. 4th ed. Baltimore, MD, The Williams and Wilkins Co., 1976. 21. DREISBACH, R. H. Handbook of pOisoning. Los Altos, CA, Lange Medical Publications, 1974. p. 275. 22. MIKLASHEVSKII, V. E. ET AL. Toxicity of chloroform administered perorally. Gigiena i sanitari_ja, 31: 320 (1966). 23. DE SALVA, S. ET AL. Long-term safety studies of a chloroform containing dentifrice and mouth-rinse in man. Food and cosmetics toxicology, 13: 529 (1975). 24. WALLACE, C. J. Hepatitis and nephrosis due to cough syrup containing chloroform. California medicine, 73: 442 (1950). 25. KUTOB, S. D. & PLAA, G. L. A procedure for estimating the hepatotoxic potential of certain industrial solvents. Toxicology and applied pharmacology, 4: 354 (1962). 26. CARDEIHAC, P. T. & NAIR, K. P. C. Inhibition by castration of aflatoxin induced hepatoma in carbon tetrachlonde-treated rat~. Toxicology and applied pharmacology, 26: 393 ( 1973). 27. NATIONAL CANCER INSTITUTE. Carcinogenesis bioassay of chloroform. Bethesda, MD, National Cancer Institute, 1976. 28. CANTOR, K. P. AssociatiOn of cancer mortality rates and trihalomethane level in municipal drinking water supplies. (Abstract) American journal of epidemw/ogy, 106: 230 (1977). 29. US ENVIRONMENTAL PROTECTION AGENCY. Federal register, 68698-68703 (1979). 30. HOGAN, M. D. ET AL. Association between chloroform levels in finished drinking water supplies and various site-specific cancer mortality rates. Journal of enVIronmental pathology and toxicology, 2: 873 (1979). 31. BuLL, R. J. ET AL. In depth biochemical, pharmacological and metabolic studies of trihalomethanes in water. Proceedings of the NCI/EPA/NIOSH Workshop on Environmental and Occupational Carcinogenesis, 1980 (in press).
PART V. AESTHETIC CONSTITUENTS AND CHARACTERISTICS
1. ALUMINIUM 1.1 General description
Aluminium compounds are abundant in nature and are often found in water. The salts of aluminium are used extensively in water treatment for the removal of colour and turbidity. Compared with the aluminium intake from food, that from water is small. Ingested aluminium salts do not appear to exert any deleterious effects on man. The incidence of discoloration in drinking-water in distribution systems increases if the aluminium level exceeds 0.1 mgflitre in the final water. A guideline value of 0.2 mg/litre in drinking-water is therefore recommended, based on aesthetic considerations. This value represents a compromise, taking into consideration that. although some discoloration may occur at this level, lower levels may be hard to achieve in certain cases where aluminium compounds are used in water treatment. In those cases, special attention should be paid to the maintenance of the distribution system. 1.2 Occurrence
Aluminium is distributed widely in nature and is a constituent of all soils, plants, and animal tissues (1-4). As a consequence of this wide natural distribution and the activities of man, aluminium is present in air, food, and water, both natural and polluted (2, 5). 1.2.1 Water
Industrial wastes, erosion, leaching of minerals and soils, contamination from atmospheric dust, and precipitation are the main pathways by which aluminium enters the aquatic environment. The level of aluminium in water varies considerably and may exceed 10 mg/litre in the vicinity of aluminium-processing plants (6). The concentration in any particular water is controlled by pH, the type and concentration of complexing agents that may be present, the oxidation state of the mineral components, and the redox potential of the system; many acidic waters contain naturally high levels of aluminium, presumably through the leaching process. In water-treatment processes, coagulation with aluminium salts, such as alums or sodium aluminate, is used extensively for removal of finely divided mineral and organic material, especially if the water is coloured or turbid. The level of aluminium in drinking-water is, therefore, 249
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affected by the use of these salts in water-treatment processes. Most of the aluminium used as coagulant is removed as insoluble aluminium salts, which settle or are removed by filtration. Careful operation of these processes is essential for the successful treatment of water, to ensure both an aesthetically pleasing product and effective disinfection (7). Although in the treatment of water the aim is to leave as little aluminium as possible in the water, some always remains. Levels in treated water have been reported to be between < 0.01 and 2 mgjlitre (2). It is worth pointing out that these values represent an "alum" dose twelve times greater. Levels of aluminium in the final water above about 0.3 mgjlitre usually reflect faults in the coagulation, sedimentation, or filtration stages of treatment. During distribution, a portion of the aluminium may sediment out and a gradual reduction throughout the length of the supply system may be observed (8). This aluminium will accumulate in the system, especially where flows are low, and, together with iron, manganese, silica, organic material, and microorganisms, form sediments that may readily be disturbed by changes in flow and appear at the consumer tap, rendering the water aesthetically unacceptable (9). In the presence of aluminium, levels of iron normally too low to cause problems may produce obvious discoloration of water. It has been shown (10) that the incidence of discoloration in water in distribution systems, and therefore the frequency of consumer complaints, increases if the aluminium level exceeds 0.1 mgjlitre in the final water.
1.3 Routes of exposure Aluminium present in drinking-water contributes only a small proportion of the estimated daily human intake. The bulk of this intake is derived from food and, based on published information (1 I~ 14), a total aluminium consumption of 88 mg per person per day has been estimated. A consumption of 2 Iitres of water daily containing 1.5 mg of aluminium per litre will provide only 3.0 mg of aluminium per person per day, i.e., less than 4% of the normal daily intake.
1.4 Health aspects Aluminium does not appear to be an essential nutrient in man. Aluminium salts are not normally absorbed from food and water, but are complexed with phosphate and excreted in the faeces (15). The chronic use of large quantities of aluminium hydroxide in the form of "antacids" can lead to excessive loss of phosphate from the system. Ingested aluminium does not accumulate substantially in the tissues, except in bone (16), whereas aluminium compounds inhaled, in the form of dust, accumulate in the lungs (17) and lymph nodes (18). Administration of aluminium to rats at the rate of 2.5 mgjkg of body weight per day for six months led to minimal systemic toxicity and
1.
ALUMINIUM
251
minimal gonadotoxic effects (19). There was no effect on the median life-span, longevity, incidence of tumours, or clinical biochemistry of rats fed aluminium in drinking-water at a concentration of 5 mg/litre (20). There was no evidence of carcinogenesis in a series of studies using different mammals and a range of aluminium salts (21). Aluminium has been associated with certain neurological disorders, such as dialysis dementia and Alzheimer's disease (20, 21). It is not clear, however, whether the presence of aluminium causes such conditions or is simply an indicator of other factors.
REFERENCES I. CorroN, F. A. & WILKINSON, G. Advanced inorgamc chemistry, 3rd ed. New York, Wiley-Interscience, 1972, pp. 261-262. 2. SoRENSON, J. R. J. ET AL. Aluminium in the environment and human health. Environmental health perspectwes, 8: 3 (1974). 3. UNDERWOOD, E. J. Trace elements in human and animal nutritwn, 3rd ed. New York, Academic Press Inc., 1971. 4. SAAKASHVILI, T. G. & KVIRIKAZE, N. A. Content of certain trace elements in human blood. Trudy Instituta Uro/., Akademia Nauk Gruzinskoi SSR, 1: 93 (1962); Chemical abstracts, 61: 4784 (1964). 5. MONIER-WILLIAMS, G. W. A/umimum in food. London, Ministry of Health, 1935 (Reports on Public Health and Medical Subjects). 6. SYLVESTER, R. 0. ET AL. Factors involved in the locatiOn and operation of an aluminium reduction plant. Proceedings of the 22nd Industrial Waste Conference. Lafayette, IN, Purdue University, 1967, pp. 441-454. 7. RIDGWAY, J. ET AL. Water quality changes-chemical and microbiological studies. In: Water distrzbution systems. Medmenham, England, Water Research Centre, 1979. 8. AINSWORTH, R. G. ET AL. Deposits, corrosion products and corrosion mechanisms in iron mains. In: Water distribution systems. Medmenham, England, Water Research Centre, 1979. 9. AINSWORTH, R. G. ET AL. The introduction of new water mto old distribution systems. Medmenham, England, Water Research Centre, 1980 (TR 146). 10. VozAR, L. Content of aluminium in the diet and its biological action. Voprosy pitanija, 21: 28 (1962). II. ZooK, E. G. & LEHMANN, J. Total diet study: content of ten minerals-aluminium, calcium, phosphorus, sodium, potassium, boron, copper, iron, manganese and magnesium. Journal of the AssociatiOn of Official Agricultural Chemists, 48: 850 (1965). 12. GABOVICH, R. D. Contents of some trace elements in the food in certain cities and towns of the USSR. Gigiena i sanitarija, 31: 41 (1966). 13. JAULNES, P. & HAMELLA, G. Presence et taux des oligo-elements dans les aliments et les boissons de l'homme. Annates de Ia nutrztion et del' aliment, 25: Bl33 (1971). 14. HAMILTON, E. I. & MINSKI, M. J. Abundance of the chemical elements m man's diet and possible relations with environmental factors. Sc1ence of the total environment, 1: 375 (1973). 15 THIENES, C. H. & HALEY, T. J. Clinical toxicology. Philadelphia, Lea & Febiger, 1972, pp. 169-170. 16. DEICHMANN, W. B. & GERARDE, H. W. Toxicology of drugs and chemicals. New York, Academic Press, 1969, p. 88. 17. HAMILTON, E. I. ET AL. Concentration and distribution of some stable elements in healthy human tissues from the United Kingdom, environmental study. Science and the total environment, 1: 341 (1973). 18. KRASOVSKII, G. N. ET AL. Experimental study of biological effects of lead and
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aluminium following oral administration. Environmental health perspectwes. 30: 47-51 (1979). 19. MAHURKAR, S. D. ET AL. Electroencephalographic and radionuclide studies in dialysis dementia. Kidney internatwnal, 13: 306 (1978). 20. ELLIOTT, H. L. ET AL. Alummium toxicity dunng regular haemodialysis. British medical journal, 1: 1101 (1978). 21. CRAPPER, D. R. ET AL. Brain alummium distribution in Alzheimer's disease and experimental neurofibrillary degeneration. Science (Washington), 180: 511 (1973).
2. CHLORIDE 2.1 General description Chloride is widely distributed in nature, generally in the form of sodium (NaCI), potassium (KCI) and calcium (CaCI 2 ) salts. It constitutes approximately 0.05% of the lithosphere (1). By far the greatest amount of chloride in the environment is present in the oceans. The presence of chloride in natural waters can be attributed to dissolution of salt deposits (2), contamination resulting from salting of roads to control ice and snow (3-7), discharges of effluents from chemical industries (8), oil-well operations (9), sewage discharges (10), irrigation drainage (1 1), contamination from refuse leachates (12), and seawater intrusion in coastal areas (1). Each of these sources may result in local contamination of both surface-water and groundwater. The chloride ion is highly mobile, however, and is eventually transported into closed basins or to the oceans (1). 2.2 Occurrence Chloride is generally present at low concentrations in natural surfacewater. Levels in unpolluted water are often less than 10 mg/litre and may often be less than I mgflitre (1 1, 13). In foods of plant and animal origin, chloride occurs naturally, generally at levels less than 0.36 mgfg (14). The addition of salt during processing or cooking, and at the table, can markedly increase the chloride level in food. 2.3 Routes of exposure Estimation of the daily intake of chloride in food is complicated by the widespread use of table-salt as a condiment. Approximately 600 mg of chloride per day are ingested in a salt-free diet (15, 16). Because of the addition of salt to food, however, the daily intake of chloride averages 6 g and may range as high as 12 g (17. 18). An average intake of chloride from drinking-water is approximately 100 mg per day (10, 17, 18). The intake from air is negligible. 2.4 Health aspects Chloride is the most abundant anion in the human body and contributes significantly, along with its associated cations, to the osmotic 253
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activity of the extracellular fluid; 88% of the chloride in the body is extracellular (19). A normal 70-kg human body contains approximately 81.7 g of chloride (19) and 45 litres of water. Water and electrolyte balance in the body is maintained by adjusting total dietary intake and excretion via the kidneys and intestinal tract. Absorption of chloride is almost complete in normal individuals. Most fluid and electrolyte absorption takes place in the proximal half of the small intestine (20). Normal daily loss of fluid is the equivalent of about 1.5-2 litres of water, together with about 4 g of chloride; 90-95% of the chloride loss occurs in the urine, 4-8% in the faeces, and about 2% in sweat. The total obligatory loss of chloride per day amounts to approximately 530 mg (19). On the basis of this estimate of obligatory loss, a daily dietary intake for adults of 9 mg of chloride per kg of body weight (630 mg for a 70-kg man) is essential (equivalent to slightly more than 1 g of table-salt per person per day). For children up to 18 years of age, a daily dietary intake of 45 mg of chloride per kg of body weight should be sufficient (19). The taste threshold for chloride in drinking-water is dependent upon the associated cation, but is usually within the range 200-300 mg of chloride per litre. Taste threshold levels for sodium chloride, potassium chloride and calcium chloride in water are 210, 310, and 222 mg/litre, respectively (21, 22). The taste of coffee particularly is affected if it is brewed with water having a chloride concentration of 400 mg/litre as sodium chloride or 530 mg/litre as calcium chloride (23). Conventional water-treatment processes do not remove the chloride ion from water and although the amount of chloride ingested daily from drinking-water is but a very small proportion of the total daily intake, a guideline value of 250 mg of chloride per litre is recommended, based on organoleptic considerations.
REFERENCES I. NATIONAL RESEARCH COUNCIL OF CANADA AsSOCIATE CoMMITTEE ON SciENTIFIC CRITERIA FOR ENVIRONMENTAL QUALITY. Effects of alkah hahdes in the Canadian environment. Ottawa, National Research Council, 1977. 2. NATIONAL RESEARCH CouNCIL. Nutrient and toxic substances in water for livestock and poultry. Washington, DC, National Academy of Sciences, 1974. 3. MURRAY, D. M. & ENNST, V. F. W. An economic analysis of the environmental impact of highway de-icing salts. Natwnal technical information service publication, 253: 268 (1976). 4. PoLLOCK, J. J. & TOLER, L. G. Effects of h1ghway de-icing salts on groundwater and water supplies in Massachusetts. Washington, DC, Department of the Interior, 1972 (US Geological Survey). 5. TERRY, R. C. Road salt, drmking water and safety. Cambridge, MA. Ballinger, 1974. 6. HuTCHINSON, F. E. Effects of highway salting on the concentration of sodium chloride in private water supplies. Research in life sciences, 15 (1969). 7. RALSTON, J. G. De-icing salts as a source of water pollutwn, Toronto, Mmistry of the Environment, 1971.
2.
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255
8. LITTLE, A. D. Inorganic chemtcal pollution of freshwater. Washington, DC. US Envuonmental Protection Agency, 1971. 9 PETTYJOHN, W. A. Water pollutiOn by oil-field brines and related industrial wastes m Ohio. Ohio journal of science, 71: 257 (1971). 10 PETTYJOHN, W A. Water quality in a stressed em·ironment, Minnesota, Burgess Publishing Co., 1972. II. BoND, R. G. & STRAUB, C. P. Handbook of envzronmental control, Vol. 3, Cleveland, Chemical Rubber Co., 1973. 12. ScHNEIDER, W. 1 Hydrologic zmplicatwns of solid-waste disposal. Washinghton, DC. Department of the Interior. 1970, pp Fl-FIO (US Geological Survey Circular 601-F). 13. National water quality data bank. Ottawa, Inland Waters Directorate, Water Quality Branch, 1976. 14. LONG, C ET AL. Biochemists' handbook. London, E. and F. N. Spon Ltd, 1961. 15. DAHL, L. K. Salt and hypertension. Amencan journal of clinical nutrition, 25: 231 (1972) 16. MENEELY, G. R. A review of sources of and the toxic effects of excess sodium chloride and the protective effect of extra potassium in the diet. Plant foods for human nutrition, 23: 3 (1973). 17. ZoETEMAN, B. C 1. & BRINKMAN, F. 1. 1. Human intake of minerals from dnnking water m I he European Communities. In: Hardness and drinking water and public health. Proceedings of the European Scientific Colloquium, Luxembourg. Oxford, Pergamon Press, 1976, p. 175. 18. Sodzum. chlorides and conductwily in dnnkmg water. Report on a WHO Working Group. Cophenhagen, WHO Regional Office for Europe, 1979 (EURO Reports and Studies, No. 2). 19 Dietary standard for Canada. Ottawa, Health Protection Branch, Department of National Health and Welfare, 1975. 20. SLADEN, C. E. Absorption of flUid and electrolytes in health and disease. In: McColl, I. & Sladen, G. E., ed. lntestme absorptwn m man. London, AcademiC Press, 1975, p. 51. 21. WHIPPLE, G. C. The value of pure water. New York, John Wiley and Sons, 1907. 22. RICHTER, C. P. & MACLEAN, A. Salt taste threshold of humans. Amencan journal of physiology, 126: I (1939). 23. LocKHARD, E. E. ET AL The effect of water impunties on the flavor of brewed coffee. Food research, 20: 598 (1955).
3. COLOUR 3.1. General description Colour in drinking-water may be due to the presence of: coloured organic substances, usually humics; metals such as iron and manganese; or highly coloured industrial wastes, of which pulp and paper and textile wastes are the most common. The primary importance of colour in drinking-water is aesthetic but the sensory effects may be regarded as a health effect. Experience has shown that consumers whose drinking-water contains aesthetically displeasing levels of colour may seek alternative, possibly unsafe, sources. Most people can detect levels of colour above 15 TCU (true colour units) in a glass of water. The removal of excess colour. prior to chlorination, will reduce the production of trihalomethanes. Taste due to chlorinated organics is also mitigated. Limiting the colour in potable water also limits the concentration of undesirable substances that are complexed with or adsorbed on to humic material. The guideline value recommended for colour in drinking-water is less than 15 TCU.
3.1.1 Source The appearance of colour in drinking-water is caused by the absorption of certain wavelengths of normal "white" light, by the presence of coloured substances, and by light scattering caused by suspended particles (1, 2). Colour measured in water that contains suspended matter is defined as "apparent colour"; "true colour" is measured on water samples from which particulate matter has been removed by centrifugation or filtration, colour then being due to humic substances in true solution (3, 4). In general, the true colour of a given water sample is substantially less than its apparent colour (4).
3.1.2 Measurement The colour of a water sample may be measured by visual comparison with a series of standard solutions containing known amounts of potassium chloroplatinate and added cobalt(II) chloride. Since the platinum-cobalt standard method was designed to analyse naturally coloured water, difficulty in comparing the colour of a water supply with standard colour solutions may be indicative of pollution. 256
3.
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One TCU corresponds to the amount of colour exhibited under the specified test conditions by a solution containing 1.0 mg of platinum per litre in the form of chloroplatinate ion (4}. A colour of 15 TCU can be detected in a glass of water by most consumers and 5 TCU will be apparent in large volumes of water, such as in a white bathtub; few people can detect a colour level of 3 TCU (5}. The colour of natural surface-water generally increases with increasing pH. This is commonly referred to as the "indicator effect" (1). Therefore, it is widely recommended that the pH of the sample be recorded together with the colour measurement to allow for this effect (4}.
3.2 Occurrence Complaints of coloured water generally approach in number those collectively concerned with taste and odour. Colour in natural waters is due mainly to organic matter, primarily humic substances, originating from the decay and aqueous extraction of vegetation into surface-water. Iron and manganese may often be present in groundwater as well as in some surface-waters and impart a colour. Another important source of iron in drinking-water is dissolution of iron pipes conveying the water. Iron and manganese can give rise to red and black water respectively. Copper solubilized from copper pipes may give rise to blue-green discoloration of sanitary ware in addition to a faint blue colour to the water in extreme cases. Highly coloured wastewaters, in particular wastes from the pulp, paper and textile industries, may create coloured water problems. A colour problem of microbiological origin is the production of "red water", a phenomenon caused by the oxidation of iron (II) to iron (III), as a result of which, the iron precipitates from solution as the hydroxide and imparts a characteristic reddish colour to the water. In severe cases, distribution lines have been blocked by the action of these "iron bacteria". Similarly, a black discoloration may be imparted to drinkingwater by the action of bacteria capable of oxidizing dissolved manganese to its insoluble oxides. This colour problem occurs more frequently in groundwater than in surface-water supplies.
3.2.1 Removal of colour
Removal of colour from water may sometimes best be effected in practice by chemical oxidation to supplement the coagulation and filtration. Data recorded in 1976 for about 200 plants in Sweden showed that colour levels ranging from < 5 to 150 TCU for untreated water were reduced to< 5 to 25 TCU for treated water (6}.
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3.3 Health aspects Limits for colour in potable water have traditionally been based on aesthetic considerations. It has been noted, however, that supply to consumers of visibly coloured water may lead them to seek a colourless, but possibly unsafe, alternative source of drinking-water (7). Other health-related criteria include the association between colour and production of some chlorinated organic compounds, interference with the water treatment, and increased chlorine consumption. Few toxicological studies of natural organic colour agents have been undertaken. Drinking-water containing a low-ash preparation of soil fulvic acid in concentrations of 10, 100, and 1000 mg/litre was supplied to male rats for periods of up to 90 days; no significant changes in body weight, food and water intake, organjbody weight ratios, or tissue histology were observed (G. C. Becking & A. P. Yagminas, unpublished observations, 1978). The same fulvic acid preparation was also given daily (for 14 days) to rats by gavage at a dosage of 1000 mgjkg of body weight; no mortality occurred at this dose level. The rate of weight gain, however, was decreased compared with that of control animals and slight changes were noted in some of the kidney enzyme concentrations. In the most relevant study to date, humic material was fed to rats in their drinking-water at two dose levels for periods of 19-35 weeks. The authors concluded that, applying a safety factor of approximately 100, drinking-water containing 2.5 mg of "humic acid" per litre would be safe for human consumption (8). Very few studies have been devoted to the comparative toxicities of trace metals and their humate complexes with reference to human health (9). It has been shown that the acute mammalian toxicities of iron, lead, barium, silver, copper, and zinc are substantially enhanced on intravenous injection as their humate complexes, but that orally ingested lead humate is at least 60% less toxic than lead acetate (10). Increases of 50 to 100 'i'o were found in the amounts of ionic material (calcium, magnesium, iron, manganese, zinc, and sulfate) that permeated the intestine in the presence of humic acid (1 1). Unfortunately, the substances studied did not include toxic trace metals. No information has been published to date on the bioavailability to mammals of the humate complexes of toxic organic substances.
3.4 Other considerations The adsorption by humic substances isolated from soil (1 2, 13) of organic compounds in amounts that can exceed their aqueous solubilities (14, 15) is a matter of potential importance meriting further investigation. Furthermore, owing to their polyanionic electrolytic properties, humic substances play an essential role in the dissolution, transport, and deposition of positively charged inorganic ions, such as
3. the heavy substances solubility substances (16).
COLOUR
259
metals. Most metals will complex on contact with humic in water. Complex formation can dramatically increase the of the metal; for example, naturally occurring humic in water may render iron up to 10 9 times more soluble
Some metals under certain circumstances form insoluble complexes with humic substances; this is the basis for the use of iron and aluminium salts in the production of drinking-water. The fate of complexed toxic metals in water undergoing treatment for production of drinking-water is a particularly relevant question. The most prevalent view is that, although toxic metals associated with the suspended solids in water may be at least partially removed, dissolved trace metals are probably removed only to a negligible extent during conventional treatment (17). Claims that dissolved humic substances cause a taste in drinking-water (18) cannot be confirmed as no recent research appears to have been done on the topic. Highly coloured, polluted water will frequently have an associated objectionable taste, but the degree to which this association is causative is unknown. It is known that the organic colouring material in water stimulates the growth of many aquatic microorganisms (19), some of which are directly responsible for the production of odour in water. The relationship between corrosion and incrustation and the humic content of water is both complex and important. Small amounts of humic substances (l-2 mgjlitre) assist in the deposition of a protective layer of calcium carbonate in distribution systems (20). Where lime has been added as a post-treatment corrective step for corrosive waters (2 1), larger amounts of humic acid may be responsible for the deposition of flow-restrictive "humus mud" in distribution systems. Water containing very little dissolved humic material can be more metal-corrosive than water containing larger amounts (22, 23). Since humic acid and certain of its metal complexes are poorly soluble at the pH of potable water, they may be partly responsible for turbidity in a water sample. Furthermore, since "dissolved" humic substances in water exist predominantly as colloidal dispersions, and since optical measurements of turbidity are influenced by particles in the colloidal size range (2), such colour in water will affect turbidity values. Difficulty in maintaining a free available chlorine residual in distribution systems may be due to the presence of organic colour in treated water. Although this fact has been known since 1949 (18), it was not until the discovery in 1974 (24, 25) of the presence of relatively large amounts of chloroform and other trihalomethanes in chlorinated water that the reaction of chlorine with dissolved humic substances was subjected to careful study. Trihalomethanes are reaction products of chlorine (and adventitiously present bromine and iodine) with humic substances; fortunately, however, methods used for coagulation remove
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AESTHETIC CONSTITUENTS AND CHARACTERISTICS
most of the organic precursors from the raw water (26). Colour can interfere with the chemical analysis of many constituents of water. In Standard methods for the examination of water and wastewater, for example, it is stated that colour must be compensated for or removed in colorimetric analytical methods (4). The metal-complexing properties of humic substances can interfere in non-colorimetric methods of analysis. Humic substances interfere with trace metal analysis when a complexing agent is employed and the sample is concentrated by extraction with an organic solvent (27).
REFERENCES I. BLACK, A. P. & CHRISTMAN, R. F. Characteristics of colored surface waters. Journal of the Amencan Water Works Association, 55: 753 (1963). 2. BLACK. A. P. & HANNAH, S. A. Measurement of low turbidities. Journal of the American Water Works Assocwtion, 57: 901 (1965). 3. SAWYER, C. N. & MCCARTY, P. L. Chemistry for sanitary engineers, 2nd ed. Toronto, McGraw-Hill Book Company, 1967, p. 299. 4. Standard methods for the examination of water and wastewater, 14th ed. Washington, DC, APHA, AWWA, WPCF, 1976. 5. BEAN, E. L. Progress report on water quality critena. Journal of the Amencan Water Works Assocwtion, 54: 1313 (1962). 6. ANON. VAV AD 76 Vattenbeskaffenhet1976 [Water quality 1976]. Stockholm, Swedish Water Works Association, 1977. 7. Public health service drinking water standards. Rockville, MD, US Department of Health, Education and Welfare, 1962, p. 21 (Public Health Service Publication No. 956). 8. JANECEK, J. & CHALUPA, J. Biological effects of peat water humic acids on warmblooded orgamsms. Archw f!ir hydrobiologie, 65: 515 (1969). 9. BROWN, V. M. ET AL. Aspects of water quality and the toxicity of copper to rambow trout. Water research, 8: 797 (1974). 10 KLOCKING, R. Influence of humic actds on the toxicity of lead. Proceedmys of the European society of toxicology, 16: 258 (1975). II. VISSER, S. A. Some biological effects of humic acid in the rat. Acta bwlogica et medzca Germamca, 31: 569 (1973). 12. FLAIG, W. ET AL. In: Gieseking, J. E., ed. Soil components. Vol. I, New York, Springer-Verlag, 1975. 13. FELBECK, G. T. JR. In: McLaren A. D. & Skujms, J .. ed. Sot! bwchemzstry. Vol 2. New York, Marcel Dekker, 1971, pp. 54-56. 14. KHAN, S. U. & ScHNITZER, M. The retentiOn of hydrophobic orgamc compounds by humic acid. Geochzmica cosmochimica acta, 36: 745 (1972). 15. HAGUE, R. & ScHMEDDING, D. Studies on the adsorption of selected polychlorinated biphenyl isomers on several surfaces. Journal of environmental science and health, Bll: 129 (1976) 16. SHAPIRO. J. Effect of yellow organic acids on iron and other metals m water. Journal of the Amencan Water Works Assocwtion, 56: 1062 (1964). 17. COMMITTEE ON WATER QUALITY CRITERIA, NATIONAL ACADEMY OF SCIENCES. Water quality cmena 1972. Washington. DC, US Government Pnnting Office, 1973, p. 51 (EPA-R-73-033). 18. McKEE, J. E. & WOLF, H. W., ed. Water quality cntena, 2nd ed. Sacramento, CA, California State Water Quality Control Board, 1963, p. 198 (PublicatiOn No. 3-A). 19. PRAKASH, A. ET AL. Humic substances and aquatic productivity. In· Povoledo, D. &
3.
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20. 21. 22.
23. 24. 25. 26.
27.
Golterman, H. L., ed. Humic substances 1972. Wageningen, The Netherlands, Pudoc, 1975, pp. 259-268. AMERICAN WATER WORKS ASSOCIATION. Water quality and treatment, 3rd ed. Toronto, McGraw-Hill Book Company, 1971, p. 311. GJESSING, E. T. Phys1ca/ and chemical characteristics of aquatic humus. Ann Arbor, Ml, Ann Arbor Science, 1976. VAN BENEDEN, G. & LECLERC, E. Les matieres humiques. Leur comportement dans les eaux ou general, leur role dans Ia corrosion des metaux. Techno/. water (Czech), 8: 225 (1964). MooRE, M. R. Plumbosolvency of waters. Nature, 243: 223 (1973). RooK, J. J. Formation of haloforms during chlorination of natural waters. Journal of water treatment and examination, 23: 234 (1974). BELLAR, T. A. ET AL. The occurrence of organohalides in chlorinated drinking waters. Journal of the American Water Works Association, 66: 703 (1974). STEVENS, A. A. & SYMONS, J. M. Measurement of trihalomethane and precursor concentration changes. Journal of the American Water Works Association, 69: 546 (1977). PAKALNS, P. & FARRAR, Y. J. The effect of surfactants on the extraction-atomic absorption, spectrophotometric determination of copper, iron, manganese, lead, nickel, zinc, cadmium and cobalt. Water research, 11: 145 (1977).
4. COPPER 4.1 General description
Copper and its compounds are ubiquitous in the environment and are thus frequently found in surface-water. The nature of the copper in water depends on the pH and carbonate concentration in the water and the other anions in solution (1). Water treatment processes usually result in the removal of trace metals from water but the copper concentration in drinking-water at the consumer's tap can be higher than in either the source water or the treated water entering the supply. Various chemical and physical characteristics of the distributed water influence the leaching of copper from the distribution system and household plumbing. Water stored in copper vessels tends to maintain the bacterial quality without deterioration (2). Copper in solution imparts a colour and an undesirable taste to drinking-water (3). 4.2 Occurrence
The copper content of soils depends on such factors as geographical location, proximity to industry, and use of fertilizers. Copper concentrations in inorganic-based fertilizers were found to range from 0.01 to 0.05 mgjg (4). The amount of copper present in food will vary with the copper content of the soil from which it derives. Foods such as vegetables, flour, and dairy and meat products normally have a copper content less than 0.01 mgjg.a Copper levels in drinking-water vary normally from 0.01 to 0.5 mgjlitre (5). 4.3 Health aspects
Copper is an essential element in human metabolism, having roles in erythrocyte formation, release of tissue iron, and the development of bone, the central nervous system, and connective tissue. Copper is usually combined with proteins: haemocuprein in the erythrocytes and ceruloplasmin in the blood plasma contain copper as an integral part of their structure; metallothionein is a copper storage protein. A number of copper-containing enzymes have been isolated, notably cytochrome oxidase, ascorbic acid oxidase, and uricase. a Department of NatiOnal Health and Welfare. Food Momtonng Survey, ProJect FMOI, Ottawa, Canada. August 1971-January 1976 (unpublished data).
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As copper is widely distributed in foods, it is unlikely that human beings, with the exception, perhaps, of infants on an exclusive milk diet, ever develop a dietary deficiency of copper. It is a beneficial adjunct to iron therapy in the treatment of nutritional anaemia in infants. In animals, however, a wide variety of clinical disorders have been associated with copper deficiency (6). Intake of excessively large doses by man leads to severe mucosal irritation and corrosion, widespread capillary damage, hepatic and renal damage, and central nervous system irritation followed by depression. Severe gastrointestinal irritation and possible necrotic changes in the liver and kidneys could occur. However, copper poisoning is rare in man and higher mammals owing to the powerful emetic action of copper. Application of copper salts to the skin is corrosive and may lead to papulovesicular eczema. Local action on the eye produces serious inflammation. Copper in water has an unpleasant, astringent taste. The taste threshold is above 5.0 mg/litre, although taste is detectable in distilled water at 2.6 mgflitre (7). 4.4 Other aspects The presence of copper in the water supply, although not constituting a hazard to health, may interfere with the intended domestic uses of water. Copper in public water supplies enhances corrosion of aluminium and zinc utensils and fittings. Staining of laundry and plumbing fixtures occurs when copper eoncentrations in water exceed 1.0 mg/litre, and this value is recommended as a guideline value.
REFERENCES I. McKEE, J. E. & WoLF, H. W. Water quality criteria. Sacramento, CA, California State Water Quality Control Board, 1963. 2. DHABADGAONKAR, S. M. Metalhc copper for disinfection of water m rural areas. Journal of Indian Water Works Association, 12: 43 (1980). 3. PAGE, G. G. ContaminatiOn of drmking water by corrosion of copper tubes. New Zealand journal of sciences, 16: 349 (1973). 4. VAN LooN, J. C. & LICHWA, J. A study of the atomic absorption determination of some Important heavy metals in fertilizers and domestiC plant sludges. Em•ironment letters, 4: I (1973). 5. ZOETEMAN, B. C. J. & BRINKMAN, F. J 1. In: Hardness of drinking water and public health. Proceedings of the European Scientific Colloquium, Luxembourg, 1975. Oxford, Pergamon Press, 1976, p. 173. 6. WHO Technical Report Series, No. 532, 1973 (Trace elements in human nutntwn: report of a WHO Expert Comm1ttee). 7. CoHEN, J. M. ET AL. Taste threshold concentrations of metals in drinkmg water. Journal of the American Water Works Assoczatwn, 52: 660 (1960).
5. HARDNESSa 5.1 General description
Water hardness is the traditional measure of the capacity of water to react with soap, hard water requmng a considerable amount of soap to produce a lather. Scaling of hot-water pipes, boilers, and other household appliances is due to hard water. Water hardness is caused by dissolved polyvalent metallic ions. In freshwater, the principal hardness-causing ions are calcium and magnesium; the ions strontium, iron, barium, and manganese also contribute (1). Hardness is usually measured by the reaction of the polyvalent metallic ions present in a water sample with a chelating agent such as EDT A and is expressed as an equivalent concentration of calcium carbonate (1, 2). Hardness may also be estimated by determination of the individual concentrations of the components of hardness, their sum being expressed in terms of an equivalent quantity of calcium carbonate. The degree of hardness of drinking-water has been classified in terms of its equivalent CaC0 3 concentration as follows: soft medium hard hard very hard 0-60 mg/litre 60-120 mg/litre 120-180 mg/litre 180 mg/litre and above.
Hardness has also been classified in terms of equivalent concentration of CaO or Ca(OH) 2 • In the Sl system, it is recommended that hardness be expressed as moles of Ca2+ per cubic metre (3). Although hardness is caused by cations, it may also be discussed in terms of carbonate (temporary) and non-carbonate (permanent) hardness (4). Carbonate hardness refers to the amount of carbonates and bicarbonates in solution that can be removed or precipitated by boiling. This type of hardness is responsible for the deposition of scale in hotwater pipes and kettles. Non-carbonate hardness is caused by the association of the hardness-causing cations with sulfate, chloride, or nitrate and is referred to as "permanent hardness" because it cannot be removed by boiling. Alkalinity, an index of the buffering capacity of water, is closely linked to hardness. For the most part, alkalinity is produced by anions or molecular species of weak acids, mainly hydroxide, bicarbonate, and a The problems associated with hardness m dnnkmg·water have also been discussed m Part lll, sectiOn 9, p 106
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carbonate; other species, such as borates, phosphates, silicates, and organic acids also contribute to a small degree if present. Whichever solute species contributes to the alkalinity of water, it is always expressed in terms of an equivalent quantity of calcium carbonate. When the alkalinity of a surface-water is due to the presence of carbonates and/or bicarbonates, the alkalinity value is usually close to the hardness value (5).
5.2 Occurrence The principal natural sources of hardness in water are sedimentary rocks, seepage, and runoff from soils. Hard water normally originates in areas with thick topsoil and limestone formations (4). Groundwater is generally harder than surface-water. Groundwater rich in carbonic acid and dissolved oxygen usually possesses a high solubilizing potential towards soil or rocks that contain appreciable amounts of the minerals calcite, gypsum, and dolomite, and consequently hardness levels up to several thousand milligrams per litre can result (4, 6). The two main industrial sources of hardness are the inorganic chemical and the mining industries (4, 7). In the building industry calcium oxide is used in mortar, stucco, and plaster. It also finds use in pulp and paper production, sugar refining, petroleum refining, tanning, and as a water and wastewater treatment chemical (8). Magnesium is also used in various processes in the textile, tanning, and paper industries. Alloys of magnesium find extensive use in moulds and die castings, portable tools, luggage, and general household goods. The salts of magnesium are used in the production of magnesium metal, fertilizers, ceramics, explosives, and medicines (9).
5.3 Health aspects As outlined in section 5.1, the major contributing factors to hardness of water are calcium and magnesium ions. There is no evidence of adverse health effects specifically attributable to high levels of calcium or magnesium in drinking-water. Apart from the domestic disadvantage resulting from the use of water possessing a high degree of hardness, another possible disadvantage may arise from the association of magnesium with the sulfate ion resulting in a water possessing laxative properties. The taste threshold for the calcium ion in drinking-water varies from I 00 to 300 mgjlitre, depending upon the anions present; for the magnesium ion the taste threshold is less than this value (10). Part Ill, which deals with the health aspects of inorganic constituents of drinking-water, should be consulted for further details on the relationship between water hardness and cardiovascular disease (see p. 109). Guideline values are not proposed for calcium or magnesium in water as
266
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a guideline value is proposed for total hardness based on aesthetic considerations.
5.4 Other aspects Soft water has a greater tendency to cause corrosion of pipes, and consequently, certain heavy metals such as copper, zinc, lead, and cadmium may be present in the distributed drinking-water (11-14). The degree to which this corrosion and solubilization of metals occurs is also a function of pH, alkalinity, and dissolved oxygen concentration. In some communities, corrosion is so severe that special precautions must be adopted with the supply (15). In areas with very hard water, household pipes can become choked with deposited scale (16); hard water also deposits incrustations on kitchen utensils as well as increasing soap consumption. Hard water can thus be both a nuisance and an economic burden to the consumer. Public acceptance of water hardness varies among communities; it is often related to the hardness to which the consumer has, over the years, become accustomed, and in many communities a water hardness greater than 500 mgjlitre is tolerated. A hardness level of about 100 mg of CaC0 3 per litre provides an acceptable balance between corrosion and the problems of incrustation, although, from aesthetic considerations, 500 mgjlitre is recommended as a guideline value (17).
REFERENCES I. Quahty criteria for water, Washington, DC, US Environmental ProtectiOn Agency, 1976 (EPA-440/9-76-023). 2. SEKERKA, I. & LECHNER, J. F. Simultaneous determination of total, non-carbonate and carbonate water hardnesses by direct potentiometry. Talanta, 22: 459 (1975) 3. GLOHMANN, A. Harte des Wassers. In: Amavis, R. et ai., ed. Hardness of drinkmg water and public health. Proceedings of the European Scienufic Colloquium, Luxembourg, 1975. Oxford, Pergamon Press, 1976, p. 129. 4. SAWYER, C. N. & McCARTY, P. L. Chemistry for sanitary engineers, 2nd ed. New York, McGraw-Hill, 1967 (Series in sanitary science and water resources engineering). 5. THOMAS, J. F. J. Industnal water resources of Canada. Water Survey Report No. I. Scope, Procedure and Interpretation of Survey Studies, Ottawa, Queen's Printer, 1953. 6. DE FULVIO, S. & 0LORI, L. Definitions and classificatiOn of naturally soft and naturally hard waters. Chemical and physical characteristics of the water m some member states of the European community. In: Amavis, R. et a!., ed. Hardness of drmkmg water and public health. Proceedings of the European ScJenUfic Colloqmum, Luxembourg, 1975. Oxford, Pergamon Press, 1976, p. 95. 7. BIESECKER, J. E. & GEORGE, J. R. Stream quality in Appalach1a as related to coalmine drainage, 1965. In: Pettyjohn, W. A., ed. Water quality in a stressed environment. Minnesota, Burgess Publishing Company, 1972. 8. McQuARRIE, M. C. Lime. In: McGraw-Hill encyclopedia of sczence and technolog}'. New York, McGraw-Hill, 1966. 9. BECH, A. V., ed. The technology of magnesium and its alloys. In: McGraw-Hill encyclopedw of sczence and technology, New York, McGraw-Hill, 1<)66. 10. ZOETEMAN, B. C J. Sensory assessment of water quality. Oxford, Pergamon Press, 1980.
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II. NERI, L. C. Some data from Canada. In: Amavis R., et a!., ed. Hardness of drinking water and public health. Proceedings of the European Scientific Colloquium, Luxembourg 1975. Oxford, Pergamon Press, 1976, p. 343. 12. SHARRETT, A. R. & FEINLEIB, M. Water constituents and trace elements in relation to cardiovascular diseases. Preventive medicine, 4: 20 (1975). 13. CRAUN, G. F. & McCABE, L. J. Problems associated with metals in dnnking water. Journal of the Amencan Water Works Association, 67: 593 (1975). 14. NERI, L. C. & HEWITT, D. Review and implications of ongomg and projected research outside the European communities. In: Amavis, R. et a!., ed. Hardness of drinking water and public health. Proceedings of the European Scientific Colloquium, Luxembourg, 1975. Oxford, Pergamon Press, 1976, p. 443. 15. MuLLEN, E. D. & RITTER, J. A. Potable-water corrosion control. Journal of the Amencan Water Works Association, 66: 473 (1974). 16. CoLEMAN, R. L. Potential public health aspects of trace elements and dnnking water quality. Annals of the Oklahoma Academy of Science, 5: 57 (1976). 17. BEAN, E. L. Quality goals for potable water. Journal of the Amencan Water Works Association, 60: 1317 (1968).
6. HYDROGEN SULFIDE 6.1 General description Hydrogen sulfide is a flammable, poisonous gas with a characteristic odour of rotten eggs (1). Hydrogen sulfide and sulfides of the alkali and alkaline earth metals are soluble in water (2). Soluble sulfide salts dissociate in water into sulfide ions, which react with the hydrogen ions to form hydrosulfide ion (HS-) or hydrogen sulfide (H 2 S). The relative concentrations of these species are a function of the pH of the water, hydrogen sulfide concentration increasing with decreasing pH (1, 3). The sulfide ion is present in appreciable concentrations above pH 10. In well-aerated water, hydrogen sulfide is oxidized to sulfate. Biological oxidation to elemental sulfur (3) occurs and sulfides form an indispensable link in nature in the "sulfur cycle" (4).
6.2 Occurrence Sulfide occurs naturally in mineral ores, oil, and coal deposits (5). Copper, lead, zinc, nickel, and other mined base metals may occur as simple or complex sulfides. Iron sulfides are often associated with these ores (1). Sulfides are also present in industrial wastes from petroleum and petrochemical plants, chemical plants, gas works, paper mills, heavy water plants, and tanneries (1, 3, 6, 7). Sulfides are generated by ~ulfate reducing bacteria (1, 3, 8- 10). The growth of sulfate-reducing bacteria in distribution systems can be a major cause of taste and odour problems in drinking-water. Sulfide levels in Mississippi river water were in the region of 0.092 mg/litre; at St Paul, Minnesota, pond and well-water (11) were found to contain 0.16 and 0.19 mgjlitre, respectively. Levels of atmospheric hydrogen sulfide naturally range from 1.5 x I0- 4 to 4.6 x 10- 4 mg/m 3 (6). In industrialized areas, levels are significantly higher. Sulfides are present in a number of raw and cooked foods. Chives and garlic contain both dimethyl sulfide and dimethyl trisulfide (12). The Smethylmethionine level in onion bulbs is less than 0.001 mgjg; in tomatoes it averages 0.003 mgjg, and in cabbage the concentration reaches 0.05 mg/g (13). Dimethyl sulfide is an important flavouring compound in British ales (0.0002-0.0037 mgjlitre) and European lager beers (0.003-0.114 268
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mg/litre) (14). It has also been found in eggs, milk, and dairy products (13).
The hydrogen sulfide content of cooked meat ranges in concentration from 0.276 mg/kg for minced beef to 0.394 mg/kg for minced lamb (15). The hydrogen sulfide concentration of heated dairy products ranges from 0.80 mg/litre in skimmed milk (0.1 lo fat) to 1.84 mg/litre in cream (30.5% fat) (16). Hydrogen sulfide and other soluble sulfides are utilized in pigment and dye manufacturing, tanning, pulp, and chemical processing. They are also employed in the production of cosmetics (17). Spring-waters that contain elevated concentrations of hydrogen sulfide are used for therapeutic baths and have been consumed for medicinal purposes.
6.3 Routes of exposure Data on the sulfide content of foodstuffs are incomplete and thus the daily dietary intake of sulfides has not been estimated. Exposure may result from the consumption of beer and ale, seafoods, cooked meat, warm milk, and asparagus and other vegetables. In England, the "maximum likely daily intake" of dimethyl sulfide from artificially flavoured sweets, soft drinks, creams and jellies was estimated to be I. 7 mg (17). The inhalation of 20 m 3 of air containing hydrogen sulfide in natural concentrations would result in a daily intake of 0.003-0.01 mg. The intake of sulfide from drinking-water cannot be estimated owing to the absence of data on sulfide levels in drinking-water.
6.4 Health aspects The alkali sulfides are absorbed rapidly from the intestine (18). Hydrogen sulfide is transformed to alkali sulfide in the blood and tissues. Excretion of sulfide is via the kidneys and lungs, whereas metallic sulfides are excreted by the bowels (18). Sulfide may also be oxidized to sulfate and thiosulfate and excreted by the kidneys (18). Hydrogen sulfide blocks the action of certain enzyme systems, some of which are involved directly in cellular oxidative processes. Such inhibitive effects have been reported for dehydrogenase (succinic), phosphatase (ATPase), oxidase (dopa oxidase), carbonic anhydrase, dipeptidase, benzamidase, and some iron-containing enzymes (6). The mechanism of enzyme inhibition is not completely understood, but is considered to be via the formation of metal sulfides, which decrease the availability of cations to the enzymes. Alkaline sulfides irritate the epithelium of the mucous membranes (18). Oral ingestion of alkali sulfides produces nausea, vomiting, and epigastric pain (1, 18). An oral dose of 10-15 g of sodium sulfide is fatal (18). Daily oral ingestion of dimethyl sulfide at a dose of 250 mgfkg of body weight over a period of 14 weeks was found to produce no ill
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effects in rats; this dose is equivalent to a daily intake of 17.5 g by a 70kg adult human (17). Inhaled hydrogen sulfide gas causes death by paralysing the respiratory centre (19); the threshold for acute inhalation poisoning is between 700 and 1000 mg/m 3 (length of exposure was not specified). Exposure levels of 1400-2100 mg/m 3 stimulate the nervous system and are fatal within 30 minutes; levels greater than 2800 mg/m 3 result in instantaneous death by means of nervous system paralysis. The progression of symptoms includes sudden fatigue, dizziness, intense anxiety, loss of olfactory function, collapse, respiratory arrest, and death. Olfactory paralysis prevents detection of hydrogen sulfide beyond concentrations of 225 mg/m 3 (6). Hydrogen sulfide at levels of 0.12 mgjm 3 in air causes mental depression; levels of 1.5-43 mgjm 3 result in conjunctivitis and visual disorders (6). Levels of 70-700 mg/m 3 produce chronic intoxication resulting in psychic changes, dizziness, confused somnolence, tachycardia, coughing, and vomiting (6, 19). Hydrogen sulfide in water has a disagreeable taste and odour and its presence is an important cause of consumer complaints. The taste and odour thresholds for hydrogen sulfide in solution are estimated to be in the range 0.05-0.10 mg/litre (1, 9). For sulfides, the taste and odour threshold is about 0.2 mgjlitre (1). It is unlikely that a person would consume a harmful dose of sulfide because of the unpleasant taste and odour at concentrations much lower than the toxic levels; consequently, the guideline value has been chosen as not detectable by consumers. 6.5 Other aspects Hydrogen sulfide in association with soluble iron causes black deposits in piping and on fixtures and produces black stains on laundered items.
REFERENCES I. McKEE, J. E, & WoLF, H. W. Water quality critena, 2nd ed. Sacramento, CA. California State Water Quality Control Board, 1963. pp. 156-7, 271-2, 277, 335-336. 2. SENKO, M. J. & PLANE, R. A. Chemical principles and properties, 2nd ed New York, McGraw-Hill Publishing Co., 1974, pp. 639-640. 3. Quality cnteria for water. Washington, DC, US Environmental Protection Agency, 1976, pp. 410-416. 4. SMITH, R. L. Ecology and field biology, 2nd ed. New York, Harper & Row Publishers, 1974, pp. 88-89, 123-128. 5. NATIONAL RESEARCH COUNCIL OF CANADA. Sulfur and liS morganic denvatwes in the Canadian environment. Associate Committee on Scientific Criteria for Environmental Quality. Ottawa, National Research Council, 1977. 6. BoORAS, S. G. Hydrogen sulfide· health effects and recommended air quality standard. Illinois Institute for Environmental Quality, 1974 (NTIS PB-233 843). 7. COLBY, P. J. & SMITH, L. L., JR. Survival of Walleye eggs and fry on paper fibre sludge deposits in Rainy River, Minnesota. Transactwns of the American Fisheries Society, 96: 278-296 (1967). 8. ADELMAN, I. R. & SMITH, L. L., JR. Toxicity of hydrogen sulfide to goldfish
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9. 10. II. 12. 13. 14.
15. 16. 17. 18. 19.
(Carassius Auratus) as influenced by temperature, oxygen, and bwassay techniques. Journal of the Fisheries Research Board of Canada, 29: 1309 (1972). AcREE, T. E. & SPLITTSTOESSER, D. F. Prevention of H 2 S in wine fermentation. New York food and life sciences bulletin, 5: 19 (1972). KADOTA, H. & IsHIDA, Y. Production of volatile sulfur compounds in microorganisms. Annual review of microbiology, 26: 127 (1972). BRODERIUS, S. J. & SMITH, L. L., JR. Direct determination and calculation of aqueous hydrogen sulfide. Analytzcal chemistry, 49: 424 (1977). CREWE, R. M. & Ross, F. P. Biosynthesis of alkyl sulphides by an ant. Nature, 254: 448 (1975). HATTULA, T. & GRANROTH, B. Formation of dimethyl sulphide from s-methylmethionine in onion seedlings (Alhum Cepa). Journal of the science of food and agriculture, 25: 1517 (1974). NIEFIND, H. J. & SPAETH, G. Some aspects of the formation of dimethyl sulfide through brewer's yeast and beer spoilage microorganisms. Proceedings of the Annual Meetmg of the American Society of Brewing Chemists, 33: 54 (1975). KuNSMAN, J. E. & RILEY, M. L. A comparison of hydrogen sulfide evolution from cooked lamb and other meats. Journal of food science, 40: 506 (1975). THOMAS, E. L. ET AL. Determination of hydrogen sulphide in heated milks by gas hquid _chromatographic head space analysis. Journal of dairy science, 59: 1865 (1975). BUTTERWORTH, K. R. ET AL. Short-term toxicity of dimethyl sulphide in the rat. Food and cosmetics toxicology, 13: 15 (1975). THIENES, C. H. & HALEY, T. J. Clinical toxicology, 5th ed. Philadelphta, Lea & Febiger, 1972, pp. 59-60. HENKIN, R. I. Effects of vapor phase pollutants on nervous system and sensory function. In: Finkel, A. J. & Duel, W. C., ed. Clinical implications of azr pollution research. Acton, MA. Public Science Group Inc., American Medical Association., 1974, pp. 193-212.
7 IRON 7.1 General description
Iron is the fourth most abundant element by weight in the earth's crust. In water it occurs mainly in the divalent and trivalent (ferrous and ferric) states (1). Both cast iron and steel pipes are employed for drinking-water distribution to the consumer. In the production of potable water, various salts of iron are used as coagulating agents (2). 7.2 Occurrence
Iron in surface-water is generally present in the ferric (Fe Ill) state. The concentration of iron in well aerated water is seldom high, but under reducing conditions, which may exist in some groundwater, lakes or reservoirs, and in the absence of sulfide and carbonate, high concentrations of soluble ferrous iron may be found (3). Concentrations of iron greater than I mg/litre have been reported to occur in groundwater (4). The presence of iron in natural waters can be attributed to the dissolution of rocks and minerals, acid mine drainage, landfill leachates, sewage, or iron-related industries. Iron is generally present at low concentrations in the atmosphere as a result of emissions from the iron and steel industry, thermal power plants, and incineration, but few data are available on levels of iron in the atmosphere. The iron content of foods varies considerably. Cereals (mean 0.0295 mg/g) and meat (0.0262 mg/g) appear to be the main dietary sources of this element (5), the iron concentration of most other .natural foods being less than 0.020 mgjg (6, 7). Levels may be somewhat higher in foods fortified with iron or in food cooked in iron utensils (8). Evidence suggests that the iron content of foodstuffs decreases during boiling (9).
7.3 Routes of exposure The daily intake of iron from typical diets in developed countries has been estimated to be in the range of 15 to 22mg (5, 9, 10). Concentrations of iron in drinking-water are normally less than 0.3 mgjlitre. and the intake from food is substantially higher than that from drinking-water. The contribution made by airborne iron to human exposure is negligible. 272
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7.4 Health aspects Iron is an essential element in human nutrition (7, 11). It is contained in a number of biologically significant proteins, for example, haemoglobin and cytochromes, and also in many oxidation-reduction enzymes. Estimates of the minimum daily requirement for iron vary from 7 to 14mg, depending upon age and sex; pregnant women may require m excess of 15 mg per day (12). The average daily requirement is considered to be I 0 mg. Individual iron requirements (which depend on age, sex, and physiological state) regulate the amount of iron absorbed from the diet, the amount varying from I to 20% (12, 13). In most individuals, about 10% of ingested iron is absorbed (7, 12). Obligatory losses (in faeces, urine, and perspiration) amount to I mg per day (12). Between 60 and 70% of absorbed iron is used in the production of haemoglobin; 5% is utilized in myoglobin production, the excess being stored primarily m the liver, bone marrow, and spleen. Iron ingestion in large quantities results in a condition known as haemochromatosis (normal regulatory mechanisms do not operate effectively) wherein tissue damage results from iron accumulation. This condition rarely develops from simple dietary overloading (7, 13, 14). However, the condition has resulted from prolonged consumption of acid foodstuffs cooked in iron kitchenware (14). Haemochromatosis has not been demonstrated in test animals receiving extremely high doses of iron (15). Small children have been poisoned following the ingestion of large quantities of iron tablets (16).
7.5 Other aspects The presence of iron in drinking-water supplies is objectionable for a number of reasons unrelated to health (4, 9, 17). Under the pH conditions existing in drinking-water supplies, ferrous salts are unstable and precipitate as insoluble ferric hydroxide, which settles out as a rustcoloured silt. Such water often tastes unpalatable and stains laundry and plumbing fixtures. The iron that settles out in the distribution system gradually reduces the flow of water. Iron also promotes the growth of "iron bacteria". These microorganisms derive their energy from the oxidation of ferrous iron to ferric iron, and in the process deposit a slimy coating on the piping. The above problems usually arise, especially in distribution systems, when the iron concentration approaches 0.3 mg/litre, and consequently, whenever possible, it would be prudent to maintain levels below this value. REFERENCES I. Quality cntena for water. Washington, DC. US Environmental Protectwn Agency, 1976.
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2. Cox, C. R. Operation and control of water treatment processes. Geneva, World Health Organization, 1964 (WHO Monograph Series No. 49). 3. HEM, J. D. Chemical factors that influence the availability of iron and manganese in aqueous systems. Washington, DC, The Geological Society of America Inc., 1972 (Special Paper 140) p. 17. 4. DART, F. J. The hazard of iron. Ottawa, Water and Pollution Control Canada, 1974. 5. MERANGER, J. C. & SMITH, D. C. The heavy metal content of a typical Canadian diet. Canadian ;ournal of pubhc health, 63: 53 (1972). 6. GORMICAN, A. Inorganic elements in foods used in hospital menus. Journal of the American Dietetic Associatton, 56: 397 (1976). 7. WATT, B. K. & MERRILL, A. L. Composition of foods-raw, processed, prepared. Washington, DC, Revised USDA Agriculture Handbook, Vol. 8 (1963). 8. BOWERING, J. & MACPHERSON SANCHEZ, A. A conspectus of research on iron requirements of man. Journal of nutrition, 106: 985 (1976). 9. ZoETEMAN, B. C. J. & BRINKMANN, F. J. J. Intake of minerals by man. In: Hardness of drinking water and public health. Proceedings of the European Scientific Colloquium, Luxembourg, 1975. Oxford, Pergamon Press, 1976. 10. KIRKPATRICK, D. C. & CoFFIN, D. E. The trace metal content of representative Canadian diets in 1970 and 1971. Canadian Institute of Food Science and Technology ;ournal, 7: 56 (1974). II. MooRE, C. V. Iron. In: Goodheart, R. S. & Shils, M. E., ed., Modern nutrition in health and disease. Philadelphia, Lea & Febiger, 1973, p. 297. 12. Dietary standard for Canada. Ottawa, Department of Natwnal Health and Welfare, 1975. 13. HoPPS, H. C. Ecology of disease in relation to environmental trace elements-particularly iron. Washington, DC, The Geological Society of America Inc., 1972 (Special Paper 140, p. 1). 14. JACOBS, A. Iron overload-clinical and pathological aspects. Seminars in hematology. 14: 89 (1977). 15. BROWN, E. B. ET AL. Studies in iron transportation and metabolism. Journal of laboratory and c/imcal medicine, 12: 862 (1957). 16. STEIN, M. ET AL. Acute Iron poisoning in children. Western journal of medicine, 125: 289 (1976). 17. Iron. In: McKee, J. E. & Wolf, H. W., ed., Water quality criteria. Sacramento, CA. California State Water Quality Control Board, 1971 (Publication 3-A, p. 202).
8. MANGANESE 8.1. General description Airborne manganese sources are widespread and its distribution over large distances is favoured by the fact that emitted manganese is predominantly associated with smaller dust particles {1). When present in natural surface-water, manganese occurs in both the dissolved and suspended forms. Anaerobic groundwater often contains elevated levels of dissolved manganese. 8.2 Occurrence Freshwater may contain from one to several thousand micrograms of manganese per litre depending on the location. Manganese concentrations in various lakes and rivers in Canada, the Federal Republic of Germany, the United Kingdom, the USA, and the USSR were found to range from 1 Jig/litre to around 600 Jig/litre (1, 2). Higher levels of manganese sometimes found in freely flowing river-water are usually associated with industrial pollution. The reducing conditions that may exist in underground water and in some lakes and reservoirs are conducive to very high levels of manganese. Concentrations in the ambient air of non-industrialized areas average 0.05 Jig/m 3 , whereas in industrialized areas concentrations up to 0.3 Jig/m 3 have been found. The manganese content of foodstuffs varies considerably. Low concentrations are found in dairy products (0.0-1.9 mgjkg), meats (0.00.8 mg/kg), and fish (0.0-0.1 mg/kg); higher concentrations occur in grains and cereals (1.2-30.8 mgjkg), nuts (0.4-35.1 mg/kg), and vegetables (0.2-12. 7 mg/kg) {1). Extremely high manganese concentrations have been found in the leaves of tea and a cup of tea can contain from 1.4 to 3.6 Jig. 8.3 Routes of exposure Uptake of manganese occurs via inhalation and ingestion from both food and drinking-water. The greatest exposure to manganese is from foods, and adults have been estimated to consume between 2.0 and 8.8 mgjday. Infants, during the first six months of life, have been shown to consume from 2.5 to 25 Jig/kg of body weight per day {1). The average daily intake of manganese in North America lies in 275
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the range 3.0-4.1 mg (average 3.6 mg) (3-5). The average daily manganese requirement for normal physiological function is estimated to be 3-5 mg (6). Manganese intake through drinking-water can vary considerably, normally being substantially lower than intake from food. Available data indicate that exposure via this source would normally be less than 0.1 mgjday, but can be an order of magnitude higher (7-9). The daily atmospheric exposure to manganese of a non-occupationally exposed population living in a non-industrialized area has been estimated to be in the range of 2-10 J.lg (9). A daily intake of 3-7 mg will give a body burden of 12-20 mg in a 70-kg man (11, 14). 8.4. Health aspects Manganese is an essential element in animals and man but only about 3 ~~ of ingested manganese is absorbed (10). It is required as a cofactor in a number of enzyme systems; it plays a role in the proper functioning of flavoproteins and in the synthesis of sulfated mucopolysaccharides, cholesterol, and haemoglobin, and in many other important metabolic processes (11). It has recently been suggested that its presence in drinking-water is inversely related to cardiovascular mortality ( 12). Absorbed manganese leaves the bloodstream quickly and concentrates in the liver. Manganese concentrated in the liver is conjugated with the bile salts. Manganese has a rather short biological half-life (10). The storage capacity of the liver for manganese is limited to about 1-1.3 mg/kg (wet weight). In animals, experimentally induced or naturally occurring manganese deficiency has resulted in a variety of symptoms (1). Although no specific syndrome due to manganese deficiency has been described in man, it has been suggested that there may be an association between manganese deficiency and disorders such as anaemia, bone changes in children (13), and lupus erythematosus. The main routes of absorption of manganese are the respiratory and gastrointestinal tracts (1 5). Owing to the low solubility of manganese in the gastric juice, only 3-4 lo of that orally administered is absorbed by the gastrointestinal tract (14). Intimately linked with the absorption of manganese is the absorption of iron (16). Anaemia leads to an increased absorption of both iron and manganese, and in anaemic individuals absorption of manganese is increased more than twofold. Absorption is inversely related to the level of calcium in the diet (5) and directly related to the level of potassium {17). Manganese in the body is regulated primarily by excretion rather than by both absorption and excretion. Some of the metal is excreted through the pancreatic secretions and some directly through the wall of the gut. Very little, 0.1-2%, is eliminated in the urine. Manganese is regarded as one of the least toxic elements. Chronic ingestion experiments in rabbits, pigs, and cattle at 1-2 mgjkg of body weight dose levels have shown no
8.
MANGANESE
277
effects other than a change in appetite and reduction in metabolism of iron to form haemoglobin (1 1). Except for one isolated incident, manganese intoxication from drinking-water has not been documented. In 1941, the cause of an encephalitis-like disease in Japan was attributed to contaminated wellwater, which had a manganese concentration of 14 mg/Iitre. However, the concentrations of other metals, especially zinc, were also excessive, and whether the high concentration of manganese alone was responsible for the disease was never unequivocally established (18). In another area in Japan, a manganese concentration of 0. 75 mg/litre in a drinking-water supply had no apparent adverse effect on the health of its consumers (19).
In rabbits, chronic parenteral administration of manganese produced marked degenerative changes in the seminiferous tubules, resulting in infertility (20). In contrast to these findings is the observation in mice that the administration of small amounts of manganese prevents the necrotic effect of cadmium on the testes (18). In other studies, it was noted that copper acts synergistically with manganese. There is no evidence that manganese is carcinogenic. On the contrary, several studies suggest that manganese may have an anticarcinogenic effect. A preliminary report cited by Spivey Fox (21) indicated that a higher incidence of human cancer in Finland was geographically related to soils low in easily soluble manganese. Manganese has been reported to inhibit aminoazo dye metabolism during carcinogenesis (18). No adverse health effects in humans were noted with daily manganese intake levels as follows (5): At•erage ( mg) Range (mg)
Food Water Air
3.000 0.005 0.002
2.0-7.0 0.0-1.0 0.0-0.029
8.5 Other aspects The presence of manganese in drinking-water supplies may be objectionable for a number of reasons unrt>Jated to health. At concentrations exceeding 0.15 mgjlitre manganese- imparts an undesirable taste to beverages and stains plumbing fixtures and laundry (22). When manganous compounds in solution undergo oxidation, manganese is precipitated resulting in problems of incrustation. Even at concentrations of approximately 0.02 mg/litre, manganese will form coatings on piping which may slough off as a black precipitate (23). The growth of certain nuisance organisms is also supported by manganese (22, 24); these organisms concentrate manganese and give rise to taste, odour, and turbidity problems in the distributed water. The guideline value for drinking-water is recommended as 0.1 mg/ litre, based on considerations of the staining properties of manganese.
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This is a compromise value and obviously, to avoid staining problems, concentrations of this metal in solution should be kept as low as possible.
REFERENCES I. Manganese, Geneva, World Health Orgamzation, 1981 (Environmental Health Criteria. No. 17) 2. National Water Quality Data Bank. Ottawa, Inland Waters Directorate, Water Quality Branch. Environment, Canada. 1976. 3. KIRKPATRICK, D. C. & CoFFIN, D. E. The trace metal content of representative Canadian diets in 1970 and 1971. Canadzan Inslltute of Food Science and Technology Journal, 7: 56 (1974). 4. MERANGER, J. C. & SMITH, D. C. The heavy metal content of a typical Canadian diet. Canadian journal of public health, 63: 53 (1972). 5. SCHROEDER, H. A. ET AL Essential trace metals in man: manganese. A study in homeostasis. Journal of chronic diseases, 19: 545 (1966). 6. KAY, H. 0. Micro-nutrient elements-a recapitulation. Journal of food technology, 2: 99 (1967). 7. US ENVIRONMENTAL PROTECTION AGENCY. Scientific and technical assessment report on manganese. Washington. DC, Office of Research and Development 1975 (Report No. EPA-600;6-75-002). 8. CRAUN, G. F. & McCABE, L. J. Problems associated with metals in drinking water. Journal of the American Water Works Association, 67: 593 (1975). 9. Guidelines for Canadian drinking-water quality. Quebec, Supply and Services, 1980 (supportmg documentation). 10. WHO Technical Report Series, No. 647, 1980 (Recommended health-based limits on occupational exposure to heavy metals: report of a WHO Study Group). II. NATIONAL RESEARCH CoUNCIL (Committee on Medical and Biological Effects of Environmental Pollutants). Manganese. Washington, DC, National Academy of Sciences, 1973. 12. MASIRON~ R. lnternatwnal studies on trace elements m the etiology of cardiovascular diseases. Nutrition reports international, 7: 51 (1973). 13. PIER, S. M. The role of heavy metals in human health. Texas reports on biology and medicine, 33: 85 (1975). 14. MENA, J. The role of manganese in human disease. Annals of climca/ and laboratory sc1ence, 4: 487 (1974). 15. RODIER, J. Manganese poisoning in Moroccan miners. Bntish journal of industrial medicme, 12: 21 (1955). 16. MENA, J. ET AL. Chronic manganese poisonmg. Neurology, 19: 1000 (1969). 17. UNDERWOOD, E. J. Trace elements in human and animal nutrition, 3rd ed. New York, Academic Press, 1971. 18. Environmental health criteria programme for manganese and its compounds (Japanese Report). Geneva. World Health Organization. 1974. 19. SuzuKI, Y. [Environmental contaminatiOn by manganese.] Japanese Journal of industnal health, 12: 529-533 (1970). 20. CHANDRA, S. & T ANDON, S. K. Enhanced manganese toxicity in iron-deficient rats. Environmental physiology and biochemistry, 3: 230 (1973). 21. SPIVEY Fox, M. R. In: Lee, D. H. K., ed. Metallic contaminants and human health. New York, Academic Press, 1972. 22. GRIFFIN, A. E. Significance and removal of manganese in water supplies. Journal of the Amencan Water Works A.ssociatwn, 52: 1326 (1960). 23. BEAN, E. L. Potable water-quality goals. Journal of the American Water Works Assonation, 66: 221 (1974). 24. WOLFE, R. S. Microbial concentration of iron and manganese in water with low concentrations of these elements. Journal of the American Water Works Assonation, 52: 1335 (1960).
9. OXYGEN, DISSOLVED 9.1 General description
The primary effect of dissolved oxygen in water is on oxidationreduction reactions, involving iron, manganese, copper, and compounds that contain nitrogen and sulfur. In certain distribution systems, there may be a tendency for the level of dissolved oxygen to fall with residence time. Although such changes are normally indicative of corrosion processes, it is also possible that microbial respiration of organic material, especially in sediments and deposits, within pipes may be responsible. Thus dissolved oxygen may decrease without a marked increase in the concentration of iron in the water (1). Conversely, water containing high levels of iron as a result of corrosion may show little depletion of dissolved oxygen content. 9.2 Relationship to other water-quality parameters
In the corrosion of iron, proportionately little oxygen is required, thus
2Fe = 2Fe 2 + +4e 4e + 2H 2 0 + 0 2 = 40H1 mg of oxygen per litre will produce 3.5 mg of ferrous iron per litre so that a large amount of iron corrosion may occur with little perceptible change in dissolved oxygen. When the available oxygen in water has been depleted, anaerobic corrosion processes proceed involving the activity of the sulfate-reducing bacteria that may be present, thus reducing sulfate to sulfide.
8H+
+8e+SO~- =
S2 - +4H 2 0
Frequently, depletion of the level of dissolved oxygen below about 80% saturation leads to an increased incidence of consumer complaints, especially regarding taste, odour, and discoloured water (1). Depletion of oxygen in drinking-water is often associated with other problems. Under anaerobic conditions microbial reduction of nitrate to nitrite (1) and also sulfate to sulfide occurs, often giving rise to odour problems. Anaerobic waters may not be exceptionally corrosive but the products of corrosion are often less adhesive to the pipe walls and hence more likely to cause consumer complaints of discoloration. Under anaerobic conditions, the concentration of ferrous iron in solution may be 279
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AESTHETIC CONSTITUENTS AND "CHARACTERISTICS
increased throughout the distribution system. Increasing the oxygen concentration in water as, for example, following contact with air in storage tanks and water towers, will cause insoluble ferric iron to be deposited resulting in discoloration of the water distant from the source of the problem (2). There are many disadvantages in distributing a water low in dissolved oxygen. It is recommended that water in a distribution system should always contain adequate dissolved oxygen. It is difficult to recommend a guideline value, however, as other constituents in the water influence the acceptable level.
REFERENCES I. RIDGWAY, J. ET AL. Water quality changes-chemical and microbiological studies. In:
Water distribution systems. mamtenance of water quality and p1peline mtegrity. Medmenham, England, Water Research Centre, 1979 2. HALL, E. S. & SMITH, I. G. Rusty water cured by oxygen injectiOn. Water sert•lces, 78: 941 (1974).
10. pH LEVEL J0.1 General description
The pH of a solution is the negative common logarithm of the hydrogen ion activity, aw: pH = -log1o (aw ). In a dilute solution, the hydrogen ion activity is approximately equal to the concentration of hydrogen ion. The pH of an aqueous sample is usually measured electrometrically with a glass electrode (1, 2). Temperature exerts significant effects on pH measurement (1, 2). 10.2 Factors influencing pH range
The pH of an aqueous system is a measure of the acid-base equilibrium achieved by various dissolved compounds and, in most natural waters, is controlled by the carbon dioxide-bicarbonatecarbonate equilibrium system (3). This system involves various constituent equilibria, all of which are affected by temperature. In pure water, a decrease in pH of about 0.45 occurs as the temperature is raised by 25 cc (4). In water with a buffering capacity imparted by bicarbonate, carbonate, and hydroxide ions, this temperature effect is modified (4). The pH of most raw water sources lies within the range 6.5-8.5 (5). Hydrogen ion concentration may be significantly altered during water treatment. Chlorination tends to lower the pH, whereas water softening using the excess lime/soda ash process raises the pH level.
10.3
Relationship to corrosion, incrustation, and other water-quality parameters
Corrosion in water mains and water treatment plants can be a large economic burden (6). In addition to the corrosion problem there is a loss of distribution capacity and the concomitant increase in pumping costs that results in cases of calcium carbonate deposition (7). Metals used in distribution systems, such as cast iron, steel, and copper, because of their thermodynamic instability, tend to corrode in contact with water. The deterioration of concrete, asbestos-cement, and cement-lined cast-iron pipe, all of which are commonly used in 281
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AESTHETIC CONSTITUENTS AND CHARACTERISTICS
distribution systems, may also occur. Natural waters contain gases, colloidal matter, and a variety of electrolyte and non-electrolyte material; these, together with pH, determine the extent of corrosion in a system (8) and define the "aggressivity" of the water. The presence of anions that form soluble compounds with a metal increase the "corrosiveness" of the water with respect to that metal, whereas anions that form insoluble compounds may increase the metal's passivity. The role of pH in the corrosion of metals used in water distribution has been summarized by Drane (8). Calcium carbonate deposition may control corrosion. Factors affecting this process are temperature, pH, total dissolved solids, hardness, carbon dioxide, and alkalinity. Under practical conditions it is extremely difficult and may be impossible to control the calcium carbonate/bicarbonate equilibrium. Accordingly, a number of semi-empirical and empirical relationships, using easily measured parameters, have been developed. The most widely used relationship is that developed by Langelier (4, 9). The effectiveness of corrosion protection, by altering pH and alkalinity, depends upon a judicious balance of the carbonate/bicarbonate equilibrium system. Water that is exactly in equilibrium, i.e., just stabilized, with respect to calcium carbonate, will normally be corrosive to both iron and steel owing to failure to deposit calcium carbonate. Supersaturated water, on the other hand, will form substantial scale unless suitably treated. This scale, depending upon its porosity and its power of adhesion to the metal, may or may not inhibit corrosion (7). Waters that are excessively hard do not usually create severe corrosion problems, but they are prone to excessive incrustation. Hard water softened by the limejsoda ash treatment will have a pH of the order of 10.9 and will possess scale-forming tendencies (10). Stabilization can be achieved by recarbonation, addition of carbon dioxide to a pH of 9. 710, or by the addition of 0.25-0.5 mg of sodium polyphosphate per litre (10). Recarbonation to pH 8.6, to stabilize the water against subsequent excessive calcium carbonate deposition in the distribution system, has been recommended (11). Biological slime on distribution pipe surfaces may prevent the removal of oxidation products from and the penetration of oxygen to the pipe walls, thus inhibiting corrosion. Alternatively, excessive growths generating carbon dioxide could create regions of locally low pH at the pipe surface. This could lead to localized corrosion even though the bulk water might possess favourable stability and agressivity indices (12). The growth of iron bacteria is very pH-dependent, occurring over the range 5.5-8.2, with an optimum pH of about 6.5 (13). Red-water complaints are often the result of sudden growth of iron bacteria, which produce ferric hydroxide as the metabolic end product. Under favourable conditions iron bacteria may develop so rapidly that severe blocking of water pipes can occur in a matter of weeks. pH is related, in several different ways, to almost every other water
10.
pH LEVEL
283
quality parameter as aqueous chemical equilibria invariably involve hydrogen (and hydroxyl) ions. The formation of gaseous hydrogen sulfide, yielding "bad-egg" odours in waters prone to sulfur contamination, is thermodynamically favoured at pH values less than about 7.0 (14). In the chlorination process, the objectionable pungent odour (1 5) of nitrogen trichloride tends to be f9rmed in greater concentrations at pH values less than 7.0 (16). It is claimed that at high pH levels drinking-water acquires a bitter taste (17). Colour intensity in a given water sample is increased by raising the pH (18). This effect, known as the "indicator effect", has led to the suggestion that all colour measurements for water quality control be carried out at the standard pH of 8.3 (19). The efficiency of the coagulation and flocculation process is markedly dependent on pH, and it is standard practice in water treatment to adjust pH so that optimum floc formation is achieved (20, 21). Filtration efficiency in certain instances is also sensitive to pH (22). Most microorganisms usually tolerate the pH range commonly found in water sources (13, 23, 24). The microbiological integrity of water is dependent upon the pH level, which influences the effectiveness of chlorine disinfection. The germicidal efficiency of chlorine in water is lower at high pH values; this has been attributed to the reduction in hypochlorous acid concentration with increasing pH (25-27). Within the pH range encountered in drinking-water, the effectiveness of both ozone and chlorine dioxide as alternative disinfectants is unchanged (16). Corrosion in the water supply system is a major source of metal contamination in drinking-water (28). Two of the potentially most troublesome metals are lead and cadmium. At all pH levels above 6 in pure water, lead is immune to corrosion. In the presence of carbonates and bicarbonates, lead is passive between about pH 4 and pH 12, but is subject to corrosion above pH 12 (29). The supply of drinking-water with a low alkalinity and a fairly low pH to households that had lead plumbing resulted in high levels of lead in the drinking-water (30). In pure water, cadmium appears passive between about pH 9 and pH 13.5, whereas, according to experimental data, corrosion is only significant below pH 6 (29).
10.4 Health aspects A direct relationship between human health and the pH of drinkingwater is impossible to ascertain because pH is so closely associated with other aspects of water quality. In an epidemiological study carried out on drinking-water supplies in which pH was one of the parameters considered, Taylor and co-workers (31) were unable to establish any significant correlation between the incidence of viral hepatitis A and finished water pH; however, it must be appreciated that this study was primarily directed towards the effects of pH on disinfection efficiency.
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In so far as pH affects the various processes in water treatment that contribute to the removal of viruses, bacteria, and other harmful organisms, it could be claimed that pH has an indirect effect on health. The recommended guideline value for pH is 6.5-8.5, although it is recognized that some problems could arise within a distribution system with pH levels below 7. REFERENCES I. pH value. In: Standard methods for the examination of water and waste water, 14th ed. Washington, DC, APHA, AWWA, WPCF, 1976, p. 460. 2. Standard method of test for pH of water and waste water. In: Annual book of ASTM standards, Part 31. PhJladelphia, American Soc1ety for Testing and Materials, 1976, p. 178. 3. GoLDMAN, J. C. ET AL Water research, 6: 637 (1972). 4. LANGELIER, W. F. Effect of temperature on the pH of natural waters. Journal of the American Water Works Associatwn, 38: 179 (1946). 5. WEBBER, W. J , JR. & STUMM, W. Mechanism of hydrogen ion buffenng in natural waters. Journal of the American Woter Works Assonation, SS: 1553 (1963). 6. HuDSON, H. E., JR. & GILCREAS, F. W. Health and economic aspects of water hardness and corrosiveness. Journal of the American Water Works Association, 68: 201 (1976). 7. McCLANAHAN, M. A. & MANCY, K. H. Effect of pH on quality of calcium carbonate film deposited from moderately hard and hard water. Journal of the Amencan Water Works Assoctation, 66: 49 (1974). 8. DRANE, C. W. Natural waters. In: Shreir, L. L., ed. Corroswn, 2nd ed. London, Newnes-Butterworths, Chapter 2. 9. LANGELIER, W. F. Chemical equilibria in water treatment. Journal of the American Water Works AssociatiOn, 38: 169 (1946). 10. DYE, J. F. & TUEPKER, J. L. Chemistry of the lime-soda process. In: Water quality and treatment. 3rd ed. Toronto, McGraw-Hill, 1971, p. 313. II. SAWYER, C. N. & McCARTY, P. L. Residual chlorine and chlorme demand. In: Chemistry for sanitary engmeers, 2nd ed. Toronto, McGraw-Hill, 1967, p. 363. 12. O'CONNOR. J. T. ET AL Deterioration of water quality in distribution systems. Journal of the American Water Works Association, 67: 113 (1975). 13. SHAIR, S. Iron bacteria and red water. Industrial water engineering, March-April: 16 (1975). 14. PoURBAIX, M. Atlas of electrochemical equilibria in aqueous solutions. 2nd ed. Houston, National Association of Corrosion Engineers, 1974, p. 545. 15. AMERICAN WATER WORKS ASSOCIATION RESEARCH FOUNDATION. Handbook of taste and odor control experiences in the US and Canada. Denver, CO, A WW A 1976. 16. MoRRIS, J. C. Chlorination and disinfection-state of the art. Journal of the Amencan Water Works Assoctation, 63: 769 (1971). 17. Statement of basis and purpose for the national secondary drinking water regulations. Washington, DC, US Environmental Protection Agency, 1977. 18. BLACK, A. P. & CHRISTMAN, R. F. Characteristics of coloured surface waters. Journal of the American Water Works Association, SS: 753 (1963). 19. SINGLEY, J. E. ET AL. Correction of color measurements to standard conditions. Journal of the American Water Works Association, 58: 455 (1966). 20. SAWYER, C. N. & McCARTY, P. L. Chemical coagulation of water. In: Chemistry for sanitary engmeers, 2nd ed. Toronto, McGraw-Hill. 1967, p. 341. 21. MAUDLING, J. S. & HARRIS, R. H. Effect of ionic environment and temperature on the coagulation of color-causing organic compounds with ferric sulfate. Journal of the American Water Works Association, 60: 460 (1968). 22. Committee Report. Coagulation-filtration practice as related to research. Journal of the American Water Works Association, 66: 502 (1974).
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23. DAVIS, B. D. ET AL. Microbzology, 2nd ed. New York, Harper and Row, 1973, pp. 9293. 24. RuooLFS, W. ET AL Literature review on the occurrence and survival of enteric, pathogenic, and relative organisms in soil, water, sewage, and sludges, and on vegetation. Sewage and industrial wastes, 22: 1261 (1950). 25. BUTTERFIELD, C. T. ET AL. Influence of pH and temperature on the survival of coliforms and enteric pathogens when exposed to free chlorine. Public health reports, 58: 1837 (1943). 26. SMITH, W. W. & BoDKIN, R. E. The influence of hydrogen IOn concentration on the bactericidal action of ozone and chlonne. Journal of bacteriology, 47: (Al7) 445 (1944). 27. SCARPINO, P. V. ET AL. A comparative study of the inactivation of viruses in water by chlorine. Water research, 6: 959 (1972). 28. CRAUN, G. E. & McCABE, L. J. Problems associated with metals in drinking water. Journal of the American Water Works Associatzon, 67: 593 (1975). 29. POURBAIX, M. Atlas of electrochemical equtbbria in aqueous solutions, 2nd ed. Houston, National Association of Corrosion Engineers, 1974, pp. 488-491. 30. McFARREN, E. F. ET AL. Water quality deterioration in the distribution system. Kansas City, MO, Water Quality Technology Conference, 1977. 31. TAYLOR, F. B. ET AL. The case for water-borne infectious hepatitis. American journal of public health, 56: 2093 ( 1966).
11. SODIUM 11.1 General description
The increased pollution of surface and groundwater during the past decades has resulted in a substantial increase in the sodium content of drinking-water in different regions of the world. Waterworks treatment processes and the practice of domestic water softening may also contribute to an increase in sodium levels in drinking-water. Sodium sulfate is employed in the manufacture of pigments and colours, the pulp and paper industries, and many other present-day manufacturing activities whose effluents contain enhanced sodium concentrations (1). Snow and ice control on roads accounts for the largest single use of sodium chloride in many countries and the quantity used for this purpose is steadily increasing. Sodium chloride is also used in the production of caustic soda, chlorine, and many industrial chemicals. Significant quantities are used in the food processing, slaughtering and meat packing, dairy, fishing, grain, and brewing industries (2). 11.2 Occurrence
Sodium is the most abundant of the alkali elements, compounds of sodium being widely distributed in nature, and constitutes 26 g/kg of the earth's crust. Soils contain sodium within the range 1-10 gjkg, it being mainly present as silicate minerals, such as amphiboles and feldspars. Some groundwaters contain high concentrations of sodium and, under certain circumstances, this can lead to increased salinity in rivers and streams. The control of ice and snow, by salting of highways, also results in an increased sodium burden in soils. Runoff may be sufficiently high to lead to contamination of public water supplies. It is estimated that between 25% and 50% of the salt used on a road infiltrates the groundwater (3). Sewage, industrial effiuents, seawater intrusion in coastal areas, and the use of sodium compounds for corrosion control and water-softening processes all contribute to sodium concentration in water because of the high solubility of sodium salts and minerals. Sodium concentrations vary considerably, depending on regional and local hydrological and geological conditions, the time of year, and salt utilization patterns. Sodium levels in groundwater vary widely but normally range between 6
286
11.
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287
and 130 mgjlitre (4). Higher levels may be associated with saline soils. In surface-water the sodium concentration may be less than I mgjlitre or exceed 300 mgjlitre, depending upon the geographical area (5-7). Reported sodium concentrations in public water supplies range from less than I mgjlitre to over 1000 mg/litre (7). Water softening can dramatically increase the sodium concentration in a supply (8). Sodium occurs naturally in all foods, varying considerably for different types, and food processing can have a marked effect on these levels. Fresh peas contain about 9 mg of sodium per kg whereas the level is 2.3 g/kg in drained canned peas and I g/kg in frozen peas (9). Fresh fruits and vegetables contain from less than 0.01 gjkg to about I gjkg, whereas cereals and cheeses may contain between 10 and 20 gjkg (10).
11.3 Routes of exposure Daily sodium intake by individuals varies considerably owing to variation in the sodium content of foods and personal variation in the use of salt as a food-seasoning agent. It was found that in a sample of 3833 people, 45% of the males and 30% of the females routinely added salt to their food (11, 12). The average daily intake of sodium for a Canadian male, aged between 20 and 64 years, has been estimated to be 3600 mg (10). Drinking-water contributes only a very small percentage of the daily intake of salt compared with the intake from food.
11.4 Health aspectsa Sodium is the most abundant extracellular cation and, together with its associated anions, contributes significantly to the osmotic activity of the extracellular fluid. Water and electrolyte balances are maintained by dietary intake in food and water and loss in urine, faeces, perspiration, and expired air. A normal 70-kg man contains approximately 69 g of metabolically active sodium and 45 litres of water. The control of water and sodium balance is achieved through a complex series of interrelated processes, involving both nervous and hormonal systems. The balance is maintained by a sodium loss rather than by control of absorption through the gut. The most important factor controlling loss is the mineralocorticoid hormone, aldosterone. Because the body has very effective methods of controlling sodium levels, sodium is not an acutely toxic metal. The deaths of 6 out of 14 infants were reported after they had mistakenly been given sodium in a concentration of 21 140 mgjlitre in their feed (13). Toxic symptoms of sodium poisoning include general involvement of the central nervous system with increase in sensitivity. There are some studies that show a a For further discussion of the health aspects of the sodmm content of dnnkmg-water, see Part Ill, section 16, p. 145.
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positive correlation between sodium intake and hypertension in man (14) and others that do not (15-17). A daily sodium intake in the- range 1600-9600 mg is generally considered to have no adverse effect on the health of normal individuals (18). To ensure a total dietary intake level of 500 mgjday would require limiting the concentration of sodium in drinking-water to about 20 mgjlitre which would incur considerable extra water treatment expense using the technologies at present available (19). To ensure that drinking-water is tasteless to the maJority of the consumers, the salt composition in the water should approximate to the salt content of saliva. The average sodium content of saliva is 300 mg/litre but may well exceed this value by a factor of two. The taste threshold of sodium in water depends upon the associated anion and the temperature of the solution. Sodium carbonate has the lowest taste threshold and the bicarbonate salt the highest. At room temperature the various threshold values for sodium were found to be about 20 mg/litre for Na 2 C0 3 , 150 mgjlitre for NaCl, 190 mgjlitre for NaN0 3 , 220 mg/litre for Na 2 S0 4 , and 420 mg/litre for NaHC0 3 . The recommended guideline value is 200 mgjlitre, which is based on the above taste thresholds and not on health considerations.
REFERENCES I. KILLIN, A. F. Sodium sulphate. In: Canadian minerals yearbook. Ottawa, Department of Energy, Mines and Resources, 1974. 2. KILLIN, A. F. Salt. In: Canadian minerals yearbook. Ottawa, Department of Energy, Mines and Resources, 1974. 3. McCoNNELL, H. H. & LEWIS, J. Add salt to taste. Environment, 14: 38 (1972). 4. BoND, R. G. & STRAUB, C. P. Genetic types of subterranean waters in relation to their salinity. In: Handbook of environmental control. Vol. 3. Water supply and treatment, 1st ed. Cleveland, OH, Chemical Rubber Co., 1973, p. 85. 5. WEILER, R. R. & CHAWLA, V. K. Dissolved mineral quality of Great Lakes waters. Proceedings of the 12th Conference on Great Lakes Research, Ann Arbor, MI, 1969, p. 801. 6. DoBSON, H. H. Principal ions and dissolved oxygen in Lake Ontario. Proceedings of the lOth Conference on Great Lakes Research, 1967, p. 337. 7. Sodium, chlorides and conductivity in drinking water. Report on a WHO working group. Copenhagen, WHO Regional Office for Europe, 1979 (EURO Reports and Studies, No. 2). 8. ELLIOTT, G. B. & ALEXANDER, E. A. Sodium from drinking water as an unsuspected cause of cardiac decompensation. Circulatwn, 23: 562 (1961). 9. MENEELY, G. R. A review of sources and the toxic effects of excess sodium chloride and the protective effect of extra potassium in the diet. Qualitas plantarum. Plant foods for human nutrition, 23: 3 (1973). 10. GoRMICAN, A. Inorganic elements in foods used in hospital menus. Journal of the American Dietetic Association, 56: 397 (1970). ll. Statement of basis and purpose for the national interim primary drinking water regulation, Washington, DC, US Environmental Protection Agency, 1975. 12. NATIONAL HEART AND LUNG INSTITUTE. The public and high blood pressure. Washington, DC, US Department of Health, Education and Welfare, 1973 (Publication No. (NIH) 74-356).
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289
13. FINBERG, L. ET AL. Mass accidental poisoning in infancy. Journal of the Amencan Medical Association, 184: U!7 (1963). 14. DAHL, L. K. Salt and hypertension. American journal of clinical nutrition, 25: 231 (1972). 15. KERKENDALL, W. M. The effects of dietary sodium on the blood pressure of normotensive man. In: Genest, J. & Koiw, E., ed. Hypertension, Heidelberg, SpringerVerlag, 1972, p. 360. 16. EVANS, J. G. & RosE, G. Hypertension. British medical bulletin, 27: 37 (1971). 17. DAUBER, T. R. ET AL. In: Stamber, J. et al., Environmental factors in hypertension. New York, Grune and Stratton Inc.,' 1967. 18. DAHL, L. K. Possible role of salt intake in the development of essential hypertension. In: Cottier, P. & Bock, K. D., ed. Essential hypertension: an internatwnal symposium. Heidelberg, Springer-Verlag, 1960, p. 53. 19. NATIONAL ACADEMY OF SCIENCES AND NATIONAL ACADEMY OF ENGINEERING. Water quality criteria, 1972. Washington, DC. US Government Printing Office, 1974.
12. SULFATE 12.1 General description
The majority of sulfates are soluble in water, the exceptions being the sulfates of lead, barium, and strontium (J). Dissolved sulfate is considered to be a permanent solute of water. It may, however, be reduced to sulfide, volatilized to the air as H 2 S, precipitated as an insoluble salt, or incorporated in living organisms (2). Sulfates are discharged into the aquatic environment in the wastes from many different industries (I, 3). Atmospheric sulfur dioxide (S0 2 ), formed by the combustion of fossil fuels and emitted by the metallurgical roasting processes, may also contribute to the sulfate content of surfacewater. Sulfur trioxide (S0 3 ), produced by the photolytic or catalytic oxidation of sulfur dioxide, combines with water vapour to form sulfuric acid, which is precipitated as "acid rain" or snow (3). 12.2 Occurrence
The concentration of sulfate in most freshwaters is very low, although levels of 20-50 mgjlitre are common in the eastern USA, Canada, and most of Europe (4, 5). The average sulfate concentration in the public water supplies in 23 of the larger cities in European communities is reported to be 64 mg/litre (range, 9-125 mgjlitre) (6). According to data collected over a 5-year period (1969-1973), at some 600 water sources supplying approximately 60% of the total population of England, Scotland and Wales, United Kingdom water supplies contain sulfate levels ranging from 4 to 303 mgjlitre.• Aluminium sulfate, which is extensively used as a flocculant for water treatment, may add 20-50 mg of sulfate per litre to the final water. Sulfate is not removed from water by conventional water treatment methods. Sulfate concentrations of bottled mineral water marketed in European communities average 223 mg/litre (range, 01182 mg/litre) (6). 12.3 Routes of exposure
Data concerning the daily dietary intake of sulfates are scarce. Sulfate compounds used in the USA as additives in foods are estimated to
° Figures quoted by P. Powell (Water quality and health division, Water Research Centre, Medmenham Laboratory, Medmenham, Marlow, Buckinghamshire, England) 290
12. (7, 8).
SULFATE
291
contribute an average of 453 mg to the daily sulfate intake of Americans Daily intake of sulfate from drinking-water, particularly if bottled mineral water is used, is extremely variable. 12.4 Health aspects Sulfate is poorly absorbed from the human intestine (8); it slowly penetrates the cellular membranes of mammals and is rapidly eliminated through the kidneys (9). The reported minimum lethal dose of magnesium sulfate in mammals is 200 mg/kg of body weight (10). Sulfate doses of 1.0-2.0 g have a cathartic effect on humans, resulting in the purgation of the alimentary canal (1). Infants ingesting sulfate equivalent to 21 mgjkg of body weight per day may also suffer from this effect. Magnesium sulfate at concentrations above 1000 mg/litre acts as a purgative in normal humans, but concentrations below this are apparently physiologically harmless (1, 10). Sensitive people are responsive to magnesium sulfate levels as low as 400 mg/litre and new users or those imbibing occasionally may be affected by concentrations in excess of 700 mg/litre. The human system adapts in the course of time to higher concentrations of sulfate in drinking-water (1 1). Taste threshold concentrations for the most prevalent sulfate salts are: 200-500 mg/litre for sodium sulfate; 250-900 mg/litre for calcium sulfate; and 400-600 mg/litre for magnesium sulfate (1, 11). Essentially on the basis of the above values, which are also allied to the cathartic effect of sulfate, a guideline value of 400 mg/litre is proposed. 12.5 Other aspects High sulfate concentrations in water may contribute to the corrosion of metals in the distribution system, particularly in waters having low alkalinity. REFERENCES I. McKEE, J. E. & WoLF, H. W. Water auality cr!leria, 2nd ed. Sacramento, CA, California State Water Quality Control Board, 1963, pp. 136, 213, 247, 270, 275-277. 2. NATIONAL RESEARCH COUNCIL, Drinking water and health, part 1. Washington, DC, National Academy of Sciences, 1977. 3. DELISLE, C. E. & SCHMIDT, J. W. The effects of sulphur on water and aquatic life in Canada. In: Sulphur and its morganic derivatwes in the Canadian environment. Ottawa, National Research Council of Canada, 1977, pp. 227-284. 4. HITCHCOCK, D. R. Biogenic contributions to atmospheric sulfate levels. Proceedings of the Second National Conference on Complete Water Re-use. Chicago, American Institute of Chemical Engineers, 1975. pp. 291-310. 5. KATZ, M. The Canadian sulphur problem. In: Sulphur and lis inorganic derivatives in the Canadian environment. Ottawa, National Research Council of Canada, 1977, pp. 21-67.
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6. AMAVIS, R., ET AL., ed. Hardness of drinking water and public health. Oxford, Pergamon Press, 1976, pp. 176-199. 7. Subcommittee on Research of GRAS (Generally Recognized as Safe) List (Phase II). Food ingredients, Washington, DC, National Academy of Sciences, 1972 (DHEW No. FDA 70-22). 8. NOVIKOV, Yu. V. & ERISMAN, F. F. The 'potable water' standard (GOST 2874-73). A new stage in development of water hygiene. Vestnik Akademii Medllsinskikh Nauk SSR, No. 3, 59 (1975). English translation No. 3, 76 (1975). 9. SENNING, A. Sulfur in organic and inorganic chemistry. Vol. 2. New York, Marcel Dekker Inc., 1972, p. 160. 10. ARTHUR D. LITILE, INC. Inorganic chemical pollutwn of freshwater. Washmgton, DC, US Environmental Protection Agency, 1971 (Water Pollution Control Research Series No. DPV 18010). II. ZoETEMAN, B. C. J. Sensory assessment of water quality. Oxford, Pergamon Press, 1980.
13. TASTE AND ODOUR 13.1 13.1.1 General description In the strict meaning of the word, the taste of water is the sensation that results from the interaction between the saliva and substances dissolved in the water, as perceived by receptors located in the taste buds. There are about 3000-10 000 taste buds in the mouth, most of which are located on the upper surface of the tongue, at its tip, sides, and rear surfaces. When "tasting" water, the senses of both gustation and olfaction are activated and it is extremely difficult to differentiate between them. Consequently, the combined effect of taste and odour is frequently classified as "taste". The taste perception is much less sensitive than that of smell (1, 2). However, water apparently free from odour may produce an offensive "taste" when taken in the mouth. In such cases, the higher oral temperature is conducive to the release into the nasal cavity of dissolved organic substances from the water. In this concentrated form, the sense of smell perceives the presence of the solutes and thus the "tasting" of water, as opposed to just smelling, is often a more sensitive sensory assessment of quality (3). In the assessment of drinking-water quality, the sensations of taste and odour are complementary. In general, the sense of taste is most useful in detecting inorganic constituents of drinking-water, while the sense of smell is more useful in detecting organic constituents. Taste tests in general have received considerable criticism (4-6). In taste threshold tests, the use of distilled water as the standard for "tasteless" water and as a mouth rinse introduces a bias into the results. Water containing salts in concentrations higher or lower than those in saliva will be perceived as different from the saliva by the sense of taste. Therefore, minimum concentrations of ions, such as sodium, chloride, calcium, and bicarbonate, are essential to make a water tasteless. Taste tests that involve large laboratory panels of judges can be timeconsuming and difficult for small treatment facilities to perform (4), and the use of consumer panels has been advocated as an alternative (3). Taste tests performed in water treatment plants may underestimate the taste of the water as delivered to the consumer because, at the plant, objectionable tastes can be masked by residual chlorine. The masking 293
Taste
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effect diminishes as the chlorine residual decreases in the distribution system (4). Chemical dechlorination of the water prior to assessment may augment the taste (4). Panels of consumers assessing water taste in their homes can overcome these problems (3). Additional difficulty arises from differences in test procedures, such as quality of dilution water, number of and motivation of panel members, and variations in the statistical treatment and interpretation of taste-test data. Suitable methods for determining the taste intensity of drinking-water are similar to those described for odour.
13.1.2 Occurrence Taste and odour problems in drinking-water supplies account for the largest single class of consumer complaints. They may occur in any type of water and at any time of year. Some are due to natural causes, others to man's industrial activity. They may be associated primarily with the raw water, the treatment method, the distribution system, or with a combination of all three. Taste and odour survey results in water treatment plants in Canada, the Netherlands, and the USA have been published (3, 7). Groundwater sources for supply normally have the fewest taste problems. The majority of the surface-water supplies are subject to seasonal variation in taste and odour problems, which suggests that such problems may be of biological origin. High concentrations of colour and turbidity in water are often associated with nonspecific taste (and odour) problems (8). Taste acuity is reported to depend on temperature (9, 10), and the degree to which taste is influenced by temperature is a function of the specific taste-causing substance (10). The growth rate of microorganisms, some of which may produce bad-tasting metabolites, is enhanced by higher temperatures, as also is the rate of formation of offensive-tasting corrosion products. Where pH controls the equilibrium concentration of the neutral and ionized forms of a substance in solution it can notably influence its taste (and odour). Several studies of the organoleptic properties of residual chlorine have been performed over the years, but further work will be required before the key questions can be answered. Under ideal conditions, the amount of free available chlorine at the consumer's tap should be sufficiently high to ensure the microbial safety of the water and sufficiently low to avoid objectionable taste and odour problems. Taste and odour thresholds of residual chlorine are thus of considerable interest and the most recent investigation into this subject found that the average taste threshold concentration of free residual chlorine increased from 0.075 mgjlitre to 0.450 mg/litre as the pH increased from 5.0 to 9.0 (11).
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The average threshold was 0.156 mgjlitre, with a range of 0.020.29 mgjlitre at pH 7.0. Although much more work is clearly called for, it is probable that most treatment plant managers are aware of the appropriate balance between the applied chlorine residual and consumer complaints. In mineralized water (1 2) and in coffee (1 3), taste thresholds for chlorine indicate that other constituents causing taste in water can influence the magnitude of the chlorine threshold concentration. Thus, the nature of the raw-water supply will be one major factor in the detectable taste threshold concentration for residual chlorine. In certain countries, it is known that consumers are assured of the safety of their water supply by the presence of a slight taste of chlorine. Many of the inorganic substances occurring in water exert an unpleasant taste at concentrations much lower than those required for acute toxic effects. Limits for such substances are, therefore, set at concentrations that reflect levels found to be objectionable to consumers. These substances are discussed briefly below. Taste thresholds in distilled water for the major cations of drinkingwater, i.e., calcium, magnesium, sodium, and potassium, have been reported to be approximately 100, 30, 100 and 300 mg/litre, respectively (3, 14). The uncertainty associated with these evaluations is largely due to the influence on taste of their associated anions. Taste threshold tests for iron, as Fe(II), have shown that the most sensitive 5% of the members of a test panel can detect concentrations of 0.04 mg/litre in distilled water but in a mineralized spring water having a total dissolved solids content of 500 mg/litre the threshold value was 0.12 mgjlitre (15). Zinc could be detected at a concentration of 4.3 mg/litre in distilled water, but only at 6.8 mgjlitre in mineralized spring water (1 5). Reliable data on the taste and odour thresholds for sulfide in water are sparse, and the situation is complicated somewhat by the influence of pH on the sulfide-bisulfide-hydrogen sulfide equilibrium. 13.1.3 Health aspects
The presence of objectionable tastes in a public water supply may cause consumers to seek alternative sources of potable water, which may or may not be subject to the same degree of microbial protection afforded by the rejected supply. This has been exemplified by a survey, conducted by the California State Department of Public Health, which found that dissatisfied consumers were large-scale purchasers of bottled water. In the Netherlands, it was shown that water possessing an offensive taste resulted in a lower consumption of tap-water (3). The taste of water unfortunately provides no assurance that such water is free of pathogens or toxic inorganic chemicals. Fortunately, median taste thresholds of inorganic substances are generally much lower than the concentrations that cause adverse health effects.
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In a public water supply, short-term changes in the normal taste may signal changes in the quality of the raw-water source, deficiencies in the treatment process, or chemical corrosion and biological growths in the distribution system. The objective is, therefore, to provide water that is free of objectionable taste for the majority (90 %) of consumers. The most direct way to verify this objective is to seek periodically the views of a selected consumer group. Laboratory panels can be applied to assess water taste using a category scale (e.g., good-not observable-weakobjectionable-bad) or by assessing the taste number by the forced choice method (see section 13.2.1., p. 297). In the latter case, it is recommended that the taste number of the drinking-water be kept below 1. Although consumer panels are most suited for assessing the taste of tap-water, laboratory panels are valuable for assessing the taste of the water during treatment. Special local circumstances may occur resulting in an unavoidable perceptible taste in the water. In such cases, local health authorities should give priority to disinfection to ensure the control of diseasecausing contaminants, such as pathogenic bacteria.
REFERENCES I. RoSEN, A. A. & BooTH, R. L. Taste and odour control. In: Water quality and treatment, 3rd ed. Toronto, McGraw-Hill, 1971, p. 225. 2. SUFFETT, I. H. & SEGALL, S. Detecting taste and odour in dnnking. water. Journal of the American Water Works AssociatiOn, 63: 605 (1971). 3. ZoETEMAN, B. C. J. Sensory assessment of water quality. Oxford, Pergamon Press, 1980. 4. BAKER, R. A. Dechlorination and sensory control. Journal of the American Water Works Association, 56: 1578 (1964). 5. BRUVOLD, W. H. Human perception and evaluation of water quality. CRC critical reviews in environmental control, 5: 153 (1975). 6. SWETS, J. A. Is there a sensory threshold? Science, 134: 168 (1961). 7. Handbook of taste and odour control experiences in the US and Canada. Denver, co: American Water Works Association, 1976. 8. RIDDICK, T. M. Zeta potential and polymers. Journal of the American Water Works Association, 58: 719 (1966). 9. Standard methods for the examination of water and waste water, 14th ed. Washington, DC, American Public Health Association, 1976, p. 121. 10. PANGBORN, R. M. & BERTOLERO, L. L. Influence of temperature on taste intensity and degree of linking of drinking water. Journal of the American Water Works Assocwtion, 64: 511 (1972). II. BRYAN, P. E. ET AL. Taste thresholds of halogens in water. Journal of the American Water Works Assoctation, 65: 363 (1973). 12. PANGBORN, R. M. ET AL. Sensory examination of mineralized, chlorinated waters. Journal of the American Water Works Association, 62: 572 (1970). 13. CAMPBELL, C. L. ET AL. Effects of certain chemicals in water on the flavour of brewed coffee. Food research, 23: 575 (1958). 14. NATIONAL ACADEMY Of SciENCES. Water quality criteria 1972. Washington, DC, US Government Printing Office, 1973 (EPA-R-73-033). 15. COHEN, J. M. ET AL. Taste threshold concentrations of metals in drinkmg water. Journal of the American Water Works Association, 52: 660 (1960).
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13.2. Odour 13.2.1 General description
The odour of drinking-water may be defined as the sensation that is due to the presence of substances having an appreciable vapour pressure and that stimulate the human sensory organs in the nasal and sinus cavities. The sense of smell will generally respond to much lower concentrations (a few micrograms per litre or less) of a substance than will the sense of taste (a few milligrams per litre or more). The odour intensity of water is usually measured in terms of its threshold odour number (TON), which is defined as the geometric mean of the dilution ratios with odour-free water, the odour of which is just detectable by a panel of judges under very carefully controlled test conditions (1). As in the case of taste threshold measurements, laboratory panel quality ratings or mean threshold values are only estimates of these values for the entire consuming population (1). Water with a TON of 2 may, depending upon its nature, stimulate more consumer complaints than other water having a TON of 4 (2). An alternative to the TON determination for odour intensity is the odour number determination by the forced choice method, which has certain advantages (3). Each member of a panel of judges is presented with a series of paired samples, one of which is a dilution of the sample being tested, the other being an odour-free control. Each panel member has to judge, for each pair, which of the two has the stronger odour. A choice must be made even when no difference is perceived. The percentage of right responses is calculated for each dilution and corrected for the 50% probability that the right flask has been indicated by chance. The odour number is taken as the dilution for which the corrected right response rate is 50%, calculated from the graph of dilution against corrected response rate. Further details of this method are given by Zoeteman (3). Good results have also been obtained by using the less time-consuming "interval scaling method", which is most suitable for drinking-water with very low odour intensity. Odour intensity measurements are usually nonspecific. However, intensity measurements for specific substances in water are normally reported in terms of their odour threshold concentrations (4). This is the concentration of the substance in water whose odour can be detected by 50% of the panel members. The wide variation in individual odour detection has been markedly illustrated by the use of odour threshold concentrations. In a large population, the most sensitive 5% is able to detect odour reliably at one-hundredth of the average odour threshold concentration (5). To obtain reliable and responsible data, therefore, odour assessment of water quality should be carried out by large panels. Smaller panels may be employed, if it is more convenient to do so, but it must be appreciated that the accuracy and reliability of the determination will thereby be reduced.
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It is important to specify the temperature at which odour intensity measurements are made, since odour intensity is related to the vapour pressure of any odour-causing substance and, hence, will be directly related to the water temperature.
13.2.2 Occurrence Water odour is predominantly due to the presence of organic substances in water. Very many odour-producing compounds have been reported in water (6, 7). Objectionable odours in drinking-water may be of either biological or industrial origin, and some odours of natural origin may be due indirectly to human activities; for example, the dumping of raw sewage into the aquatic environment enhances biological growth, which may produce odorous products. Natural odours tend to be described as earthy, musty, or sour, on the one hand, or as fishy, grassy, or cucumber-like, on the other, involving compounds such as geosmin and decanal (8-11). Industrially derived odours often smell like such substances as petroleum or creosote or have a medicinal odour. Typical examples in this category are naphthalene and the chlorinated benzenes and phenols (5). Groundwater tends to have fewer odour problems, although odours are not restricted to any single type of water nor to any particular season of the year. Odours may also be produced under stagnant water conditions in low-flow sections of distribution systems or in raw- and finished-water reservoirs. Water-purification processes may convert substances with weak odours (such as amines and phenols) into substances possessing very intense odours (such as chloramines and chlorophenols) (12). The proliferation of nuisance organisms, such as iron and sulfur bacteria, in distribution systems may also be a source of odour. The nonspecific fishy, grassy, and musty odours normally associated with biological growth tend to occur most frequently in warm surfacewater in the warmer months of the year (10, 11). Surveys of taste and odour problems in Canada and the USA identified 50 so-called "nuisance" organisms responsible for odours in drinking-water. The very intense musty odours of substances produced by the actinomycetes group of organisms can be a major source of odour contamination in public water supplies. It has been recommended that raw-water supplies be monitored for actinomycetes (1, 10). Although such monitoring in some localities has positively correlated the presence of actinomycetes with odour problems, other treatment plants have found the converse to be true.
13.2.3 Health aspects Odour in potable water is almost invariably indicative of some form of pollution of the water source or of malfunction during water treatment
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299
or distribution. Odours of biological origin are indicative of increased biological activity, which may include an increased loading of dangerous pathogens on the system. Odours of industrial origin are associated with pollution of the source-water with commercial waste products, some of which may be toxic. Sanitary surveys should include investigations for potential or existing sources of odour, and attempts should always be made to identify the source of an existing odour problem. Some chemical contaminants, of concern because of their toxic properties, may also cause odour problems. The threshold odour for hydrogen cyanide in water, for example, has been reported to be O.OOI mg/litre (6). A value for cyanide in drinking-water based on this information would be one-hundredth the value recommended elsewhere in these guidelines (see Vol. I, p. 55). In this and other examples, the sense of smell is more sensitive than the best available analytical instrumentation. Except possibly for chlordane, the odours of pesticides in water arc too weak to permit their detection at or below their recommended guideline values (3). Ideally, drinking-water should have no observable odour to any consumer. However, owing to the large differences in individual odour sensitivity within a population, a more realistic objective is to provide a water free of objectionable odour for the large majority of the population (e.g., 90 '!'~). The most direct way of achieving this objective is through the cooperation of a large consumer panel (e.g., 100 consumers) in a supply area who are asked to make periodic assessments of water odour and taste in their homes. Participants should indicate their observations on a category scale (e.g., good-not perceptible-mildly objectionable-bad). Laboratory panels tend to be more critical in assessing water odour and taste. However, laboratory panels of I0-20 trained persons can also indicate if a water is aesthetically acceptable to the majority of the consumers (3). If the odour number is measured at room temperature by the forced choice method using a selected laboratory panel, it is recommended that the objective value be less than I, unless local circumstances demand a disinfection practice requiring perceptible free chlorine residuals.
REFERENCES I. Standard methods for the examination of water and wastewater, 14th ed. Washington, DC, APHA, AWWA, WPCF, 1976, p. 75. 2. BAKER, R. A. Dechlorination and sensory control. Journal of the American Water Works Association, 56: 1578 (1964). 3. ZOETEMAN, B. C. J. Sensory assessments of water quality. Oxford, Pergamon Press, 1980. 4. BAKER, R. A. Threshold odors of organic chemicals. Journal of the American Water Works Association, 55: 913 (1963). 5. ZoETEMAN, B. C. J. & PiET, G. J. Cause and identification of taste and odour compounds in water. Science of the total envzronment, 3: 103 (1974).
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6. VAN GEMERT, L. J. & NETTENBREIJER, A. H., ed. Compilation of odour threshold values in air and water. Voorburg, National Institute for Water Supply; Zeist, Netherlands, Central Institute for Nutrition and Food Research, TNO, 1977. 7. STAHL, W. H., ed. Compilation of odor and taste threshold values data. Philadelphia, American Society for Testing and Materials, 1973 (ASTM Data Series Publication No. OS 48). 8. Handbook of taste and odour control experiences in the US and Canada. Denver, CO, American Water Works Association, 1976, p. XIV-I. 9. ZoETEMAN, B. C. J. & PIET, G. J. On the nature of odours in drinking water resources of the Netherlands. Science of the total environment, 1: 399 (1972/73). 10. MORRIS, R. L. ET EL. Chemical aspects of Actinomycetes metabolites as contributors of taste and odour. Journal of the American Water Works Association, 55: 1380 (1963). II. MCKEE, J. E. & WOLF, H. W., ed. Water quality criteria. 2nd ed. Sacramento, CA, California State Water Quality Control Board, 1963 (Publication No. 3-A). 12. BURTTSCHELL, R. H. ET AL. Chlorine derivatives of phenol causing taste and odour. Journal of the American Water Works Association, 51: 205 (1959).
14. TEMPERATURE 14.1 General description
In general, the rates of chemical reactions decrease with decreasing temperature. The relative concentrations of reactants and products in chemical equilibria can also change with temperature. Temperature can, therefore, affect every aspect of the treatment and delivery of potable water. 14.2 Physical aspects
Cool drinking-water is preferable to warm. The intensity of taste is greatest for water at room temperature and is significantly reduced by chilling or heating. Increasing the temperature will also increase the vapour pressure of trace volatile compounds in drinking-water and may lead to increased odour. Turbidity and colour are indirectly related to temperature as the efficiency of coagulation is strongly temperature-dependent. The optimum pH for coagulation decreases as temperature increases (1). In order to achieve the most economical use of a coagulant, therefore, jar tests should be carried out at the temperature of the treated water and not at room temperature (2). As temperature decreases, the viscosity of water increases, and the rates of sedimentation and filtration decrease. The efficiency of colour and turbidity removal by coagulation, sedimentation, and filtration may be less under winter temperature conditions than in summer. The decreased efficiency of turbidity removal by filtration at lower temperatures is possibly due to a reduction in floc-strength or average particle size (3). Filtration through activated carbon is also affected by temperature; the adsorptivity of activated carbon increases as the temperature drops (4). 14.3 Microbiological aspects
The microbiological characteristics of drinking-water are related to temperature through its effect on water-treatment processes, especially disinfection, and its effect on both growth and survival of microorganisms. In general, disinfection is aided by increased temperature. Working with Escherichia coli, Butterfield and co-workers observed a 5-fold 301
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increase in the bactericidal effectiveness of chlorine between 20 and 25 oc compared with that between 2 and 5 oc (5). In a US Army study, Ames & Smith found a 9-fold increase in effectiveness between 8 oc and 40oC (6). Chambers reported that the influence of temperature on the effectiveness of chlorine disinfection was insignificant at pH values between 7.0 and 8.5, but at higher pH values over the temperature range 4 oc to 22 °C, a 4- to 8-fold increase in effectiveness was observed (7). Similar results have been obtained with viruses (8). It has been reported that the inactivation of Mycobacterium fortuitum by ozone increases with temperature (9). Coagulation and sedimentation of water reduce the number of suspended microorganisms and, as discussed earlier, temperature affects these processes. At a given pH, higher temperature leads to greater dissociation of hypochlorous acid. The magnitude of this effect on the germicidal efficiency of chlorinated water is, however, of secondary importance to the larger, and opposite, effect of increased germicidal action at higher temperatures. Published information is somewhat equivocal about the effect of temperature on bacterial survival in water (10). Seasonal variations in coliform counts in raw-water sources have been observed (11). However, temperature would be only one of a number of factors leading to this variation. At low temperatures, viruses can survive considerably longer than bacteria and survival times of up to 6 months have been reported for poliovirus in tap-water at low temperature (12). However, an epidemiological study of viral hepatitis A in 13 cities in the USA showed no correlation between infection rate and raw water temperature (13). The survival time in water of the cysts and ova of parasitic worms is shortened by higher temperatures. For example, Schistosoma ova die in 9 days at 29-32 °C, in 3 weeks at 15-24 °C, and in 3 months at 7 oc (14).
The growth of nuisance organisms is enhanced by warm water conditions and could lead to the development of unpleasant tastes and odours. 14.4 Chemical aspects The rate of formation of trihalomethanes in chlorinated drinkingwater increases with temperature (15) and is perhaps the single most important factor influencing seasonal variation in trihalomethane concentrations (16). The effect of temperature on corrosion in water treatment systems demonstrated that corrosion increased as a function of temperature (17). Sodium hydroxide adjustment of the pH halved this increase over the same temperature range. At temperatures below 10 °C, however, water containing sodium hydroxide showed a higher corrosion rate than the untreated water. The corrosion rate is also a function of the dissolved
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oxygen concentration in the water. Dissolved oxygen variation with temperature is small compared with the much larger (and opposite) change in corrosion rates cited above. The dissolved oxygen content, however, plays an unimportant role in the temperature dependence of corrosion. The solubility product of calcium carbonate decreases with temperature. At low alkalinities, however (50 mgjlitre as calcium carbonate), the decrease in pH with increased temperature actually increases the solubility of calcium carbonate. This effect on the saturation index tends to reduce incrustation by carbonate and at the same time increases the agressivity of the water, leading to increased corrosion in hot-water systems (3).
REFERENCES I. MAUDLING, J. S. & HARRIS, R. H. Effect of ionic environment and temperature on the coagulation of color-causing organic compounds with ferric sulfate. Journal of the American Water Works Association, 60: 460 (1968). 2. CAMP, .T. R. ET AL. Effects of temperature on rate of floc formation Journal of the American Water Works Association, 32: 1913 (1940). 3. AMERICAN WATER WORKS AsSOCIATION. Water quality and treatment, 3rd ed. Toronto, McGraw-Hill, 1971, pp. 89, 305. 4. WEBER, W. J. & MORRIS, J. C. Equilibria and capacities for adsorption on carbon.
5. 6. 7. 8. 9. 10. II. 12. 13. 14. 15. 16. 17.
Journal of the Sanitary Engineering Division, Proceedings of the American Society of Civil Engineers, 90: (5A3) 79 (1964). BUTTERFIELD, C. T. ET AL. Influence of pH and temperature on the survival of coliforms and enteric pathogens when exposed to free chlorine. Public health reports 58: 1837 (1943). AMES, M. & WHITNEY-SMITH, W. Journal of bacteriology, 47: 445 (1944). CHAMBERS, C. W. An overview of the problems of disinfection. Symposium on wastewater treatment in cold climates. Saskatoon, Canada, University of Saskatchewan, 1974, p. 423 (EPS 3-WP-74-3). WHITE, G. C. Disinfection: The last line of defense for potable water. Journal of the American Water Works Association, 67: 410 (1975). FAROOQ, S. ET AL. Influence of temperature and UV light on disinfection with ozone. Water research II: 737 (1977). RuDOLFS, W. ET AL. Literature review on the occurrence and survival of enteric, pathogenic, and relative organisms in soil, water, sewage, and sludges, and on vegetation. Sewage and industrial wates, 22: 1261 (1950). RAo. S. S. & HENDERSON, J. Summary report of microbiological baseline data on Lake Superior 1973. Ottawa, Environment Canada, Inland Waters Directorate, 1974, (Scientific Series No. 45). HEALTH AND WELFARE CANADA. Microbiological quality of drinking water. Ottawa, Health and Welfare Canada, 1977, (77-EHD-2). TAYLOR, F. B. ET AL. The case for water-borne infectious hepatitis. American Journal of public health, 56: 2093 (1966). Temperature. In: Water quality criteria, 2nd ed. Sacramento, CA, California State Water Quality Control Board, 1963, p. 283. STEVENS. A. A. ET AL. Chlorination of organics in drinking water. Journal of the American Water Works Association, 68: 615 (1976). SMILLIE, R. D. ET AL. Organics in Ontario drinking water, Part II. Toronto, Ontario Ministry of the Environment, 1977. MULLEN. E. D. & RiTTER, J. A. Potable water corrosion control. Journal of the American Water Works Association, 66: 473 (1974).
15. TOTAL DISSOLVED SOLIDS 15.1 General description The total dissolved solids (TDS) in water comprise inorganic salts and small amounts of organic matter. The principal ions contributing to TDS are carbonate, bicarbonate, chloride, sulfate, nitrate, sodium, potassium, calcium, and magnesium (1). Total dissolved solids influence other qualities of drinking-water, such as taste, hardness, corrosion properties, and tendency to incrustation.
15.2 Occurrence Total dissolved solids in water may originate from natural sources, sewage effluent discharges, urban runoff, or industrial waste discharges. Waters in contact with granite, siliceous sand, well-leached soil, or other relatively insoluble material have TDS levels of less than 30 mgjlitre (2). Waters in precambrian shield areas generally have TDS levels less than 65 mg/litre (3). Waters in areas of palaeozoic and mesozoic sedimentary rock have higher TDS levels, ranging from as little as 195 to 1100 mg/litre (3); carbonates, chlorides, calcium, magnesium, and sulfates are the principal ions present (2, 4). In addition to these natural leaching processes, sewage and industrial wastes may lead to further increases. Under arid conditions, the TDS of the smaller streams may increase to levels of 15 gjlitre (4). Elsewhere, levels in excess of 35 g of TDS per litre have been recorded in briny waters (3). The use of salt for snow and ice control on roads during winter weather contaminates both surface and groundwater sources, increasing the TDS of waters noticeably in some countries. In the winter of 1969-70, Denmark applied 203 000 tonnes of salt to its road system; 2. 5 million tonnes were used in Canada in 1974, and in the USA, salt consumption for road de-icing in 1970 was calculated as 9 million tonnes-all contributing to the TDS levels of streams, waterways, and groundwater supplies (5).
15.3 Health aspects There is no evidence of deleterious physiological reactions occurring in persons consuming drinking-water supplies that have TDS levels in 304
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excess of IOOOmg/litre (2, 4, 6). The results of certain epidemiological studies would appear to suggest that TDS in drinking-water may even have beneficial health effects. The common dissolved mineral salts are claimed to affect the taste of water (7-JJ). The effects that many of these minerals have on taste have been discussed in the separate reviews of these constituents and in the section dealing with taste (see p. 293). Bruvold et a!. (8) have rated the palatability of drinking-water according to the TDS level thus: Excellent: Good: Fair: Poor: Unacceptable: less than 300 mg/litre between 300 and 600 mg/litre between 600 and 900 mg/litre between 900 and 1200 mgjlitre greater than 1200 mgjlitre
Water with extremely low TDS levels may also be unacceptable because of its flat, insipid taste.
15.4 Other aspects
Certain components of TDS, such as chlorides, sulfates, magnesium, calcium, and carbonates, affect corrosion or incrustation in water distribution systems (2). Total dissolved solids are not generally removed in conventional water-treatment plants. Although no deleterious physiological effect has been recorded with total dissolved solids in water above I 000 mg/litre, it was considered that it would, as a rule, be unacceptable to exceed this level, which is recommended as a guideline value. REFERENCES l. Quality criteria for water. Washington, DC, US Environmental Protection Agency,
1976 (EPA-440 /9-76-023). 2. RAINWATER, F. H. & THATCHER, L. L. Methods for collection and analysis of water samples. Geological Survey Water-Supply Paper, Washington, DC, US Government Printing Office, 1960. 3. GARRISON INVESTIGATIVE BoARD. Water quality report (Appendix A). Garrison Diversion Study, Report to the International Joint Commission: US-Canada, Windsor, Ontario, 1977. 4. DURFOR, C. J. & BECKER, E. Constituents and properties of water. In: Pettyjohn, W. A., ed. Water quality in a stressed environment. Minnesota, Burgess Publishing Company, 1972. 5. Sodium, chlorides and conductivity in drinking-water supplies. Copenhagen, WHO Regional Office for Europe, 1979 (EURO Reports and Studies, No. 2). 6. ONGERTH, H. J. ET AL. The taste of water. Public health reports, 79: 351 (1964). 7. BRUVOLD, W. H. & PANGBORN, R. M. Rated acceptability of mineral taste in water. Journal of applied psychology, 50: 22 (1966).
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BRUVOLD, W. H. ET AL. Consumer attitudes toward mineral taste m domestic water. Journal of the Amencan Water Works Association, 59: 547 (1967). 9. BRUVOLD, W. H. Scales for rating the taste of water. Journal of applted psychology, 52: 245 (1968). 10. BRUVOLD, W. H. Mineral taste and the potability of domestic water. Water research, 4: 331 (1970). II. BRUVOLD, W. H. & ONGERTH, H. J. Taste quality of minerahzed water. Journal of the American Water Works Associatwn, 61: 170 (1969).
16. TURBIDITY 16.1 General description
Turbidity in water is caused by the presence of suspended matter, such as clay, silt, colloidal organic particles, plankton, and other microscopic organisms. Turbidity is an expres.>ion of certain light-scattering and light~absorbing properties of the water sample. It is a parameter whose significance is to a large extent dependent on the measurement technique. The total intensity and angular distribution of light scattered from turbid water represent the overall effects of intraparticle and interparticle interactions. They depend, in a complex manner, on such factors as the number, size, shape, and refractive index of the foreign particles, and on the wavelength of the incident light. Complex though the factors are, a number of generalizations can be made (1, 2). Five methods may be used in the measurement of water turbidity, but only two of these, nephelometry and turbidimetry, form the basis of present standard methods (3-6). Historically, turbidity measurements in the waste- and drinking-water fields have been based on the Jackson candle turbidimeter (7). Jackson candle turbidity is an empirical quantity based on the measurement in a special graduated vessel of that depth of sample that is just sufficient to extinguish the image of a burning standard candle observed vertically through the sample. The Jackson turbidity unit (JTU) is defined in terms of that depth; a depth of 21.5 em corresponds to 100 JTU (5, 6). The Jackson candle turbidimeter is applicable only to turbidities greater than 25 JTU and as such has limited applicability to the monitoring of drinking-water. Improved instruments, such as the Patterson turbidimeter (3), using electrical light sources and mirror optics, can measure lower values. As an alternative to JTU, turbidimeters can be calibrated in terms of the concentration of suspended solids (mg/litre) that gives rise to a certain turbidity (a gravimetric definition). Diatomaceous earths are commonly used to form the standard suspensions. This type of definition (sometimes called the Fuller's earth scale) is arbitrary and is specific for the type and particle size of the particular clay used (8).
The nephelometric method is the current method of choice for turbidity measurement (6, 9-11). Nephelometric turbidimeters measure the intensity of light scattered at 90° to the path of the incident light. Differences in the physical design of such turbidimeters will cause differences in measured turbidity values. 307
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In an attempt to minimize such differences, the light source equipment and detector geometry and the calibration method are specifically defined (6). Suspensions of formazin polymer as turbidity reference standards have been almost universally adopted by the water industry (7). A suspension of formazin, formed by the interaction under specified conditions of hydrazine sulfate (50 mgjlitre) with hexamethylaminetetramine (500 mgjlitre), has a defined turbidity of 40 nephelometric turbidity units (NTU) (6, 10). These are also known as formazin turbidity units (FTU). When measured on a candle turbidimeter this standard suspension has ~ turbidity of about 40 JTU (6). As defined by the above methods, turbidity represents a nonspecific measurement of suspended solids concentration. Electronic particle counters have recently become available and these are capable of accurately counting and recording the number of suspended particles as a function of size. Generally there is a telationship between turbidity (in the range 0.2-1 NTU) and particle counts; good point-for-point agreement between the two methods does not exist (1 1).
16.2 Occurrence The particles that cause turbidity in water range in size from colloidal dimensions (approximately I 0 nm) to diameters of the order of 0.1 mm. They may be divided into three general classes: clays; organic particles resulting from decomposition of plant and animal debris; and fibrous particles, e.g., asbestos minerals (1 2). Clay particles generally have an upper particle-size limit of about 0.002 mm diameter. Soil particles derived from the land surface by erosion constitute the major part of suspended material in most natural waters. The coarser sand and silt fractions are wholly or partially coated with organic material. Phyllosilicate clay particles as well as non-clay material, such as iron and aluminium oxides and hydroxides, quartz, amorphous silica, carbonates, and feldspar, constitute the clay fraction (12). Clays and organic particles are also often found together as a "clay-organic" complex (12). Humic substances have a much higher ionic exchange capacity than inorganic clays (13), and in many instances the effect of humic components predominates. Organic turbidity resulting from the accumulation of higher microorganisms may occur in such large amounts that waters become unsightly and turbid. Examples of turbidity due to microorganisms are the summer blooms of blue-green algae in surface-water, algal debris, and the detritus from iron bacteria in distribution systems (red-water is such a manifestation) (14). Raw-water turbidity can vary from less than I NTU to greater than 1000 NTU. Removal of turbidity may be achieved by simple filtration, or more effectively, by a combination of coagulation, sedimentation, and filtration. Filtration through sand beds or other single medium filters can
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consistently produce a water with a turbidity of 1 NTU or less. Continuous monitoring of turbidity throughout the treatment stages is a valuable aid in attaining such a performance. 16.3 Relationship with other water-quality parameters
The turbidity of water is related to or affects many other indicators of drinking-water quality. The particulate matter may also be a source of nutrients and protection for some microorganisms. There is evidence that a large part of the colour in water arises from colloidal particles, 50% of such colour being due to a "colloidal fraction" of humic substances (15). True colour is, therefore, defined as the colour of water from which the turbidity has been removed (16). The relationship between high turbidity, in both raw and filtered water, and taste and odour has also long been recognized, and suspended particulate matter in a potable water supply renders the water unattractive to the consumer (1 7). The presence of turbidity can have a significant effect on the microbiological quality of drinking-water. The detection of bacteria and viruses in drinking-water may be complicated by the presence of turbidity. In water, microbial growth is most extensive on the surfaces of particles and inside loose, naturally occurring, floc and floc formed during coagulation treatment (see Part I, "Microbiological aspects"). This growth is facilitated because nutrients are adsorbed on to surfaces and attached bacteria are thus able to grow more efficiently compared with those in free suspension (18, 19). Similarly it has been demonstrated that river silt readily adsorbs viruses (20). In the water-treatment process of coagulation, bacteria and viruses become trapped in the floc formed and are removed along with turbidity (21, 22). Breakthrough of filter beds by floc is also accompanied by an increase in virus penetration, even though the turbidity of the finished water remains below 0.5 JTU (23).
Particulate matter, whether organic, inorganic or due to higher microorganisms, can protect bacteria and viruses from the action of disinfectants. Sanderson & Kelly reported the presence of coliform organisms in water with turbidities ranging between 3.8 and 84 NTU, even after treatment with chlorine producing free chlorine residuals between 0.1 and 0.5 mg/litre and a minimum contact time of 30 minutes. 0 Neefe and co-workers showed that chlorination of drinkingwater deliberately contaminated with faecal matter was, by itself, insufficient protection against viral hepatitis A (24). Only by coagulation and filtration prior to chlorination could the water be rendered safe to drink. In laboratory tests, the presence in water of various clays and humic acid was shown to protect "Klebsiella aerogenes" (invalid) from disinfection by ultraviolet light (25). a Comments following a paper by Clark, N. A. et al., reference 21.
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Consumption of highly turbid, chlorinated water may be a dangerous health risk (26-29). The adsorptive capacity of some suspended particulates can lead to the entrapment of undesirable inorganic and organic compounds present in the water and in this way, turbidity can bear an indirect relationship to the health aspects of water quality. Most important in this respect is the organic or humic component of turbidity (30-33). The strength of some metal-humate complexes in the turbidity fraction may complicate the analytical measurement of trace metals in natural waters resulting in an underestimation of the metal (34). Organic molecules are also adsorbed by natural organic matter. Herbicides such as 2,4- D, Paraquat, and Diquat can be adsorbed on to clay-humic acid particulates, the adsorption being greatly influenced by metal cations present in the humic material (35). The presence of turbidity, therefore, may also interfere with the detection of biocides in water samples. In so far as turbidity is used as a measure of the efficiency of the removal of particulate matter throughout the purification process of water, low turbidity in the finished product is an indication of the effectiveness of the coagulation, sedimentation, and filtration processes. 16.4 Health aspects
A turbidity in excess of the guideline value of 5 NTU is generally objectionable to consumers (36). The perception of higher turbidity in water at the consumer's tap than in that entering the distribution system may indicate post-treatment contamination, corrosion, or other distribution problems. Consequently, as excessive turbidity can protect microorganisms from the effects of disinfection, stimulate the growth of bacteria in the water, and itself exert a significant chlorine demand, it is vitally important in producing safe drinking-water, using chlorine as disinfectant, that turbidity should be kept low, preferably below 1 NTU.
REFERENCES I. BLACK, A. P. & HANNAH, S. A. Measurement of low turbidities. Journal of the American Water Works Associatwn, 57: 901 (1965). 2. McCLUNEY, W. R. Radiometry of water turbidity measurements. Journal of the Water Pollution Control Federation, 47: 252 (1975). 3. EDEN, G. E. The measurement of turbidity in water. Procedings of the Society of Water Treatment and Examination, 14: 27 (1965). 4. AWW A TASK GROUP. Progress toward a filtrability index test. Journal of the American Water Works Association, 51: 1539 (1959). 5. Standard methods of test for turbidity of water, 01889-71 (1977). In: 1980 Annual book of ASTM standards, Part 31. Philadelphia, American Society for Testing and Materials, 1980, p. 260.
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6. AMERICAN PUBLIC HEALTH AsSOCIATION. Standard methods for the examination of water and wastewater, 14th ed. Washington, DC, American Water Works Association, 1976, p. 131. 7. HACH, C. C. Understanding turbidity measurement. Industrial water engineering, 9: 18 (1972). 8. PACKHAM, R. F. The preparation of turbidity standard. Proceedings of the Society of Water Treatment and Examination, II: 64 (1962). 9. National interim primary drinking water regulations. Washington, DC, US Environmental Protection Agency, 1976, p. 12 (EPA-570/9-76-003). 10. Turbidity. In: Methods for chemical analysis of water and wastes. Washington, DC, US Environmental Protection Agency. 1976, p. 295 (EPA-625-6-74-003a). II. Beard, J. D. & Tanaka, T. S. A comparison of particle counting and nephelometry. Journal of the American Water Works Assocwtion, 59: 533 (1977). 12. NATIONAL RESEARCH CoUNCIL. Dnnking water and health. Washington, DC, National Academy of Sciences, 1977, Chapter IV. 13. NARKIS, N. & REBHUN, M. The mechanism of flocculation processes in the presence of humic substances. Journal of the Amencan Water Works Association, 67: 101 (1975). 14. MAcKENTHUN, K. M. & KEUP, L. E. Biological problems encountered in water supplies. Journal of the American Water Works Assocwtwn, 62: 520 (1970). 15. PEMMANEN, V. Humus fractions and their distribution in some lakes in Finland. In: Povoledo, D. & Golterman, H. L., ed., Humic substances, their structure and function m the biosphere. Wageningen, The Netherlands, Pudoc, 1975, p. 207. 16. AMERICAN PUBLIC HEALTH AssOCIATION. Standard methods for the examination of water and wastewater, 14th ed. Washington, DC, American Water Works Association, 1976, p. 64. 17. ATKINS, P. F. & TOMLINSON, H. D. EvaluatiOn of daily carbon chloroform extracts with CAM. Water sewage works, II 0: 281 (1963). 18. BROCK, T. D. Principles of microbial ecology. New Jersey, Prentice-Hall Inc., 1966, pp. 72-74. 19. STOTZKY, G. Influence of clay minerals on microorganisms. III. Effect of particle size, cation exchange capacity, and surface area on bacteria. Canadian journal of microbiology, 12: 1235 (1966). 20. BERG, G. Removal of viruses from sewage, effluents, and waters. 2. Present and future trends. Bulletin of the World Health Organization, 49: 461 (1973). 21. CLARKE, N. A. ET AL. Human enteric viruses in water: source, survival and removability. In: Proceedings of the International Conference in Water Pollutwn Research, London, 1962. Advances in water pollution research, 2: 523 (1964). 22. FoLIGUET, J. M. & DONCOEUR, F. Elimination des enterovirus au cours du traitement des. eaux d'alimentation par coagulation-floculatian-filtration. Water research, 9: 953 (1975). 23. ROBECK, G. G. ET AL Effectiveness of water treatment processes in virus removal. Journal of the American Water Works Association, 54: 1275 (1962). 24. NEEFE, J. R. ET AL. Inactivation of the virus of infectious hepatitis in drinking water. American journal of public health, 37: 365 (1947). 25. BITTON, G. ET AL. Effect of several clay minerals and humic acid on the survival of Klebsiella aerogenes exposed to ultraviolet irradiation. Applied microbiology, 23: 870 (1972). 26. DENNIS, J. M. 1955-56 Infectious hepatitis epidemic in Delhi, India. Journal of the American Water Works Association, 51: 1288 (1959). 27. SYMONS, J. M. & HoFF, J. C. Rationale for turbidity maximum contaminant level. Presented at 3rd Water Quality Technology Conference, Atlanta. Washington, DC, American Water Works Association, 1975. 28. HuDSON, H. E. High-quality water production and viral disease. Journal of the American Water Works Association, 54: 1265 (1962). 29. TAYLOR, F. B. ET AL. The case for water-borne infectious hepatitis. American journal of public health, 56: 2093 (I 966).
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30. ScHNITZER, M. & KAHN, S. U. Humic substances in the environment. New York, Marcel Dekker Inc., 1972, pp. 204-251. 31. CHAU, Y. K. & LUM-SHUE-CHAN, K. Measurement of complexing capacity of lake waters. In: Povoledo, D. & Golterman, H. L., ed., Humic substances, their structure and function in the biosphere. Wageningen, The Netherlands, Pudoc, 1975, p. II. 32. OLIVER, B. G. Heavy metal levels of Ottawa and Rideau River sediments. Environmental science and technology, 7: 135 (1973). 33. RAMAMOORTHY, S. & RUST, R. R. Mercury sorption and desorption characteristics of some Ottawa River sediments. Canadian journal of earth sciences, 13: 530 (1976). 34. GARDINER, J. The chemistry of cadmium in natural water-1. A study of cadmium complex formation using the cadmium specific-ion electrode. Water research, 8: 23 (1974). 35. KAHN, S. U. Adsorption of 2,4-D from aqueous solution by Fu1vic acid-clay complex. Environmental science and technology, 4: 236 (1974). 36. Public health service drinking water standards, Rockville, MD, US Department of Health, Education and Welfare, 1962, p. 21.
17. ZINC 17.1 General description Zinc is an abundant element and constitutes approximately 0.04 g/kg of the earth's crust (1). The most common zinc mineral is sphalerite (ZnS), which is often associated with the sulfides of other metallic elements, for example, lead, copper, cadmium, and iron (2). The natural zinc content of soils is estimated to be between 1 and 300 mg/kg (3). The atmospheric concentrations of zinc vary considerably depending on such factors as proximity to point sources. In rural locations, zinc concentrations are typically between 10 and 100 ngjm\ while levels in urban areas most commonly fall within the range of 100 to 500 ng/m 3 (4).
The carbonates, oxides, and sulfides of zinc are sparingly soluble in water, while the highly soluble chloride and sulfate salts tend to hydrolyse to form zinc hydroxide and zinc carbonate. As a result the concentration of zinc in natural water is generally low. Adsorption on to sediments further depletes the levels of dissolved zinc (5). 17.2 Occurrence The concentration of zinc in tap-water can be considerably higher than that in surface-water owing to the leaching of zinc from galvanized pipes, brass, and zinc-containing fittings. Zinc concentrations in tapwater generally vary between 0.01 and 1 mgjlitre (6). 17.3 Routes of exposure Zinc is important nutritionally and in the USA comprehensive tables of the zinc content of foods are available (7, 8). Meats and dairy products are richest in zinc, while cereals and nuts are also important sources (8). The zinc content of some important food groups is (9): beef, pork, and lamb milk fish and seafood legumes and wheat leafy vegetables and fruits 20-60 mg/kg 3-5 mgjkg in excess of 15 mg/kg 15-50 mg/kg less than 2 mg/kg (fresh weight)
Food is by far the largest source of zinc for humans. The average daily intake for a "normal" man is reported to be 12 mg (10). The 313
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average daily intake of zinc from drinking-water probably does not exceed 400 /.lg. Air is a negligible source of zinc for man. 17.4 Health aspects Zinc is an essential element for both animals and man and is necessary for the functioning of various enzyme systems, including alkaline phosphatase, carbonic anhydrase, and alcohol dehydrogenase (1 1). More than 70 zinc metallo-enzymes are known (1 2). The recommended dietary intake of zinc, depending upon age and sex, is between 4 and 15 mg/day. Pregnant women and mothers of newborn babies require up to 16 mgjday (1 3). In Egypt and the Islamic Republic of Iran, an endemic zinc deficiency syndrome (among young men) has been reported (14, 15). This syndrome, characterized by retarded growth and other signs of immaturity, including anaemia, is probably caused by low intestinal absorption of zinc. A complete cure is effected by oral administration of large daily doses of zinc as the sulfate (16). In man and animals, zinc absorption is affected by many factors, for example, intake of protein, vitamins, and metals (17). Low zinc intake and low body weight increase zinc absorption, whereas high oral doses of zinc, calcium, and phytate reduce uptake; the fraction of ingested zinc that is absorbed is difficult to determine because zinc also appears to be excreted into the gut (18, 19). However, zinc does not accumulate in tissues and the proportion absorbed is thought to be inversely related to the amount ingested (22). In human serum and plasma, zinc levels are about 1 mg/litre, whereas in whole blood the concentration is about five times higher owing to a high concentration (10 mgjlitre) in the red blood cells (20). The highest body concentration of zinc occurs in the prostate (lOOmg/kg wet weight), but high levels also occur in bone, muscle, liver, and pancreas (21). Zinc may be considered nontoxic. The low toxicity of zinc and efficient homoeostatic control mechanisms make chronic zinc toxicity from drinking-water and dietary sources an unlikely hazard in man. Symptoms of zinc toxicity in humans include vomiting, dehydration, electrolyte imbalance, abdominal pain, nausea, lethargy, dizziness, and lack of muscular coordination (17). Acute renal failure caused by zinc chloride has been reported (23). Daily doses of 150 mg of zinc interfere with copper and iron metabolism because zinc is a metabolic antagonist of both these metals. However, where dietary intake of copper and iron are adequate there is little problem even with high zinc doses. Zinc is also a metabolic antagonist of cadmium. High zinc intakes may, therefore, be expected to afford some protection against the toxic effects of cadmium exposure from the environment (24). Taste threshold tests indicate that 5% of a population distinguished between zinc-free water and that containing zinc at a level of 4.3 mgjlitre
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315
(as zinc sulfate) (25). The detection levels with other zinc salts were somewhat higher. 17.5 Other aspects Zinc imparts to water an undesirable astringent taste; in addition, water contammg zinc at concentrations in excess of 5.0 mg/litre may appear opalescent and develop a greasy film on boiling. This value is recommended as a guideline value. To avoid any of the problems mentioned, however, the level of zinc in water should be kept well below this value.
REFERENCES I. BROWING, E. Tox1city of industrial metals. 2nd ed. London, Butterworths, 1969, p. 348. 2. Quality cntena for water. Washington, DC, US Environmental Protection Agency, 1976, p. 481. 3. LEVINSON, A. A. IntroductiOn to exploration geochemistry. Calgary, Apphed Publishing Co., 1974, p. 44. 4. NRIAGU, J. 0., ED. Zinc in the environment. Part I: ecological cycling. New York, John Wiley & Sons, 1980. 5. HEM, J. D. Zinc. In: Study and interpretation of the chemical characteristics of natural water. Washington, DC, US Geological Survey, 1970, p. 125 (Water-Supply Paper 1473). 6. ZOETEMAN, B. C. J. & BRINKMAN F. J. J. Human intake of minerals from drinkingwater lD the European communities. In Hardness of dnnkmg water and pubbc health. Proceedings of the European Scientific Colloquium, Luxembourg, 1975. Oxford, Pergamon Press, 1976, p. 173. 7. MuRPHY, E. W. ET AL ProvisiOnal tables on the zinc content of foods. Journal of the American Dietetic Association, 66: 345 (1975). 8. FREELAND, J. H. & CouSINS, R. J. Zinc content of selected foods. Journal of the American Dietetic AssociatiOn, 68: 526 (1976). 9. WHO Technical Report Series, No. 532, 1973 (Trace elements in human nutrition; report of a WHO Expert Committee). 10. WARREN, H. V. Some trace element concentrations in various environments. In: Howe, G. M. & Loraine, J. A., ed. EnVIronmental medicme. London, William Heinemann Medical Books Ltd., 1973, p. 9. II. PARISIC, A. F. & VALLEE, B. L. Zinc metalloenzymes: characteristics and significance in biology and medicine. American journal of clinical nutrition, 22: 1222 (1969). 12. Symposium on trace elements. Medical clinics of North America, 60: 4 (1976). 13. FooD AND NuTRITION BOARD. Recommended d1etary allowances. 8th revised version. Washington, DC, National Academy of Sciences, 1974, pp. 99-101. 14. PRASAD, A. S. ET AL. Syndrome of iron, anemia, hepatosplenomegaly, hypogonadism, dwarfism and geophagia. Amencan journal of medicine, 31: 532 (1961). 15. HALSTED, J. A. ET AL Zinc deficiency in man. The Shiraz experiment. American journal of medicine, 53: 277 (1972). 16. MICHAELSON, G. Zinc therapy in acrodermatitis enteropathica. Acta dermatologica, 54: 377 (1974). 17. PRASAD, A. S. & 0BERLEAS, D., ed. Trace elements in human health and disease. Vol. 1. zinc and copper. New York, Academic Press, 1976, p. 470. 18. BECKER, W. M. & HoEKSTRA, W. G. The intestinal absorption of zinc. In: Skorylla, S.C. & Valdron-Edwards, D., ed. Intestinal absorption of meta/10ns, trace elements and radionuclides. New York, Pergamon Press, p. 229.
316
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AESTHETIC CONSTITUENTS AND CHARACTERISTICS
19. HONSTEAD, J. F. & BRADY, D. N. The uptake and retention of 32P and 65Zn from the consumption of Columbia River fish. Health physics, 13: 455 (1967). 20. SUNDERMAN, F. W. In: Goyer R. A. & Mehlman. M. A., ed. Advances in modern toxicology. Washington, DC, Hemisphere Publishing Corporation, 1976. 21. HALSTED, J. A. ET AL Journal of nutrition, 104: 345 (1974). 22. HETH, D. A. ET AL. Effect of calcium, phosphorus and zinc on zinc-65 absorption and turnover in rats fed semipurified diets. Journal of nutrition, 88: 331 (1966). 23. CSATA, S. ET AL. Akute Niereninsuffizienz als folge einer. Zinkchloridvergiftung. Zeitschrift fiir Urologie, 61: 327 (1968). 24. UNDERWOOD, E. J. Trace elements in human and animal nutritiOn, 4th ed. New York, Academic Press, 1977, p. 545. 25. CoHEN, J. M. ET AL Taste threshold concentrations of metals in drinking water. Journal of the American Water Works Association, 52: 660 (1960).
PART VI. RADIOACTIVE MATERIALS
1. INTRODUCTION The levels of radioactivity in drinking-water recommended in the standards published by WHO in 1970 and 1971 {1, 2) were based on data available from the International Commission on Radiological Protection (ICRP) over the period 1959-1966 inclusive. However, since then additional information has become available (3-5) and has been taken into consideration in the preparation of the present guidelines. The recommended guideline values of 0.1 Bq/lltre for gross alpha activity and I Bq/litre for gross beta activity are based upon an adult drinking-water intake of 2 litres per day. It is recommended that levels of activity exceeding these values be reported to the competent authorities to determine what action, if any, is required. Procedures for measuring gross alpha and beta activities as well as individual radionuclides are described elsewhere (6). Radioactive materials are introduced into the environment from a number of sources-naturally occurring and man-made. The naturally occurring sources include those substances produced by cosmic rays, which may find their way to water courses with rainfall and runoff, and those present in the rocks and soil, such as uranium-238 and its daughters radium-226 and radon-222. The man-made radionuclides are those resulting from fallout from nuclear tests, nuclear power production, and medical and other uses of radioactive materials. The dose of natural radiation that a person receives depends upon a number of factors, such as the height above sea level at which he lives, the amount and type of radioactive nuclides in the soil, and the amount he takes into his body in air, food, and water. The growth in nuclear applications has led to more widespread environmental releases, which in turn add to the amount of radioactive substances in surface- and groundwater and could have a direct effect on radioactivity levels in water sources used for public water supply.
319
320
VI.
RADIOACTIVE MATERIALS
2. BASIC CONSIDERATIONS In assessing radiation exposure, the recommendations of IAEA (7) and ICRP (4, 5) were followed. ICRP has recently introduced the concept of "detriment" in order to identify and, if possible, quantify all deleterious effects. "Detriment" is defined as the mathematical "expectation" of harm i.ncurred from an exposure to radiation, taking into account not only the probability of each type of deleterious effect, but also the severity of the effect. The ICRP considers that knowledge of the absorbed dose is insufficient by itself to predict either the severity or the probability of deleterious effects on health resulting from irradiation under unspecified conditions. The dose-equivalent limits for radiation protection recommended by the ICRP do not apply to or include natural radiation exposure, except in so far as certain components are augmented by man's activities. Clearly, however, the ICRP recognizes that there is no sharp dividing line between levels of natural radiation that can be regarded as "normal" and those that are elevated owing to human activities or choice of environment. The ICRP has de-veloped a system of dose limitation that requires the following: "(a) no practice shall be adopted unless its introduction produces a positive net benefit; (b) all exposures shall be kept as low as reasonably achievable, economic and social factors being taken into account; and (c) the dose equivalent to individuals shall not exceed the limits recommended for the appropriate circumstances by the Commission". 2.1 Dose-response relationship
The detrimental health effects of exposure to radiation are either somatic, i.e., those that become manifest in the exposed individual, or hereditary, i.e., those that affect the exposed individual's descendants. For some somatic effects, such as carcinogenesis, and for the hereditary effects at dose levels involved in radiation protection, the probability of an effect occurring, rather than its severity, is regarded as a function of dose, without threshold (stochastic effects). For other somatic effects the severity varies with the dose and the effect appears above a threshold dose (non-stochastic effects). The aim of radiation protection is to prevent detrimental nonstochastic effects and to limit the probability of stochastic effects to a level deemed acceptable. The latter aim will be achieved by the application of the system of dose limitation presented in section 2.2. The prevention of non-stochastic effects is achieved by setting up doseequivalent limits at values sufficiently low that the threshold dose would not be reached even if the exposure persisted for the whole lifespan. The level of radioactive material in drinking-water required to reduce the
2.
BASIC CONSIDERATIONS
321
incidence of stochastic effects to an acceptable level will automatically exclude the possibility of non-stochastic effects. The dose-equivalent represents a method of quantifying dose that allows a better correlation between radiation exposure and the induced deleterious effects, more particularly the delayed stochastic effects. A basic assumption of the ICRP is that, in the range of exposure concerned, the probability of a stochastic effect occurring is proportional to the dose received. However, different tissues in the body have different sensitivities to radiation and hence the ICRP has introduced doseequivalent weighting factors to provide measures of equal risk. The sum of the weighted dose-equivalent factors for each tissue gives a measure of the total risk and is referred to as the effective dose-equivalent. Moreover, in the case of long-lived radionuclides metabolized in such a way that they remain in the body for appreciable time periods, the resulting exposure may extend over many years. The committed effective dose-equivalent (HE, 50) is defined as the total effective dose-equivalent incurred in the 50 years following intake of the nuclide. It is this measure of exposure that is relevant to the present discussion; in what follows, the term "dose" may be used for brevity. 2.2 Dose-equivalent limits The Commission's recommended dose limits apply to two categories of exposure-occupational and general. The limitation of the doseequivalent refers to the sum of the annual dose-equivalents contributed by external sources and committed dose-equivalents from radioactive materials taken into the body during any year. Dose-equivalent limits established for occupational exposure are regarded as upper limits. Limitation of the dose-equivalent for members of the public is a more theoretical concept, mainly intended to ensure that it is unlikely that an individual will receive more than the specified dose-equivalent. Usually, the effectiveness is checked using sampling procedures and statistical calculations and by control of sources from which exposure is expected to arise. The basic safety standards for radiation protection, based on ICRP's recommendations have set the effective dose-equivalent limit for the individual member of the public at 5 mSv in a year as applied to the average dose-equivalent in the critical group.a However, in cases where the doses to the same individuals approach this limit over many years, it would be prudent to take measures to restrict their lifetime doseequivalent to a value corresponding to an average annual effective doseequivalent of 1 mSv.
a The ICRP define "critical groups'' as groups of the population with characteristiCS causing them to be exposed at a higher level than the rest of the exposed population from a given practice. These groups may be used as a measure of the upper hmlt of the individual doses from a proposed practice.
322
VI.
RADIOACTIVE MATERIALS
2.3 Implications for drinking-water quality To apply the above principles to drinking-water, a measure of the potential radiation exposure from the water is necessary. While it may be possible to make some estimate on the basis of information already available-authorized discharges of man-made radioactive materials from nuclear installations, geologiCal data on natural radioactivity, etc.confirmation of the level of radioactivity requires direct measurement. Where it has been shown that the level is acceptably low, the necessary frequency of further measurements may be decided following a review of the particular circumstances applying and consultation with the competent authorities. In the case of new water supplies, information concerning the activity levels in the raw water would be required. Levels in the final water after treatment should be checked. If appropriate, the levels found in drinking-water should be related (by a competent authority) to the total exposure from all sources to which the population served by the water treatment plant is subjected.
3. SOURCES OF RADIATION EXPOSURE The basic criterion for estimating the level of exposure to which individuals are subjected is established on the dose limitation system recommended by the ICRP (4, 5). Exposure may result from naturally occurring radionuclides at natural levels or at levels augmented by man's activities, for example, the use of phosphate fertilizers and discharge of mine waters, and from artificial radionuclides introduced into the environment, such as fallout from nuclear tests, releases from nuclear power facilities, and discharges as a result of the use of radionuclides in medicine, industry, and research (8). Exposure levels from natural and man-made radioactivity are, in fact, regularly evaluated on a global basis, in so far as is possible, by the United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR), whose most recent report was published in 1977 (8). An examination of the data contained therein shows that drinking-water is a relatively minor constituent of total radiation exposure.
4.
GUIDELINE VALUE FOR GROSS ALPHA AND GROSS BETA ACTIVITY
323
4. GUIDELINE VALUE FOR GROSS ALPHA AND GROSS BETA ACTIVITY The radionuclides of interest were identified on the basis of those present in the natural environment as well as those resulting from man's activities. They are identified basically as the alpha and beta emitters, some of which have radioactive daughters. Radium-226 is typical of the naturally occurring alpha-emitting radionuclides of interest and strontium-90 is among the man-made beta emitters. However, it is not always necessary to identify specific radionuclides present when the concentrations are low. In such cases, measurements of gross alpha and gross beta activity may serve to demonstrate that the radiotoxicity level is acceptable. Gross alpha and gross beta activity measurements are of particular interest for routine monitoring purposes. Ideally, to use the gross activity screening procedure, the reference levels for alpha and beta activity will require to meet two criteria: (a) they should be such as to ensure that, irrespective of the individual nuclides contributing to the gross activity, the associated exposure will be low enough not to necessitate further detailed analyses and consideration; (b) they should be sufficiently high to ensure that the vast majority of drinking-water supplies satisfy such reference levels and hence the need for detailed analyses can be avoided. In 1979, reference levels of 0.1 and 0.8 Bq were recommended (3) for gross alpha and gross beta activity, respectively. These levels appear to satisfy the second criterion. However, it is necessary to ensure that with the advancement of knowledge, these levels continue to satisfy the first criterion. To examine this aspect, the nuclides possibly contributing must be considered individually. Table 5 sets out the nuclides concerned and indicates, for each individual nuclide, the potential exposure from drinking two litres of water per day with contamination levels of 0.1 Bq per litre for alpha emitters and I Bq per litre for beta emitters. Accurate calculations of the dose actually received by a person drinking water containing radon have not yet received general acceptance, although UNSCEAR and ICRP are at present reviewing this topic. However, it is known that the drinking-water pathway is of minor significance. In the measurement of gross alpha and gross beta activity, it is understood that both radon and tritium (a nuclide of low radiotoxicity) are excluded. Where locally elevated levels of either of these nuclides are suspected, the competent authorities should be consulted. It is accepted that some of the transuranium isotopes are more toxic than the nuclides listed in the table; however, the occurrence of transuranium isotopes in drinking-water at appreciable concentrations is
324
VI.
RADIOACTIVE MATERIALS
rare. Where their presence is suspected, guidance should be sought from the competent authorities as to any special precautions that have to be taken. Table 5. Potential exposures to various alpha and beta emitters from drinking two litres of water per day for one year HE,50 Nuclide
HE,50[Bq (Sv)• 4.36 7.63 3.05 7.4 7.07 6.32 x1o- 7 x10-B x1o- 7 x1o- 7 x10-B x10-B
from one year's ingestion (mSv)
Alpha emitters 21op 0 224Ra 22sRa 2J2Th 2J4U 2JBU
73 Bq ( = 0.1 Bqflitre)
0.032 0.006 0.022 0.054 0.005 0.005 730 Bq ( = 1 Bqflitre)
Beta emitters so co B9Sr gosr 1291 1311 1J4Cs 1J7Cs 21opb 22BRa
6.97 x1o- 9 2.17 x1o- 9 3.6 x1o-B 7.4 x10-B 1.4 x10-B 1.98x1o- 8 1.36x1o- 8 1.36x1o- 6 3.3 x1o- 7
0.005 0.002 0.026 0.054 0.010 0.014 0.010 0.993 0.214
• Data from the supplements to ICRP Publication 30 (5).
Of the alpha-emitting nuclides included in the table, thorium-232 is unlikely to be a major contributor in its own right to gross alpha activity. Hence, it would appear that attributing the alpha activity entirely to radium-226 is a conservative approach. Considering the beta emitters, iodine-129 can be eliminated with the same proviso as for the transuranium isotopes. Radium-228 and more particularly lead-210 are the most toxic of the nuclides listed. However, these nuclides will normally represent only a small fraction of the gross beta activity, except in cases when the radon concentration in the water is high. Hence, assigning the beta activity entirely to strontium-90 would appear to be a conservative assumption, in the absence of high radon concentrations. It can be seen that attributing 0.1 Bq of alpha activity per litre to radium-226 and 1 Bq of beta activity per litre to strontium-90 would imply an exposure of 0.048 mSv per year for a daily drinking-water intake of 2 litres. The actual exposure associated with these concentrations will not exceed 0.048 mSv per year, and it is highly improbable that in practice even this dose level would be incurred.
5.
RADON
325
According to ICRP data, this dose corresponds to a total risk in the range I0- 7 to I0- 6 per year, an order of magnitude less than what would be "likely to be acceptable to any individual member of the public" (4). This order of magnitude fully allows for the fact that drinking-water represents only one part of the general exposure. Hence, levels of I Bqjlitre for gross beta activity and 0.1 Bq/litre for gross alpha activity, as quoted above, are recommended as the reference levels for screening purposes.
5. RADON Data from several countries show radon activity up to I 0 3 Bqjlitre in groundwater sources (deep wells) used as drinking-water supplies by some communities. Since the radon in the water is easily lost during handling, it is difficult to assess what quantity is ingested. Therefore, it is not possible to make accurate calculations of the dose actually received by a consumer drinking water containing radon, although there have been many attempts to do so (3). It should be noted, however, that, as far as radon is concerned, the health risk from inhalation of the atmosphere in the room where water is drawn from a tap is greater than the risk due to ingestion of the water (8).
REFERENCES I. European standards for drinking-water, 2nd ed. Geneva, World Health Organization, 1970. 2. International standards for drinking-water, 3rd ed. Geneva, World Health Organization, 1971. 3. Radiological exammat10n of drinking-water. Report of a WHO Working Group. Copenhagen, WHO Regional Office for Europe, 1979 (EURO Reports and Studies No. 17). 4. INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION. Recommendations of the International Commission on Radiological Protection. Annals of the ICRP, 1 (3): 1-53 (1977) (ICRP Publication 26). 5. INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION. Limits for intakes of radionuclides by workers. Annals of the ICRP, 2-8, (1979-1982) (ICRP Publication 30 and supplements). 6. MITCHELL, N. T. Radiological examination. In: Suess, M. J., ed. Examination of water for pollution control, vol. 2, Oxford, Pergamon Press, 1982, chapter 5. 7. INTERNATIONAL ATOMIC ENERGY AGENCY. Basic safety standards for radiation protection. Vienna, IAEA, 1982 (Safey Series No. 9). 8. UNITED NATIONS SciENTIFIC COMMITTEE ON THE EFFECTS OF ATOMIC RADIATION. Sources and effects of ionising radiation. New York, United Nations, 1977.
INDEX
INDEX Aesthetic quality of drinking-water, 249-318 see also Colour; Odour; Taste; Turbidity Aldrin, acceptable daily intake, 197 health effects, 196-197 metabolism, 195-196 routes of exposure, 194-195 Algae, 53-56 Alkanes, chlorinated, 155-161 Alkylbenzenes see Benzene Alpha emitters, 323-325 exposure from water, 324 Aluminium, health effects, 250-251 occurrence, 249-250 recommended guideline value, 249 routes of exposure, 250 Amoebiasis, 40-43 Ancylostoma, 48-50 Argyria, 143 Arsenic, health effects, 65-66 metabolism, 65 occurrence, 63 routes of exposure, 63-64 Asbestos, absorption and distribution, 70-71 health effects, 71-73 occurrence, 68-69 routes of exposure, 69-70 Asbestosis, 72-73 Ascaris lumbricoides, 50 Bacterial pathogens, waterborne, detection, 27-29 transmission, 3-4 see also individual bacteria Bacteriological monitoring, 4-9, 1529 see also Bacteriological samples; individual bacteria Bacteriological samples, bottles for, 12 collection, 13-14 concentration, 27 329
disinfectant neutralization, 13 transport and storage of, 15 Balantidiasis, 41, 43 Balantidium coli, 39-40, 42-44 Barium, health effects, 78 metabolism, 78 occurrence, 76 routes of exposure, 76-77 Benzene (and lower alkylbenzenes), health effects, 231-232 metabolism, 231 occurrence, 229 recommended guideline value, 232 routes of exposure, 229-231 Benzo[a]pyrene, health effects, 184 metabolism, 184 occurrence, 182 recommended guideline value, 185186 routes of exposure, 182-183 Beryllium, health effects, 81-82 metabolism, 81-82 occurrence, 80 routes of exposure, 80-81 Beta emitters, 323-325 exposure from water, 324 Biomass, 55 Bladder cancer, association with trihalomethanes, 244 Bromodichloromethane, 240-241 concentration in water, 241 see also Chloroform; Trihalomethanes Bromoform, 240-241 concentration in water, 241 see also Chloroform; Trihalomethanes Cadmium, health effects, 88-89 metabolism, 87-88 occurrence, 84 routes of exposure, 85-87 Calcium see Hardness
330
INDEX
Campylobacter fetus, 28-29 Carbon tetrachloride, health effects, 157-158 metabolism, 156-157 occurrence, 155-156 tentative guideline value, 158 Carcinogenicity, of arsenic, 65-66 of asbestos, 71-73 of benzenes, 227, 231 of beryllium, 82 of cadmium, 89 of chlorinated alkanes, 157-158, 160 of chlorinated ethenes, 164-165, 168, 173-174, 178-179 of chloroform, 243-244 of chromium, 95 of fluoride, I 04 of nickel, 126 of nitrosamines, 133 of pesticides, 196, 20 I, 206, 211, 215, 220 of polynuclear aromatic hydrocarbons, 184-185 of selenium, 139 of trichlorophenol, 237 Carcinoma, hepatocellular, see Hepatocellular carcinoma Cardiovascular disease, and water hardness, 109 associated with barium, 78 Caries, dental, see Dental caries Central nervous system depression, by benzene, 231 by chloroform, 243 by copper, 263 by tetrachloroethene, 178 Chlordane, health effects, 200-201 metabolism, 199-200 occupational exposure to, 199, 201 occurrence, 197 recommended guideline value, 201 routes of exposure, 198-199 Chloride, health effects, 253-254 occurrence, 253 recommended guideline value, 254 routes of exposure, 253 see also Salt Chlorinated alkanes, 155-161 Chlorinated ethenes, 162-179 Chlorinated phenols, see Chlorophenols
Chlorination, effect of temperature, 301, 302 effect of turbidity, 11-12, 309-310 effect on cyanide, 97 formation of trihalomethanes, 54, 240, 259-260, 302 see also Disinfection Chlorine residual, I 0-11 neutralization of, 13 Chlorobenzene (monochlorobenzene), health effects, 222, 223 concentration in water, 222 metabolism, 223 occurrence, 221- 222 recommended guideline value, 223224 routes of exposure, 222-223 Chlorobenzenes, 221-228 Chlorodibromomethane, 240-241 concentration in water, 241 see also Chloroform; Trihalomethanes Chloroform, concentration in water, 241 health effects, 242-245 metabolism, 242 occurrence, 240-241 recommended guideline value, 244245 routes of exposure, 241-242 Chlorophenols, health effects, 234 occurrence, 233-234, 235 recommended guideline value, 233 see also Pentachlorophenol; 2,4,6Trichlorophenol Chromium, health effects, 95-96 metabolism, 94-95 occurrence, 91-92 routes of exposure, 92-94 Clostridia perfringens (C. welchii), 7, 24-25 Clostridia, sulfite-reducing, definition, 24 detection by membrane-filtration, 24-25 detection by multiple-tube method, 24 Coliform organisms, 5-7 definition, 16 detection of, 16-22 apparatus, 20
INDEX
331
Coliform organisms (continued) confirmatory tests, 18, 21 culture media, 18, 21 differentiation, 22 membrane-filtration, 19-22 multiple-tube method, 17-19 see also Bacteriological monitoring; individual coliforms Colony counts, 8-9, 20-21, 26 Colour, 256-260 caused by free-living organisms, 55 effect of pH, 283 health aspects, 258 measurement of, 256-257 occurrence, 257 recommended guideline value, 256 removal, 249, 257 source, 256 Copper, health effects, 262-263 occurrence, 262 recommended guideline value, 263 Culture media, 18, 21, 23, 24-25, 26 Cyanides, health effects, 99 metabolism, 98-99 occurrence, 97 recommended guideline value, 99 routes of exposure, 98 2,4-D, health effects, 219--220 metabolism, 219 occurrence, 218- 219 recommended guideline value, 220 routes of exposure, 219 DDE, see DDT DDT (total isomers), health effects, 193-194 metabolism, 192-193 occurrence, 191 routes of exposure, 191-192 Dental caries, association with selenium, 138, 139 prevention by fluorides, 103 Dibromochloromethane, see Chlorodibromomethane Dichlorobenzenes, health effects, 225227 occurrence, 221-222, 224-225 routes of exposure, 224-225 recommended guideline values, 227 Dichlorodiphenyltrichloroethane, see DDT
1,2-Dichloroethane, health effects, 160-161 metabolism, 160 recommended guideline value, 161 routes of exposure, 159-160 I ,1-Dichloroethene, health effects, 167-168 metabolism, 166-167 recommended guideline value, 168169 routes of exposure, 166 I, I '-(2,2-Dichloroethenylidene)bis[4chlorobenzene] (DDE), see DDT Dichlorophenols, see Chlorophenols 2,4-Dichlorophenoxyacetic acid, see 2,4-D Dieldrin, acceptable daily intake, 197 health effects, 196 metabolism, 195-196 occurrence, 194-195 routes of exposure, 195 Disinfectants, residuals, 10-11 neutralization of, 13 Disinfection, effect of turbidity on, 11-12 efficiency, I 0 see also Chlorination Dissolved oxygen, 279-280 Dose-equivalent limits, radiation, 320, 321 Dose-response relationship, radiation, 320 Dracontiasis (dracunculosis), 48 Dracunculosis, see Dracontiasis Dracunculus, 46-49 Dysentery, amoebic, 40-43 Endotoxins, detection of, 55-56 Entamoeba histolytica, 39-45 Escherichia cob, 5-7, 16, 18, 19, 22,28 Ethenes, chlorinated, 162-179 see also I, 1-Dichloroethene; Tetrachloroethene; Trichloroethene; Vinyl chloride Faecal pollution, 3-8, 16, 17, 32-33 Faecal streptococci, 5, 7, 22-24 definition, 22 detection, by membrane-filtration, 23 by multiple-tube method, 22-23
332
INDEX
Faecal streptococci (continued) differentiation, 23-24 Faecal (thermotolerant) coliforms, 5, 6-7 Fasciolids, 51 Fluoride, health effects, 103- I 04 metabolism, I 02 routes of exposure, I 00-102 Fluorosis, 103-104 Free-living organisms, 53-56 Gamma-HCH, see Lindane Gastroenteritis, 33, 54 Giardia spp., 6, 39, 40, 41, 42, 43, 4445 Giardiasis, 41, 42, 43 Guideline values, see under individual contaminants; individual organisms Guinea-worm infection, see Dracontiasis Haemochromatosis, 273 Hardness, health effects, 108-110, 265 causes, 106, 264 occurrence, I 06- I 08, 265 pH and, 282 see also under Cardiovascular disease HCB, see Hexachlorobenzene Helminths, infections caused by, 48. 50, 52 monitoring, 48, 50, 51 recommended guideline values, 4849, 51, 52 species, 47 transmission, 46-48, 49-50, 51-52 see also individual helminths Hepatitis, 33 Hepatocellular carcinoma, associated with particular contaminants, 173, 179, 201, 211, 237, 244 Heptachlor and heptachlor epoxide, health effects, 210-211 metabolism, 209-210 occurrence, 208 recommended guideline value, 211 routes of exposure, 208-209 Hexachlorobenzene, health effects, 204-206
metabolism, 204 occurrence, 203 recommended guideline value, 206 routes of exposure, 203-204 Humic material, 256, 257-260, 308310 Hydrocarbons, polynuclear aromatic, see Polynuclear aromatic hydrocarbons Hydrogen sulfide, health effects, 269270 occurrence, 268-269 pH and, 283 routes of exposure, 269 Hypernatraemia, 149-150 Hypertension, associated with cadmium intake, 88, 89 associated with sodium intake, 150151 Indicator organisms, 4-9, 44-45 see also Coliform organisms Invertebrates, 53-56 Iron, health effects, 273 occurrence, 272 oxygen, dissolved, and, 279 routes of exposure, 272 "Iron bacteria", 257, 273, 282 Jackson turbidity unit (JTU), 307-308 Laboratory facilities and safety, 16 Lead, health effects, 116-118 metabolism, 115-116 occurrence, 111 routes of exposure, 112-115 Lindane, acceptable daily intake, 214 health effects, 214-215 metabolism, 213 occurrence, 212 recommended guideline value, 215 routes of exposure, 212-213 Macroinvertebrates, 53-56 Magnesium, see Hardness Manganese, health effects, 276-277 occurrence, 275 recommended guideline value, 277 routes of exposure, 275-276 Membrane-filtration method, 19-22, 23, 24-25, 26
INDEX
333
Meningoencephalitis, 40, 41, 43-44 Mercury, health effects, 122-123 metabolism, 122 occurrence, 120-121 routes of exposure, 121-122 Methaemoglobinaemia, 132-133 Methoxychlor, health effects, 218 metabolism, 217-218 occurrence, 217 recommended guideline value, 218 routes of exposure, 217 Microbiological aspects, 3-36 see also individual bacteria; Viruses Minimum infective dose, 4, 33-34, 44 Monitoring, of bacteria, 4-5, 8, 15-29 of free-living organisms, 55-56 of helminths, 48, 50, 51 of protozoa, 44 of viruses, 34-35 Monochlorobenzene, see Chlorobenzene Mottling of teeth, I 03 Multiple-tube method, 17-19, 22-23, 24, 25-26 Mutagenicity, of chlorinated ethenes, 165, 167, 172, 178 of I ,2-dichloroethane, 160-161 of nickel, 126 of pesticides, 200-201, 210, 214-215 of polynuclear aromatic hydrocarbons, 184 of trichlorophenol, 237 Naegleria, 39, 40, 43-44 Nephelometric turbidity unit (NTU), 12, 307-308 Nickel, health effects, 126-127 metabolism, 126 occurrence, 124 routes of exposure, 125-126 Nitrate and nitrite, health effects, 132133 metabolism, 131-132 occurrence, 128-129 routes of exposure, 129-131 see also Nitrosamines Nitrite, see Nitrate and nitrite Nitrosamines, 133 Nuisance organisms, 9, 298
Odour, causes, 55, 233, 298-299 dissolved oxygen and, 279 pH and, 283 temperature and, 301 testing for, 297, 299 Odour threshold number, 297, 299 Organisms, free-living, 53-56 Organoleptic properties, see Aesthetic quality of drinking-water Oxygen, dissolved, 279-280 Pathogens, bacterial, see Bacterial pathogens Pentachlorophenol, health effects, 238 metabolism, 238 occurrence, 237 recommended guideline value, 238 routes of exposure, 238 Perchloroethylene, see Tetrachloroethene Pesticides, 190-220 see also individual pesticides pH, 281-284 Phenol, health effects, 234-235 occurrence, 233-234, 235 recommended guideline value, 233 Phenols, chlorinated, see Chlorophenols Physical aspects, see Colour; Temperature; Total dissolved solids; Turbidity Pipes, asbestos-cement, 72 cadmium, 84, 85 corrosion of, 257, 281-283, 302-303 encrustation of, 303 lead, Ill- 112 nuisance organisms in, 9 poly(vinyl chloride), 162 zinc, 313 Plankton, 53-56 Pollution, faecal, 3-8, 32-33 Polynuclear aromatic hydrocarbons (P AH), health effects, 184-185 metabolism, 184 occurrence, 182 recommended guideline value, 185186 routes of exposure, 182-183 Poly(vinyl chloride), see Vinyl chloride Porphyria cutanea tarda, 205
334
INDEX
Presumptive coliform test, 17-19 Protozoa, infections caused by, 43-44 monitoring, 44 transmission, 39-43 see also individual organisms Pseudomonas aeruginosa, 8, 25-26 Quality of water and indicator organisms, 8-9 Radiation exposure, dose-equivalent limits, 320, 321 dose-response relationship, 320 from water, 324 recommended guideline values, 320, 321-322, 323, 325 stochastic effects, 320-321 Radioisotopes, 323-325 Radionuclides, 323-325 Radon, 325 Salmonella, 27 Salt, 145, 253-254, 304 see also Sodium Samples, see Bacteriological samples; Virological samples Schistosoma, 48-50 Selenium, health effects, 137-139 metabolism, 136-137 occurrence, 135 routes of exposure, 135-136 Shigella, 27-28 Silver, health effects, 143-144 metabolism, 143 occurrence, 141 routes of exposure, 141-143 Sodium, health effects, 149-151, 287288 metabolism, 148-149 occurrence, 145, 286-287 recommended guideline value, 151, 288 routes of exposure, 146-148, 287 see also Salt Sodium chloride, see Salt Softness, of water, 112, 264, 266 see also Hardness Solids, total dissolved, see Total dissolved solids Spirometra, 46-48
Stochastic effects, of radiation, 320321 Streptococci, faecal, see Faecal streptococci Sulfate, health effects, 291 occurrence, 290 recommended guideline value, 291 routes of exposure, 290-291 Sulfide, see Hydrogen sulfide see Sulfite-reducing clostridia, Clostridia, sulfite-reducing Taste, causes, 55, 259, 294-295 testing for, 293-294, 296 see also Taste threshold Taste threshold, of chloride, 254 of hydrogen sulfide, 270 of manganese, 277 of residual chlorine, 294-295 of sodium, 288 of sulfate, 291 of zinc, 314-315 Temperature, 301-303 Teratogenicity, of pesticides, 205, 210 of polynuclear aromatic hydrocarbons, 184 of trichloroethene, 172-173 of vinyl chloride, 165 Tetrachloroethene, health effects, 178179 metabolism, 177-178 occurrence, 176 routes of exposure, 176-177 tentative guideline values, 179 Tetrachlorophenol, see Chlorophenols Tobacco, inorganic constituents of, 64, 77, 86, 93, 107, 113, 125, 130 Toluene, health effects, 231-232 metabolism, 231 occurrence, 229 routes of exposure, 229-231 Total coliforms, see Coliform organisms Total dissolved solids, occurrence, 304-305 . recommended guideline value, 305 Trichlorobenzene, 221-222 Trichloroethene, health effects, 172174 metabolism, 171 occurrence, 170
INDEX
335 measurement, 307-310 recommended guideline value, 12, 310
Trichloroethene (continued) routes of exposure, 170 tentative guideline value, 174 1,1 '-(2,2,2-Trichloroethylidene) bis [4chlorobenzene], see DDT 2,4,6-Trichlorophenol, health effects, 237 metabolism, 237 occurrence, 234, 235, 236 recommended guideline value, 237 routes of exposure, 236 Trichlorophenols, see Chlorophenols; 2,4,6-Trichlorophenol Trichuris trichiura, 50 Trihalomethanes, health effects, 242245 metabolism, 242 occurrence, 240-241 recommended guideline value, 244245 routes of exposure, 241-242 see also individual trihalomethanes; see also under Chlorination True colour unit (TCU), 256, 257 Turbidity, causes, 55, 307-310 effect on disinfection, 11-12, 309310
Vibrios, cholera and non-cholera, 29 Vinyl chloride, health effects, 164-165 metabolism, 164 occurrence, 162 routes of exposure, 163 Viral infections, 33-34 Virological samples, 34-35 Viruses, monitoring, 34-35 recommended guideline values, 3334 temperature and, 302 transmission, 32-33 Waterborne bacterial pathogens, 3-4 Yersinia enterocolitica, 28
Zinc, health effects, 314 occurrence, 313 recommended guideline value, 315 routes of exposure, 313-314
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饮用水水质准则 第二卷
卫生基准及其补充资科
世界卫生组织
编
秦伍慧王有森吴联熙
译
人民卫生出版社
⑩
世界卫生组织委托中华人民共和国卫生部由人民
卫生出版祉出版本书中文版
Guidelines For Drinking-Water Quality Volume 2 Health Criteria and Other Supporting Informa tion Wor ld Health Organization Geneva 1984
饮用水水质准则 第二卷
卫生基准及其补充资料 秦怪慧 等译
人民卫生出版社出版 (北京市崇文区天坛西里 10 号)
人民卫生出版社印刷厂印刷
新华书店北京发行所发行
787 x 1092 毫米
32 开本
15 印张
4 插页
324 千字
1986 年 12 月第 1 版
1986 年 12 月第 1 版第 1 'l1.:印刷
目录 序言………………. .
第-章微生物学指挥 1.饮用水的细菌学质量……………………… .2 1. 1 介水传播的病菌…………………………….
2 3
1. 2 利用指示菌的理由…………………………..
1. 3 障碍性生物……………………………… ..9 1. 4 消毒…………………………………….
.10
1. 5 细菌检验用水样的采集、贮存和运输…………… '13 1. 6 检出与计数污染指示菌的建议方法……………… 16
参考文献……………………………………. '32 2. 饮用水的病毒学质量… 2.1 概述…………………………………….
'39
2.2 接触途径………………………………… '39 2.3 对健康的影响……………………………… 40 2.4 建议的依据……………………………….
'41
2.5 病毒检验方法……………………………… 42 2.6 对阳性结果的解释和评价……………………. '43
参考文献……………………………………. '43
第二章生物学指挥
1.原生动物
2.
.蠕虫………. .
Cl l
2.1 第一组(龙线虫、叠宫缘虫〉…………………… 54 2.2 第二组(血吸虫、钩虫、板口线虫)……………… 58
2.3 第三组(蚓虫、鞭虫、类圆线虫、烧虫、片吸虫、 膜壳缘虫、棘球拗〉………………………… 60
3.
自由生活生物……………………………… m
参考文献……………………………………. .67 第三章 与健康有关的无机成分
1. 2.
~申….
.
石棉………………………
3. 4. 5.
顿…· 镀……· 铺……·
6.
恪……"
7.
氟化物…
8. 9. 10.
氟化物…· 硬度…· 铅……………………………………… 152
1 1.来 12. 镇………………..........................
.170
13.
硝酸盐和亚硝酸盐……..
14. 15.
晒… 银……………………………………… 191
16.
铀…
第四章 1.氯代炕怪·
与健康有关的有机成分
(2)
1. 1 四氯化碳………………………………… 214
1. 2 2.
1 ，公二氯乙烧…………………………….
.220
氯代乙烯………………………………… 225 2.1 氯乙烯………………………………….
.225 .231
2.2
1 ，卜二氯乙烯…………………………….
2.3 三氯乙烯………………………………… 238
2.4 四氯乙烯………………………………… 248
3. 4.
多环芳怪 (PAH)
………………………… .258
农药……………………………………… 272 4.1 滴滴涕(全部异构体)……………………….
.273
4.2 艾民剂和狄氏剂…………………………… 278 4.3 氯丹…………………………………… .282
4.4 六氯苯……· 4.5 七氯和环氧七氯…………………………… 300 4.6 林丹…………………………………… .307
4.7 甲氧滴滴涕……………………………… .316
4.8 2 ，←滴…………………………………. .318 5. 氯苯……………………………………… 322 5.1 氯苯(一氯苯〉…………………·‘………… 323
5.2 二氯苯…………………………………. .325
6.
苯和低级烧基苯…………………………… 331
7. 酷和;氯盼………………………………… 337 7.1 具有毒理学意义的氯盼……………………… 337
7.2 2 ，也 6- 三氯盼……………………………. .340 7.3 五氯盼………………………………….
.342
8.
三卤甲烧………………………………. .3 -1 6
第五章感官指标和特性 1.铝………………………………………. .358
(3)
2. 氧化物…· 3. 颜色………
4. 5. 6. 7. 8.
铜… 硬度………· 硫化氢………… 铁………… 锤………
9.
溶解氧…………………· pH 值…·
10. 12. 13.
1 1.纳 硫酸盐………………. 味和臭…………….，
.
13.1 昧…· 13.2 臭………
14.
温度……………………..
.…………… .435
15.
溶解性总固体…
16. 1.虫度……· 17. 铮… 第六章放射性物质 1.前言…………………………..............
.460
2.
基本考虑………………………………… 461 2.1 剂量一反应关系…………………………… 461 2.2 剂量当量限值…………………………….
.463
2.3 对饮水水质的要求………………………… '463
3. 辐射照射的来源…………………………… 464 (-IJ
4.
对总 α 和总自放射性的建议值……………….-1 64
5.
氧…·
参考文献…
(5)
序 曾于 1971 和 1970 年分别公布了"国际饮用水标准 h 和 "欧洲饮用水标准川。现已对这些标准进行了评价，修改和
合并，并以"饮用水水质准则"为题，分三卷发表。各卷包括 下列简要内容 z 第-卷 z 建议。包括建议值及其主要的依据和监测要求。为
确保符合建议值，还尽可能推荐了治理措施。该卷包括饮用 水的微生物学、生物学、化学、感官和放射性指标各部分。 第二毒z 卫生基准及其补充资料。该卷包括为推荐建议值而
进行审议的饮水污染物和其它成分的卫生基准。此外，还提
供了检测水中污染物和控制措施的资料。总之，综述了现有 作为建议值依据的毒理学范流行病学和临床资料。
第三卷 z 小型集中式结水的饮用水水质控制准则。该卷主要 针对农村的小型集中式给水问题。包括对这种给水的污染评
价和控制技术的资料，如采样和分析的简便方法、卫生调查 和在这些区域进行调查和控制饮水水质的其它方法。本卷主
要涉及饮水在细菌学方面的安全性。 第二卷综述了确定建议值的依据，汇总和评价了饮水成 分对健康和感官影响的现有资料，并为建立和执行国家标准
以及从事这方面工作的研究人员提供了一种方便的参考文献
来源。该卷详细阐述了在第一卷中提到的对健康危害的资 料，应该认为它是第一卷的关键性补充文件。 a 国际饮用水标准，第三版，世界卫生组织，日内瓦，
1971.
b 欧洲饮用水标准，第二版，世界卫生组织，司内瓦， 1970.
(1)
本卷的第一和第二章论述了饮水的微生物学和生物学指 标。除讨论介水细菌病原体，包括使用指示菌的原理外，还 提供了关于监督要求和确保饮水细菌学质量可能性的详细资
料。在各节中，包括了水样的收集、贮存和运送，检测各种 微生物的建议方法以及消毒方法。此外，还简要地阐述了饮
水中的病毒。 虽然未提出生物污染物的建议值，但是谈到了致病性原 生动物和蠕虫，也讨论了给水中自由生活的生物。然而由
于缺乏标准检验方法，故未推荐建议值。因此其重点是放 在水源的一般防护，以尽量减少给水中的生物所带来的卫生 问题。 在准备该"准则"的过程中，专业组考虑了大量与健康有
关的有机物和无机物。他们详细综述了 37 种无机物和 46 种 有机物，确定了 9 种无机物和 15 种有机物的建议值以及 3 种有机物的试行建议值。第三章和第四章汇总了在确定建议
值时所依据的有关健康影响的资料和真它补充资料。此外， 还综述了某些其它的化学物和成分，但是由于各种原因，认
为不适于推荐基于健康影响的建议值。包括石棉、坝、镀、 硬度、镇、亚硝酸盐、银、纳、氯乙烯、以及某些氯酷和氯 苯。 综合的资料通常包括下列项目: 要来源和在水中的浓度 J
(a) 化学物的概述，主
(b) 人的接触途径〈水、食物、空
气等) ，包括其相对意义，
(c) 代谢(吸收、分布、存留、排
泄和生物转化) J (d) 对健康影响的资料，包括对有害生物 学作用的阐述及其对健康意义的评价、最敏感的危险人群的
鉴别、剂量和效应关系以及效应的发生率。每项总结均包括 有关参考文献。 (2)
仔细审议了 22 种饮水中的成分和指标以确定其对水感 官性状的影响。推荐了 15 种成分和指标的建议值。本卷第五
章综合了确定建议值所依据的基础资料，包括来源、存在、 接触途径以及对健康和其它方面的影响，也提供了温度、溶 解氧和 pH 对饮水水质的影响资料。(:
第六章论述了饮水中的放射性物质，这部分是与国际放 射防护委员会(l CRP) 密切合作进行准备的。它阐述了确
定总 α 和总 S 放射性建议值的依据以及提供了在实践中使用 这些建议值的指导。 "准则"第一和第二卷的准备历时三年之久，包括了近 30
个世界卫生组织成员国，大批科学家和 10 个专业组会议的 积极参加。这些机构和科学家的工作是使"准则"圆满完成的 关键，在此谨致谢意，他们的各字列于第一卷附录 1 。此外， 世界卫生组织环境卫生基准署的各国机构、各国际组织和各
位专家给予了极大帮助，他们的连续参加有效地促进了工 作。该项工作的梅调人是世界卫生组织总部的 H.
Galal
Gorchev 博士和世界卫生组织欧洲办事处的 w. Lewis 先
生。 衷心感谢丹麦国际开发署 (DANIDA) 和联合国环境 规划署 (UNEP) 对世界卫生组织提供的经济援助，从而
使"准则"的准备成为可能，并为本卷的出版提供了资金。也 应感谢美国环境保护局委派 Galal 期间所作的努力。 Gorchev 博士在两年
巳经认识到，一且具备新的资料，这些建议值的依据将
需要重新审定和修改，并可能推荐新的或修改的建议值。在 将来，国际化学品安全署(l PCS) 将能促进对给水中化学 物质的这种审定，该机构是联合国环境规划署、国际劳工组 (3)
织(l LO) 和世界卫生组织的一个合作机构。它具有两项主 要目标: (a) 评价化学物对人体健康和环境的影响 J
(b)
制订各类化学吨的接触限量(例如日许量以及空气、水、食物
和工作环境的最高容许浓度或理想浓度) ，这些化学物包括食 品添加剂、工业化学物、天然有毒物质、塑料、包装材料和 农药。特别值得指出的是，最终的专业组会议已经认识到， 在未来若干年内将可能具备有关下列化学物潜在健康危害的
新证据?届时将需重新考虑这些物质，包括石棉、铺、硝酸 盐/亚硝酸盐、镇、氯仿、其它三卤甲挠、三氯乙烯、四氯乙烯和 四氯化碳。
(~)
第→章微生物学指标
1. 饮用水的细菌学质量 有关饮用水的最常见、最广泛的危险是直接或闽接被污 水、其它污物、或人与动物粪便污染。如果这种污染是新的， 又假如污染源中有肠道传染病带菌者，就可能有某些活的致
病因子。喝下去或是在调制某种食品时使用这种污染的水， 就会造成更多的传染病例.
1.1
介水传播的病菌
粪便污染饮用水会导入各种肠道病原体一一-细菌的，病 毒的以及寄生虫的-一其出现与当时居民区中存在的传染病 和病原携带者有关。肠道病菌广泛分布于全世界。已知在污 染的饮用水中出现过的有各种沙门氏菌、志贺氏菌、产肠毒
素大肠埃希氏菌、霍乱弧菌、小肠结肠炎耶尔森氏菌和胎儿 弯曲杆菌。这些细菌能引起从轻度胃肠炎到重笃以及有时是
致死性的荆疾、霍乱或伤寒等轻重不同的疾病。 在环境中天然存在而且不吾成是病原体的真它细菌也可 能偶尔引起疾病。饮用水中有这种菌可以主要在局部或全身 天然防御机能降低的人中引起感染，这种情况似乎大多出现
于极老年人、极年轻人以及像烧伤或免疫抑制疗法的住院病 人之中。供病人饮用和洗澡用的水，如含有大量假单胞菌、
黄杆菌、不动杆菌、克雷白氏菌、以及沙雷氏菌之类的细菌 时可能引起多种传染病，可波及皮肤、眼、耳、鼻和咽喉粘膜。 病菌的传播方式，其中包括摄取被污染的水和食物，与
-
2
一
受感染的人或兽接触，以及暴露于气榕胶等。水在肠道细菌 性传染病传播途径上的意义，可因疾病及地方情况而显著不 同。虽然志贺氏菌有介水传播的，但志贺氏菌病的主要传播 途径一般不是水，而是在拥挤的生活条件下人与人的接触， 对比之下，霍乱一般是介水传播的，而沙门氏菌病是介食物
传播的。 各种介水病原体，都有一个引起人类感染所必需的最小 感染量，这个量的幅度很广。对伤寒沙门氏菌来讲，吃下相 当少的菌就能发病 F 福民志贺氏菌则需要数以百计的细菌， 而各种血清型沙门氏菌一般要上百万个细菌才能引起胃肠
炎。同样，对于产毒细菌，例如产肠毒素性大肠埃希民菌和 霍乱弧菌，可能需要多达 10 3 个细菌才能引起疾病。感染剂
量的大小也因人的年龄、营养状态以及接触时的全身健康状 态不同而异。不应低估饮用水以外传播途径的意义，因为只
靠提供安全给水而不同时改善卫生和个人习惯必定不会防止 传染。初步的应用卫生学教育是必不可少的。
1.2
利用指示菌的理由
由于认识到细菌性传染病可能介水传播，从而建立了常 规检验方法，以保证拟供人饮用的水不受粪便污染。虽然现
在有能力查出水中存在的多种病原体，但分离和计数的方法 常很复杂，很费时间。所以对饮用水中每一种可能因污染而
出现的病菌都进行监测是不切实际的。更合乎逻辑的办法是，
检查人与其他温血动物粪便中正常存在的细菌做为粪便污染 和水处理与消毒效果的指标。这种菌的存在意味着粪便的存 在，也意味着有肠道病原体的可能。反之，没有粪便共栖菌
则意味着大概也没者病原体。检查这种粪便污染指示菌，就 一 3 一
成了质量控制的一个手段。对于未处理水细菌质量的监测， 不仅在估计污染程度，就连选择最佳水源和需要的处理上也 很重要。
细菌学检查提供了最敏感的试验去发现新的因而可能是 危险的粪便污染，所以能以常规化学分析所缺乏的敏感性和
特异性提供水质的卫生评价。重要的是，水要定期而频繁的 检验，因为污染可能是断断续续的，不可能靠单份样品检查
出来。根据这个道理，经常地用简单方法检查饮用水，比不 经常地用比较复杂或成套试验更重要。一定要永远优先保证， 即使人力和设备受到限制，也要坚持常规的细菌检验。
必须懂得，一次细菌学分析最多能证明以特定培养方法 于检验当时在既定水样中有或没有污染或指示粪便污染的细
菌。此外，常规细菌学检查的结果一定要依据对该给水-一一包
括它的水源、处理和配水一一的全面了解加以解释。每当条 件变化已导致给水水质恶化，即使仅仅加大了污染的可能性，
细菌学检查的次数也应增加 F 这样，从仔细挑选的部位采集 一系列样品便可鉴定出危险之所在并得以采取治理措施。一 旦卫生监测(包括感观〉表明某一给水已明显受到污染，不
管细菌学检查结果如何，必须采取治理措施。对于无管道的 农村给水，卫生监测常只能采取定期检验的形式。
1.2.1
指示粪便污染的细菌
用正常肠道细菌做粪便污染指标而不用病原体本身，这 在监测和评价给水的微生物学安全性上已被普遍接受 E134 理想上，发现这种指示菌应当意味着所有有关病原体都可能
存在。指示菌应当在粪便中大量存在，而不存在或只少量存 在于其他来源，应易于分离、鉴定和计数，并且应不在水中
-4-
繁殖，还应在水中比病原体存活时间长，对氯等消毒剂的抵 抗力也要比病原体强。实际上任何一种细菌都不会完全达到 这些标准，虽然其中一部分已被大肠菌群特别是大肠埃希氏 菌所满足，可做为人或动物粪便污染的基本指标。其它能满 足这些标准某些部分的微生物，虽达不到大肠菌群的同等程
度，但在某些情况下也能做为粪便污染的补充指标。有或没
有特定粪便指示菌意义的大小，因各种细菌、特别是它与粪 便特殊联系的程度而有所不同。 用做粪便污染指示菌的细菌有 z 整个大肠菌群、大肠埃 希氏菌和那些称为"粪便大肠菌群"的大肠菌群，粪便链球 菌以及亚硫酸盐还原性梭菌，特别是产气英膜梭菌。厌气菌
如双岐杆菌和拟杆菌等在粪便中比大肠菌群还多，但没有可 供常规检查和计数的方法。其他菌群在环境中还有非粪便来 源，有的甚至可以在水环境中生长，因而降低了它们存在与
粪便污染关系的可信性。充分鉴定这种指示菌需要一系列试 验，这在常规监测中是不现实的。因此，水细菌学家们开发 的指示菌种属和菌群的定义是实用的而不是分类学的，而且
大部分是以多管法或捷膜过滤技术从水中检验与计数为依
据。
1. 2. 1. 1
犬肠菌群〈总大肠菌群〉
大肠菌群很久以来被人认为是饮用水水质适宜的细菌性 指标，这很大程度上是由于这些菌易于从水中检出和计数。 其主要特征是能在 35 0C 或 37 0C 培养下发酵乳糖，所以包括 大肠埃希氏菌、拘梅酸杆菌、以及克雷白氏菌等几个种。大 肠菌群不应在处理过的给水中检出，如果发现有，则意味着 处理不当或消毒不够，或是处理后又污染 (2) 。在这个意义上，
大肠菌群检验可用于做处理效果的一种指标。尽管大肠菌群
-5-
与饮用水中病毒的存在可能没有直接相关，在公用给水的微 生物学质量监测上，大肠菌群检验的应用仍不可少 E330 已知
许多寄生虫的包囊比大肠菌群对消毒的抵抗力更强。刚刚消 过毒的地表水中没有大肠菌群，却未必意味着没有贾第虫、 阿米巴或其它寄生虫的包囊。此外，大肠菌群不只来自温血 动物粪便，也还能来自植物和土壤 (4-6) 。在一定条件下，大肠
菌群也能从非金属构成物取得营养。因此有少数大肠菌群存
在(每 100 毫升水中 1-10 个菌)，特别是在未经处理的地 下水中，只要没有粪便大肠菌群，卫生意义可能不大。 粪便(耐热〉犬肠菌群
1. 2. 1. 2
这是些能在 44.0 'C或 44.5 'C下发酵乳糖的大肠菌群 z
其中包含埃希氏菌属，偶尔还有几株肠杆菌、拘梅酸杆菌和 克雷自氏菌。这些菌中，只有大肠埃希氏菌专一地来自粪便，
总是在人、动物、和禽类的粪便中大量存在，极少在未受粪 便污染的水或土壤中出现。充分按现代分类学鉴定大肠埃希
民菌，要用很大一套试验，对于水的常规检验是不现实的。 因此，检查并鉴定出这些细菌是粪便细菌或推测是大肠埃希
氏菌就被认为提供了足够评价粪便污染的信息。粪便大肠菌 群在配水系统中的复苏的似不会有，除非有充分的细菌营养
〈生化需氧量 BOD 大于 14 毫克/升)，水温高于 13 'C，而 且无游离余氯。
1. 2.2
其他的粪便污染指标
如果有某种怀疑，特别在发现有大肠菌群而没有粪便大 肠菌群或大肠埃希氏菌时，可用其他指示菌去肯定污染是否 来自粪便。这些第二位的指示菌，包括粪便链球菌和亚硫酸
盐还原梭菌，特别是产气英膜梭菌。 6 一
1. 2.2.1
粪便链球菌
水中出现粪便链球菌，一般表示有粪便污染 (8 ， 9) 。这些
链球菌通常存在于人兽粪便之中故而得名。其中包括粪链球
菌、屎链球菌、坚忍链球菌、牛链球菌、鸟链球菌、以及性
质上介于它们之间的中间型。这些菌极少在污水中繁殖，对 消毒剂的抵抗力略强于大肠菌群，但因其在盐类浓度不大的 水中存活时间很长 E102 ，例如，这种情况吁出现于混合供水时，
所以很少主张把这种指标用于饮用水质量控制。加之，除非 在常规方法中包括菌株鉴定，饮用水中每 100 毫升中不得超
过 100 个粪便链球菌这个数目的意义就会因粪链球菌液化亚
种的广泛存在而被贬低。做为辅助指示菌使用时，只要数据 收集充分，粪便大肠菌群与粪便链球菌之比(人粪便 为 >3 :1，其他动物为 <0.7: 1)在严重被粪便污染的原水水 源，确定污染源的位置是有用的。此外，还可用于评价大肠
菌群检验的可疑结果，特别是查到大肠菌群而查不到粪便大 肠菌群时的意义。这在干管维修之后核查配水系统中的水时
也有意义。
1. 2.2.2
亚硫酸盐还原性梭菌
为厌气性产芽胞菌，其中最有特征的是产气英膜梭菌
(魏氏梭菌) ，通常存在于粪便中但数量比大肠埃希氏菌少得
多。梭菌芽胞在水中存活的时间比大肠菌群长，如果消毒药 浓度、接触时间和 pH 不合适时抵抗较大。它在消毒过的 水中存在，可能表示处理不充分 (1 1)。
然而不能指望考虑用这种菌作配水系统常规监测，因为 它易于存活和累积，因而离污染源很远地方都可以长时间检
出，从而有假阳性的危险。
一 7 一
1. 2.3
水质的指挥
除菌落计数外，有人倡议 (12 ， 13) 用其他微生物〈包括绿肤
杆菌〉评价饮用水的卫生质量。但不论这些菌的检验，也不 论菌落计数，对于卫生质量的常规监测都不是必需的。只在
某些情况下给配水系统的一般清洁程度和评价瓶装水水质提
供指示上有价值。
1. 2.3.1
铜绿假单胞菌〈绿腺杆菌)
此菌常出现于人粪便，只是数目比大肠菌群少很多。它 是对婴幼儿和高龄老人而且已疾病缠身者的条件致病菌，常 从尿路感染和皮肤烧伤的病人分离出来 (4) 。此菌在未处理
水中出现，常同时伴有大肠菌群存在。但在饮用水中它可以 在没有大肠菌群时出现 05 ， 16)，对此部分解释是该菌有利用
配水系统和泵房等构筑物中某些物质支持生长的能力 (17) 。
虽说不应忽视此菌在饮用水中的存在，却没有人用它做有无 粪便污染的常规水检验。检验绿服菌在某些情况下是有价值
的，例如重新调制补液合液、婴儿食品、药房调剂、以及对 医院给水和瓶装水的监测等 (18) 。
1. 2.3.2
菌落计数
菌落计数可用做水中细菌总含量的评价。它并不代表水
中所有微生物的总数，而只是那些在特定培养条件下能在营 养培养基上形成菌落的细菌数。它在发现粪便污染上价值不 大，因而也不应该认为饮用给水安全性的评价上是必需的，
虽然地下水源水菌落计数的偶然增加可能是该水污染的早期 标记 (19) 。菌落计数在评价水处理过程特别是混凝、过洁、或
消毒效果上很好用，在处理过的水中细菌密度要尽可能保持
很低。也可以用来评价配水系统的清洁度和完整性以及该水 8 一
是否适用于制造食品和饮料，要减少变质危险，菌落数应当 很低。菌落计数的主要价值在于比较从同一给水的定期样品 所得结果，能检查出某一特定部位上异常的显著变化.
1. 3
障碍性生物
这是一群形态学和生理学上形形色色的生物，其中包括
浮游性和附着性藻类、真菌、甲壳虫、原虫和放线菌，以及
铁细菌和硫细菌。这些生物可产生令人不快的嗅、色、昧和 浊度，并因堵塞捷网或滤池而干扰处理过程。此外，有些浮
游生物本身虽然无害，却能包藏病菌，保护它们抵御氯消 毒。大多数障碍性生物通过常用水处理过程就可以相当容易 地控制住。不仅如此，世界卫生组织水的美学特征基准巴将
可能由障碍性生物引起的嗅、色、昧和浊度问题间接地收纳 进去〈见本书第五章〉。水中某些生物的出现可能是废铁
腐蚀或支持微生物生长的构筑物材料受到生物降解的标 志 (17~。促使微生物〈有时包括大肠菌群和绿肤杆菌)生长营 养成分的物质，包括塑料、橡胶、连结化合物和管线材料等
非金属材料。这些生物虽然通常可经处理去除， f!!可以在配
水系统的材料上或在沉淀物和粘液中定居，其所在之处可以
支持二柑水虱属或其他动物体的生长。滞留消毒有助于控制 这种困扰，但间或对数节管道进行冲洗或用聚氨脂 j包沫栓做 机械刷洗可能是必要的。 、障碍性生物也会包缠井节使地下水源出问题，降低给水
的感观质量并减少产量。实际上，障碍性生物的出现可表示 该蓄水层有有机污染 (20) 。
障碍性生物因其多样性和出现的难预科性，无法建议做 常规监测，尽管细菌学家应该知道它能降低水质。对障碍性
-9-
生物规定某种数量界限是不切实际的。
1. 4
消毒
饮水消毒的基本原则是保证破坏病原体，对进入配水系 统的病原体构成防护性屏障，并抑制管道环境中细菌的复苏。
因为消毒在保障饮水卫生质量上非常重要，一定要经常测量 消毒剂的浓度，并且最好做连续记录。对于小型给水，特别
是已知处于危险状态的，最好有简单的消毒和测量设备。
1. 4.1
消毒效果
消毒剂的比较放呆，可以用达到同等消毒效率所需的相 对浓度，也可以用消毒剂相同浓度所产生的相对消毒效率来
表达。但是由于各种微生物性质不同，以及 pH 、温度和水
的化学特征等试验条件标准化上的困难，只能对不同消毒剂 的比较效果做概括地叙述。在这种限度之内，消毒剂可按其 效果分类。这样，使用氯、二氧化氯或臭氧是更可取的，虽 然用氯时 pH 应小于 8.0 。由于氯股只是慢性杀菌剂，所以
不主张水处理时用它做为主要消毒剂，虽然在接触时间较长
的配水系统中维持余氯上可能是有用的。 同样，按递减顺序，各种类型微生物的相对抵抗力及其 大致存活情况开列如下 E 原生动物包囊〉肠病毒〉肠杆菌。
虽然灭活肠病毒所需时间与灭活肠忏菌相比有明显不同，但 保证给水微生物学安全所需最低限度的消毒剂残留量和接触
时间可以容易达到。因此，建议来自有潜在污染水源的水应 经常消毒，以保证杀灭某些微生物，包括可能比粪便指示菌
抵抗力更强的-些病毒。
一 10 一
1.4.2
消毒剂残固 1 坷
除消毒效果外，另一个重要的考虑是，这些药剂在饮用
水贮存与分配期间做为残留消毒剂的保留能力。除臭氧外， 所有其他实用消毒剂〈氯、二氧化氯、氯胶等)都能提供长 时间残留，继续对进入配水管网的出厂水进行细菌控制。但
氯牍这种慢性杀菌剂，一且决定使用，就应从整个配水系统收 集到足够的细菌学数据进行仔细评价，以证明控制细菌复苏 的效果以及对来自交叉连接的中等污染(1%污水)所提供 的保护能力 (21) 。
所有来自地表水源的给水，均应以消毒做为最低处理的
前提。凡执行消毒，整个配水系统都应保持有可测出的余氯。 维持与监测余氯有两个好处。余氯能抑制配水系统内微生物 生长，还可以对因交叉连接或渗漏而进入的污染提供一定防
护。余氯的突然消失是配水系统中进入可氧化物质或处理过 程效力不足的一个及时的指标。如果是用氯，游离余氯最好
维持在 0.2.......0.5 毫克/升，并对全系统逐日监测。当给水的 余氯低于某特定部位预期的常规量时，则应考虑包括增加投 氯、冲洗以及卫生监测在内的治理措施，因为丧失余氯可以 标志管网中进入污染物。要保证全系统维持余氯，加强或补
充投氯可能是需要的。已知游离余氯量过多可能与某种水体
中的有机物质反应产生臭味。此时主管部门或卫生官员应支
持对处理和配水的必要的改善，并规定适当的余氯浓度，做 为临时措施以保证用水的微生物学安全。
1.4.3
浊度的影响 a
有效的消毒取决于消毒剂和被消毒微生物之间有足够时 a 关于浊度对饮用水水质影响的较全面讨论，葫参考本书第五章，第 16节。
间的接触。各种细菌学和病毒学研究曾证明，水中各种类型 颗粒物质使微生物抵御消毒作用的程度迥然不同。总的来说，
精土与水混凝剂之类无机颗粒即使有保护作用也是很小。另 一方面，有机颗粒物质不论是细胞碎片、垃圾，也不论是线 虫或甲壳动物之活的或死的机体都能对其携带的微生物提供
明显保护。其赋予保护的程度，在很大程度上取决于颗粒物 质的性质而不是烛度计表示的浊度数量 E2230
实施消毒的全过程中，浊度必须始终很低，最好低于 1
油度单位 (NTU) ，始终不得大于 5NTU ，否则即使有足够 保证杀菌数量的余氯存在，消毒效果还会被颗粒物质干扰， 部分由于消毒剂的花费，部分由于掩蔽了微生物。水的?虫度 过大也会干扰细菌学检验，特别是用滤膜过滤法时。低 j虫度，
特别是以混凝法处理时，只有仔细控制和操作，保证混凝剂 量和 pH 适当，保证絮块层稳定，并通过对压头丧失和浊
度的监测保证过惊最佳运行时才能达到。滤器的反冲洗十分 重要，保证不致发生i虫度的渗漏。凡必须以混凝、沉淀和过 滤才能保证去除颗粒物质时，都必须坚持消毒以保证水的微 生物学安全。饮用水中凡有的有机物，在配水期间都参与耗
氯使游离有效余氯减少，尤其是在配水系统的盲端段。希望 定期冲洗干管以免有机物积聚。有机j虫度也可以起营养源的
作用而引起配水管网中细菌的生长，特别是在流动缓慢的部 位。细菌生长通过生物絮凝作用促使铁的积聚。其结果形成 粘液、碳酸钙和附着于管壁的其他碎渣，从而引起水质恶化。
细菌代谢的副产物或粘质内的分解可以引起味臭问题。7]<压 的起伏也能冲掉积聚的粘液与沉渣。
一 12 -
1. 5
细菌检验用水样的采集、 贮存和运输
‘ dr
‘
必须注意确保样品能代表检查的水，保证采样期间不窍
生意外污染。因此，采样人员应受过训练，了解工作的责任 性质。样品，有鲜明标签，附有采样地点、时间、水的性质以 及其他有关资料，并立即送至化验室分析不得拖延 (23-21>)号
1. 5.1
样晶榄
,.'
样品和样品瓶的大小取决于拟做的分析，但做常规大胳
菌群检查，不论用多管法或滤膜过滤法，通常 200 毫升足够。 特殊检测时可能需要较大样品。应使用无菌的洁净玻璃瓶，
瓶子应有磨砂玻璃塞或螺口盖，瓶颈以上应盖上纸或铝捕以 防污染。能耐受高压灭菌的聚丙烯:瓶也可以用 E2330A
1. 5.2
消毒剂的中和
i ‘
如果待检水有可能含氧、氯腊、二氧化氯或臭氧时，每 个 100 毫升大小的瓶中应加 0.1 毫升1. 8% 碗代硫酸纳以中 和残留的消毒剂。这至少应能中和 5 毫克/升有效一氯，故适
于常规采样。在残余消毒剂可能稍多的特殊情况下，则需补 加硫代硫酸纳。这个浓度的硫代硫酸铺，不论水样是否含氯如
贮存时不会对大肠菌群，包括大肠埃希民菌在内，给予显著 影响 (24) 。
采取消过毒的水样时，应于采样同时测定采样点的残余 消毒剂浓度。
,",""
13-
1.5.3
采样方法
在同一地点为不同目的采取几个样品时，应第一个采取 供细菌学检查的样品以避免发生采样点受污染的危险。 样品瓶在灌注样品前应保持不开盖。应以一只手取下瓶
塞，采样期间任何东西都不得与瓶塞或瓶咀接触。应将样品
瓶用另一只手靠底部抓住。样品瓶不得临时洗涮，水样灌到 加塞(盖〉后尚留有少许空间的程度为止。 如果不能把样品采入瓶中，应使用一具灭菌过的不锈钢 罐，可以用在罐内点燃燃料酒精的方法灭菌。灌满后将水样 倒入样品瓶。
1. 5.3.1
直接从水源采样
凡直接从河流、小溪、湖泊、水库、泉水或浅井采样 时，必定以获取被检水体的代表样品为目的 (23) 。因而采样
不可离堤岸太近、或离排放点太远，或从水底淤泥处取样。
应避免在不流动水区采样。 采样时应将样品瓶最好瓶口向下捺入水面下 15~30 厘
米避开浮渣，然后将瓶咀抬起迎着水流方向，如果水不流动， 应将瓶水平地捺入水中。 从井或湖泊了水库或水塘深处采样时，应使用特殊加重 的无菌采样瓶或罐。 从血吸虫病流行区采取未处理水水样时，应戴防水手套， 以免与水直接接触。 井水采样 r、、
1. 5.3.2
从装有手动泵或机械泵的井采取样品时，采样前应连续
开动先冲掉管道，中的死水。如有可能应当烧一下泵咀，最好
使用喷灯或丁烧火炬。采样前泵出更多数量的水全部丢掉， 一 14 一
然后让水直接由泵流入瓶内。
1. 5.3.3
处理厂采样
因为经常要从水处理厂采样，应遍设采样龙头，使处理 过程的各个阶段都能得到监测。装水龙头的水管应当很短，
如果不得不用长管子时最好让水不断流出。未处理或经部分 处理水的管道，应定期清洗去除粘质和淤积物。 如果没有安装水龙头，可把采样品沉入水池或管道内采 取样品。
1. 5. 3. 4
蓄水池呆样
只要可能，在所有用于贮水的池、塘、公用水库都应安 装采样龙头。如果未安龙头，则应以加重的罐或瓶沉入池中 采样。要非常小心不要在采样期间使水污染，并应避开有浮
沫的表层和有淤泥的底层。
1. 5.3.5
水龙头呆样
多数样品要由水龙头采取，如各处理工序、贮水槽、用 户室内、公用竖管等样品。公共卫生和给水当局应按商定计
划选择采样点。从配水系统采样时，必须非常注意水龙头的
选择。选择的龙头一定要干净，而且应直接从公用干线供水。 可能需要从供给高层大楼的水槽采取补充样品。应避免使用
阀杆和阀盖之间漏水的龙头，因为水会从龙头外部流下来污 染样品。水龙头外部安装的橡皮(塑料〉水咀或滤器都应摘 掉，并让水流几分钟保证在采样前把管子里的水冲掉。采样
前烧灼龙头应看做是适宜的方法。为避免从居民龙头采样引 起问题，给水当局应当考虑在配水系统的战略重点上安装有
防护的采样龙头。
1. 5. 3. 6
消防栓采样
公用给水的样品应从适宜的水龙头采取直接来自干管的 15 一
给水。如果做不到，则可以从大街上消防水栓采取。但需特 别仔细，包括水栓的冲洗和消毒。
1.5.4
样晶的运输和保存
保持样品不见光线、并最好在 4 "C至 10 "C间而不冻结条
件下冷藏，可以保证水的细菌含量在贮存时所发生的变化减 至最低。样品不得受致冷剂污染。采样后应尽快进行检验， 最好在 24 小时内。凡拖延检验，在解释结果时都必须予以考 虑，并在报告中说明。 如果无法避免拖延，则应考虑有无现场过滤样品的可能， 然后将谑膜放在转运培养基上于密封容器中运送邸， 24) 。将
睡膜转种在常用培养基上以正常方式做最终检查之前，可以
满意地保存在转运培养基上最多 3 天。
1.6
检出与计数污染指示菌的 建议方法
有两种基本方法可用于检出和计数水中指示菌，即多管
法和稳膜过撞法。前者是将已知水量加于多个含有适当液体 培养基的营中，后者是将一定量水样以臆膜过施将细菌截留
于臆膜表面。然而一定要重视，用一种方法、一种培养基得 到的结果与另一种方法和培养基所得结果之间，即使对同一
菌群，严格说也没有可比性。所以一旦选定一种方法，便要 一直使用下去，保证结果的可比性。
1.6.1
实验童设备与安全
要保证结果的可靠和可重复性，检验饮用水特别是检验
少数粪便指示菌时，适当的装备和设备是重要的。清洁而安 一 16 一
适的工作条件，不仅在防止样品的交叉污染上，就是对实验 室人员的健康和安全也是必要的。应当注意良好的实验室习 惯，包括工作人员的训练，孵箱和水浴的正确操作，以及培 养基的细心制作和使用质量控制方法等 (Zð)。
1.6.2
大团菌群幢验
因为大肠菌群在粪便中大量存在，而且浓度低到每 100
毫升 1 个尚可检出，所以是粪便污染的敏感指标。
1. 6.2.1
大肠菌群定义
大肠菌群〈总大肠菌群〉一词适用于能在胆盐或其他有 类似生长抑制性表面活性剂存在下生长，并能在 35 0C 或 37 "C 发酵乳糖在 24---48 小时内产酸、产气和产蓝的革兰氏阴性 杆菌。大肠菌群还是氧化酶阴性，且不产生芽胞。在 44 "c或
44.5 "C温度下具有同样性质的大肠菌群称为粪便〈耐热〉大
肠菌群。在 44 0C 或 44.5 C 下，发酵手L 糖和其他基质如甘露 0
醇，产酸、产气并分解色氨酸，形成昭|睐的粪便大肠菌群看 做推测性大肠埃希氏菌。证实大肠埃希氏菌，需要证明甲基
红试验阳性，不产生乙眈基甲基原醇，不能利用拘梅酸做惟 一碳源。这个定义不是分类学定义而是水检验工作中实用的
工作定义，它涉及好几个菌属的成员。因此在分类学上鉴定 为"类大肠菌"的某些细菌，在水质检验工作中未被承认。
例如厌氧生长又不发酵乳糖的类大肠菌，以及偶尔出现的非
耐热株大肠埃希氏菌。但这些都是非典型的，且在数量上被 反应典型的菌株大大超过，因而在实际上不会影响到对结果
的解释。如何选择供检出和确证大肠菌群的试验，应看做是 根据水的类型、检验的目的和实验室能力所做专业判断的一
部分。
1. 6.2.2
检出大肠菌群的技术 N
大肠菌群检出和计数所用的基本方法有两种，即多管法 和 i虑膜过滤法 CI ， 23 28) 。这两种方法不能得出严格可比的结
果。真理由之一是，滤膜上的计数看不到乳糖产生的气体， 但实用上却会得到可比的信息。
对于未处理水来说，检验大肠菌群和检验粪便大肠菌群 可能是水的微生物学质量的一个适当的基准。控制水处理过 程时，出厂水中不应检出大肠菌群(总大肠菌群)。给水中总 会多少有大肠菌群，重要的是此时应尽可能采取证实和鉴别 工作，以便断定是否是来自粪便的污染并有助于追踪来源。
<a)
多管法
初步试验是推测性的，因为看到的产酸产气反应有时可 能是别的一些细菌或多种细菌结合的产物。得出大肠菌群这 一推测之后，要将看到的反应以进一步确证与鉴别培养基做 的补充试验加以证实。假阳性反应率既取决于该被检水的细
菌群，也取决于使用的培养基。 将适当水量接种于几支培养管，便可以得到一定水量中
存在推测性大肠菌群的估计数。至于培养，假定每个接种时 接受了 1 个或 1 个以上活菌的管都会出现生长和适于所用培
养基的阳性反应。只要出现几管阴性结果，就可以从得出阳 性反应的管数估算出原水样中细菌的最可能数 <MPN)o 统
计学概率表就是用于这一目的，此表连同 95% 可信限列于卷 1 附录 2 0
MPN 法可应用于一切类型水体，特别是出度高的水体。
设备要求相当便宜、简单，而且阳性反应易于读出。但此法 仅提供某个样品中的细菌估计数，因而易有明显的固育误
差 E2430 应当用确证培养基同时也用固体培养基做传代培养，
-
18 一
保证做进一步鉴别试验之前获得纯菌。 被检水量 E 以液体培养基试验时的样品量取决于被检水
中细菌的猜测含量。质量好的水，
1 个 50 毫升量和 5 个 10
毫升量就足够。如果水质可疑或不明，要用 1 个 50 毫升， 5
个 10 毫升和 5 个 1 毫升量。 50 毫升和 10 毫升者要加在同量 双料培养基中，而 1 毫升者则加于 5 毫升单料培养基。污染 严重的未处理水应予稀释以便获得几管阴性，这样才能取得 明确的 MPN 结果。 培养基的选择 z 能用于推测性大肠菌群试验的有 4 种培
养基 z 改良无机盐谷氨酸盐培养基 (MMGM) ，十二烧基磺 酸纳膜际肉汤 (LTB) ，麦康开氏肉汤和乳糖肉汤c1 ， 23-28) 。
MMGM 是化学成份固定的培养基，具有可被大肠菌群利用 的有限营养。麦康开肉汤和 LTB 的选择性分别于有胆盐
和表面活性剂(十二皖基磺酸铀) ;乳糖肉汤是一种无选择性 培养基。接种后要将选择性培养管放在 35---37OC 培养 24 士 2 小时及 48 :1: 3 小时后观察特定的阳性反应。发酵手L 糖产酸产 气的检查方法如下 z 一切培养基中可放一倒置的小管 (Ðu
rham 氏管〉捕捉气体，而 MMGM ，麦康开肉汤和乳糖肉 汤还加有 pH 指示荆用以证明产酸。判做阴性的糖管在被丢
弃之前，应轻轻扣打以释放溶解的气体。应确信在 24 及 48
小时内呈阳性反应的所有各管部者大肠菌群存在。这种裁定 时间限制排除了不常见的产气缓慢的大肠菌群成员，但是这
种菌总的来说卫生学意义不大。芽胞菌的存在会造成假阳性 反应，但一般在下一步确证试验就被消除了。
确 ìlE试验 z 推测性阳性反应中有无大肠菌群应予证实， 因而进一步要做粪便大肠菌群和大肠埃希氏菌的鉴别试验。 将 24 及 48 小时内呈现推测反应阳性各管直接向含有乳 ~ 19 一
糖的选择培养基传代，于相同于推测试验所用温度下培养
48::!::3 小时。由于原来的反应可能含有杂菌，所以需要选择 性培养基，也可以使用煌绿(乳糖〉胆汁肉汤 (BGB) 口， 23 ， 20 。
气体的产生证实原推测性试验中有大肠菌群的存在。阴性反 应各管应轻轻扣汀，以释放溶解的气体。查阅 MPN 表即可
得出 100 毫升原样品中确证大肠菌群含量的估计数。 为了证实粪便大肠菌群的存在，应从推测性反应阳性各 管用 BGB 或含有胆盐的 EC. 肉汤传代 (23) ， 44 土 O.25 "C
培养 24 士 2 小时飞产生证实粪便大肠菌群在原来管中存在， 并可重新得到证实的 MPN 值。此外，推测性大肠菌群计数 也可以经接种于膜陈水管来做，
44 "c或 44.5 "C培养 24:!: 2 小
时后还可以用来测昭|垛形成试验 (23μ3 。如果是阳性，则应以
各种生化试验证实 EC 的存在。习惯上，只用一管培养基 〈于 44 C 既可证明产气也能证明呵!味形成)便可确证推测性 0
EC
(29)。为了避免与缺乏半乳糖苦透酶 EC 菌株有关的问
题，曾使用甘露醇作为发酵基质。如果只有一项反应阳性
〈或产气或 psJ 垛)，则应分开用两支培养管重新测验。
(b)
惊膜过滤法
水中大肠菌群的数目，也可以用谑膜过滤一定量水样的 方法予以确定。， 24~28) 。谑膜通常由数种纤维素醋所构成，典
型孔径为 0.45μ ，它能把样品中存在的大肠菌群和许多其他
细菌截留下来。然后把洁、膜面向上放在选择性培养基上培养。 待滤膜上生长出典型的产酸或产醒菌落，则依据培养温度， 作推测性大肠菌群或推测性粪便大肠菌群计数。由于滤膜上 • Escherichia coli =大膀埃希氏菌的缩写。 b 如按《水与废水检查标准方法 F 规寇。町，应将 44 土 o .25 C全部换成 0
44.5 土 0.2 "C。
一 20 一
无法检查产气，故假定凡产酸产醒的菌落也都产气。但下一 步确证时使用的技术是要证明产气的。结果的表达是按 100 毫升原样品中的菌数。
实际上，滤膜过滤技术所得结果可以与多管法的结果相 比。但是如把样品过滤两个洁、膜，一个培养在 35 "C或 37 "C， 另一个用 44 cc 或 44.5 "C培养，那么确证的方法就多少受到 简化，因为在这种较高温度下直接计算存在的粪便大肠菌群 数是可能的。
滤膜过拮、法的一个优点是取得结果敏捷，从而允许迅速 采取治理措施，恢复正常运转。大多数水的检验都可以使用
此法，只是高出度水例外，因为惊膜会被堵塞往往使滤过的 水达不到充分水量。滤膜也不适用于含大肠菌群少而在该培
养基上生长的杂菌多的水，因为杂菌长满宿、膜会干扰大肠菌 群生长。如果水中气单胞菌之类发酵乳糖而不产气的细菌占
优势，因其结果的假阳性率高，对滤膜上所有推测性大肠菌 群菌落都应进行证实。氧化酶试验将有助于消除这种假阳性 结果。好气性芽胞忏菌在液体培养基中会造成假的推测反应，
但在滤膜上不会。滤膜技术稍改一下便可支持受伤细菌的检 出。在较低温度下做前培养或用选择性小的培养基使受伤的 细菌恢复过来并开始生长，而后的试验完全按正常方法进 ~:::;: (23 , 20 f丁。
滤、膜过谑技术是直接做单个菌落计数，但仍有统计学误
差。加之，能够检查菌落形态和直接做传代，从而减少来自 混合培养假阳性反应的可能性。尽管要求的设备数量不多， 但比起多管法的还是要贵一些。滤膜可以重复使用，只要充 分洗涤和煮沸灭菌后仍然完整，但只能再用于同一培养基。
此外，应该知道不同的撞膜有不同的特性，选择恰当，不仅 一 21 一
对目的细菌的生长，而且对有关水都是十分重要的。 滤膜过滤法的结果不一定与多管法的相同，尽管实际上 两者结果一般可以相比。因此应当用两种方法做适当系列的 平行试验以证实滤膜过滤技术适于有关水质，这是很重要的。 过滤设备和技术 z 该装置是将一块饶结的多孔玻璃板架 在固定于底座的硅橡胶垫圈上，底座上可以连接一个带刻度
的漏斗。烧结或多孔板上放臆膜。临用时，将谑膜架安装在 有侧咀的三角瓶上，侧咀与真空系统相连。可以把几个谑膜
架安在一具多板管上，可同时过谑几个样品。水过晤后取下 i虑膜，面朝上放在适当的琼脂培养基上，或是放在平皿中饱 蘸液体培养基的纸垫上，以适当温度培养。关于此装置的详 细情况可从文献查到(1， 23-28 ， 30) 0
培养后应在明亮处检查谑膜。菌落的外观取决于所用培
养基，但符合特征的全部菌落不问大小都应计数。必要时将 单个菌落传代，用液体培养基做确证试验，或接种于固体培
养基上求纯化后再进一步做鉴别试验。 为了检查水样的大肠菌群(总大肠菌群)和粪便大肠菌
群(推测性 EC) ，需要各用一个法膜，适当的培养基，以及 不同的培养温度。 待检水的量 z 大肠菌群总数和粪便大肠菌群的计数要分 两份来做，每份水要 100 毫升。除非每 100 毫升样品中大肠 菌群含量超过 100 个，否则每个试验都要过洁、 100 毫升样品。 太脏的水样，选择的过滤量希望在滤膜上长出的菌落数在
10~100 个之间。如果这个量小于 10 毫升，则应加入无菌 的稀释液使过滤量超过 10 毫升，用四分之一浓度林格氏液，
0.1% (W /V) 蛋白脏水，或缓冲稀释用水等等均可。 培养基的选择 z 很多种培养基都可供以惊膜过滤法检查
-
22 一
大肠菌群使用。其中，手L 糖十四烧基疏酸铀琼脂〈313 ，乳糖 TTC 十四皖基硫酸铀琼脂 E312· ，以及十二皖基硫酸铺乳糖肉 渴邸， 29)都可用于 35""""37OC 大肠菌群计数和 44 C 粪便大肠 0
菌群计数。远藤型培养基只应在 35 .......， 37OC 大肠菌群计数时使 用，而 44 0C 粪便大肠菌群计数时只能用 MFCb 肉汤。尽管这
些培养基全都是靠发酵乳糖来检出推测性大肠菌群，但其反 应特性却各有不同。菌落在远藤培养基上的典型金属光泽取 决于醒类的形成。 确证试验 z 拟证实的单个菌藩数量要根据水和检验目的 而定。所得信息与多管法时的推测性阳性反应相似。证实大 肠菌群的存在，一定要以 35 0C或 37 0C在 48 小时内证明发酵
乳糖产气，证实粪便大肠菌群则需 44 0C或 44. 5 C 24 小时。 0
此外， 44 OC 利用色氨酸产生月i 昧则证实有推测性 EC 。证实 的方法稍与多管法不同，滤膜上的菌落有时在 BGB 肉汤之 类选择性培养基上生长不好，所以不宜直接作确证口， 23 ， 24) 。
再者因为滤膜一般用 35 0C或 37 0C与 44 0C 或 44.5OC 两种温度
培养，所以只要证明每个菌落在 44 0C下产气就能直接确证， 而不必再做较低温度的试验 (23 ， 20 0 35 0C或 37 0C的产气可以 用乳糖蛋白陈水 (24)或十二烧基硫酸铀膜际肉汤 (23 ， 20证明， 而 44 0C则需用乳糖蛋白陈水或 EC 肉汤 (23 ， 24) 。应该用膜脏 水做明喋形成试验 (23 ， 24) 。
最好对滤膜上所有推测性菌落都能检查，可是这并不总
是实际可行的。然而处理过的水不应有大肠菌群，所以应对
这类样品来的全部菌落都应做进一步调查。对于处理前的原 a lactoac TTC tergitot 酸纳琼脂。 agar= 乳糖氯化2 、 3 、 5-三苯四氮嗖十四烧基硫
b MFC = Medium for faecal coliforml organisml =粪便大膀菌群培 养基.
-
23-
水，应将样品稀释以便在滤膜上获得可控制的菌落数。此时
可望检查一些可疑菌落(最少 10 个即可〉以确定污染的性质 和程度 c
1. 6.2.3
大肠菌群的鉴别
由于证实鉴定饮用水中任何推测性 EC 非常重要，需要
做进一步鉴别试验，必要时包括使用市售成套鉴定试剂。在 温带气候，可以预期这些试验能证实凡在 44 'C产酸、产醋、 产气或昭|昧的细菌实际上就是 EC 。而热带地区，其他细菌， 例如卫生学意义较小的肠杆菌属细菌 (32-34) 能呈现推测性的
EC 反应，但可以用这些试验加以鉴别。鉴别试验一定要用
纯培养物来做，纯培养物是从确证培养基上分离下来，继培 养于无选择性平板培养基上。然后用典型菌落作昭|味、甲基
红、 v-p 、拘梅酸、氧化酶等试验，必要时加用其他生化反 应 (27 ， 28) 。其结果，连同确证培养基的结果，按1. 6.2.1 节
规定做 EC 鉴定。同样的试验也可用于鉴别其他大肠菌群。
1. 6.3
粪便链球菌的撞出
多管法和 i虑膜过滤法都可以用在本群细菌的推测性检 出，只是应用这些技术时也会碰到以前对大肠菌群时看到的 同样限制。这些试验的结果还需要证实。， 23 ， 24 ， 26-28) 。
1. 6.3.1
粪便链球菌的定义
粪便链球菌属于 Lancefield 血清型的 D 群和 Q 群，其
中包括粪链球菌及其变种，屎链球菌，坚忍链球菌，牛链球 菌，以及它们之间的中间型。还包括马链球菌和鸟链球菌。 它们能在 45 'C在 40% 胆汁和能仰制大肠菌群和大多数革兰 氏阴性菌的浓偶氨化锅中生长。其接触酶阴性。许多菌种能
耐受 60 C30 分钟加温，并能在 pH9.6 含 6.5% 氧化铀培养 24 一
0
基中生长。
1. 6.3.2
多管法
把适量水加于单料或双料葡萄糖〈右旋糖)偶氮肉 汤。， 23 ， 24)中于 35 "C或 37. ::!::O. 5 0C 培养 48::!:: 3 小时，需要时 可延长至 72 小时 (24) 。呈现产酸、浑?虫并常伴有沉淀的管，
被看做含有推测性粪便链球菌。看不到产气。 证实阳性管中确有粪便链球菌的方法是，接种到乙基紫 偶氮肉汤中 35. 士 0.5 "C培养 48::!:: 2 小时。 确证培养基一般要求接种量大 (23 ， 24) 。粪便链球菌确证
反应阳性管呈现紫蓝色沉淀伴有浑 j虫。另一种确证方法是把 推测性阳性管用 Pfizer 民选择性肠球菌琼脂 (PSE) 平
板传代， 35::!::0.5 "C培养 24 小时后生长出带棕色晕的棕黑色 菌落，即证实有粪便链球菌的存在 (23) 。
100 毫升原样品中推测性和确证性粪便链球菌的计数，
如前大肠菌群时所述，使用卷 1 附录 2 的概率表。
1. 6.3.3
滤膜过滤法
以滤膜过惊技术做粪便链球菌计数时，一般使用 KF 琼 脂及 rn-肠球菌琼脂 (Slanetz
&
Bartley) ，这两种培养
基都含有偶氮，但碳水化合物不同。 rn-肠球菌琼脂的选择
性与培养温度有关，先在 37 士 0.5 "C培养 4 小时再在 44 士 。: 25 "C培养 44 土 3 小时，此法对粪链球菌和屎链球菌特别有 选择性 (35) 。此项试验对实际工作有好处，在证实有粪便污
染上〈特别是人的)可以做为大肠菌群试验的补充 (24) 。粪便 链球菌在这两种培养基上的总检出率者时很低 (1 口， 35) 。一般
从水检出全部粪便链球菌时最常用 KF 培养基，因其检出率 和选择性都很好 (36 ， 37)0 KF 琼脂如用煮沸而不用高压灭菌，
则对其选择性更好。滤膜放在 KF 琼脂上在35 "C或 37 土 0.5 0C
-
25 _,
培养 48 土 3 小时 (23' 。由于此培养基既稳定又对粪便链球菌 有选择性，在进行实验室正式试验之前，可将、捷膜放在它上 面运输达 3 天之久。以滤膜过滤法检出粪便链球菌时的培养
基一般有很强选择性，只要 KF 或 m- 肠球菌琼脂上长出红色 菌落，一般不用进一步证实就可以想到是粪便链球菌。必须 做确证试验时，能证明能在 40% 胆汁存在下于 44 "c生长并 且接触酶反应阴性即可。 粪便链球菌的密度应以每 100 毫升原水样中的菌落数表 示之。
1. 6. 3. 4
粪便链球菌的鉴别
如果把粪便链球菌试验用来估计污染来源，则需做种属
鉴定。对全部分离物做初步鉴定之后，凡接触酶阴性的都应 进一步鉴定 E 即能在 45 "C、 pH9.6 、以及 6.5% 氧化锅中生
长，能分解七叶苦和淀粉，并能发酵糖、还原 0.1 %亚甲兰 牛乳。
1.6.4
亚磺酸盐还原梭菌的撞出
亚硫酸盐还原梭菌，特别是产气英膜梭菌(即魏氏梭菌)， 也可以用做粪便污染指示菌 (38) 。这群细菌的特征是能形成
芽胞和亚硫酸盐还原成硫化物。靠产生黑色硫化铁沉淀做推
测性梭菌检出标志的培养基就是利用这一性质。芽胞的检出， 应将水样加热至 75 .......， 80 "C 10 分钟使繁殖型菌细胞灭活 (24) 。
1. 6.4.1
亚硫酸还原梭茵的定义
本群细菌系由一群厌气性忏菌组成，它们都有使亚碗酸 盐还原、产生芽胞、革兰氏染色阳性、以及接触酶阴性等特
性。至于产，气英膜梭菌还发酵乳糖、廉糖和蔷薇醇产气，在
石蕃牛乳中产生典型"汹涌泉障"反应，水解明肢，以及产 26-
生卵磷脂酶和酸性磷酸酶。此菌无动力。
多管法和 i虑膜过滤法都能用于亚硫酸盐还原梭菌的检
出。
1. 6.4.2
多管法
与大肠菌群时的多管法相似，只是不用吾产气。为保持 厌氧状态，整个培养时间必须使用螺口帽瓶，其中装满鉴别 性强化梭菌培养基 (DRCM) 之类含有葡萄糖和亚硫酸盐
的肉汤 (20 。在单料和双料培养基中接种适量样品， 35 cC 或
37::!:: O. 5 'C培养 48 土 3 小时。出现黑变的各管被认为含有亚 硫酸还原菌，如果样品在试验前加过热，贾可断定是产芽胞
的亚硫酸盐还原性棱菌。 证实产气芙膜梭菌的存在，只须把推测性反应阳性瓶中 的生长物接种于石恋牛乳管中 (22) ， 35. 或 37 土 O.5 'C培养 48
土 3 小时即可。典型的"汹涌发酵"反应和出现酸性，即证 明是产气英膜梭菌。需要进一步确证，可作动力试验和硝酸 盐还原试验 (39) 。
1. 6.4.3
滤膜过滤技术
水的亚硫酸盐还原梭菌计数上还没有哪一种培养基被普 遍接受，但建议使用亚硫酸铁琼脂(l SA) (38) ，磺胶嘈睫-多 粘菌素-亚硫酸盐琼脂 (SPS)(39) ，产气英膜梭菌培养基
(mCP) (40) 。各培养基的培养时间都是 0
24 小时。培养温度
因培养基而异， SPS 为 37 C ， mCP 为 45 0C ，而 ISA 为
48 'C。厌氧条件是生长所必需，这对于 ISA 及 SPS 可用 琼脂覆盖的方法达到，但对 mCP 因需要将平板连续暴露于
氨蒸气来证明磷酸酶减少，所以不能用覆盖洁，只能在厌氧
环境中培养。 如果过滤前将样品加热到 75-80 CI0 分钟，则对亚硫酸 一 27 → 0
还原梭菌有很高选择性。只有使用 ISA 时才需要对产气英
膜梭菌的存在进行证实，因为 SPC 和 mPC 都含有一定的 抗菌素，所以对该细菌极有选择性。尽管如此，含亚硫酸盐
的培养基还是经受了时间考验而被推选做标准培养基，此外 它还有一个长处，使得滤膜过滤技术在使用和操作方式上与 MPN 法相似。
1. 6.5
铜操假单胞菌(操踉抨菌〉
对铜绿假单胞菌可下定义并据下列鉴别点易与其他产荧
光的假单胞菌区另IJ: 产生色素，在 42 C 生长，水解酷蛋白， C
以及利用有机碳为特殊碳源。
1. 6. 5. 1
铜绿假单胞菌的定义
铜绿假单胞菌为革兰民阴性无芽胞忏状细菌，它可在 42 'C生长，能产生绿服菌素和荧光色素，氧化酶及接触酶均
为阳性，还原硝酸盐为亚硝酸盐，液化明胶，水解酷蛋白但 不能水解淀粉，氧化葡萄糖，并将乙酷胶还原为氨。
1. 6.5.2
多管法
培养基含有天门冬眈肢会促使色素形成，它构成目前多 管法所用改良 Drake 氏肉汤的基础 (23) 。培养基中加入乙醇
(20 毫升/升〉可防止 35 'C或 37 'C培养期间其他革兰氏阴性
菌的生长。 将适量样品加于单料或双料肉汤管中在 35 'C或 37 "C培 养 4 天。对于有生长的各管，应逐日检查色素产生并在紫外
线下出现荧光的情况。推测反应中的铜绿假单胞菌，应通过 传代在澳 f七十六烧基主甲腔牛乳琼脂或乙雕股肉汤 (23) 中予
以证实。在牛乳琼脂巾 4 1. 5 0C 培养 24 士 2 小时后酷蛋白的水 解与蓝绿色荧光色素的产生，以及在乙眈脐肉汤Jil 35 oC 或 2 l!一
37 "C 36 小时出现紫色表示产生险性条件，即证实杳铜绿假 单胞菌存在。 滤膜过滤技术
1. 6.5.3
通常使用加乙醇的改良 King 氏 A 培养基 (23) 。过滤后，将
谑膜放在 King 氏 A 培养基上 35 0C 或 37 "C培养 48 :l:: 3 小时。 有荧光和着绿色素的菌落都被看做是铜绿假单胞菌。应将菌 落接种在牛乳琼脂上， 4 1. 5 C 培养 24 小时如出现酶蛋白水 0
解，产生绿色素及荧光者即可确证 E232.
1. 8. 8
菌蕾计数
目前计算水中细菌含量的方法一般有平板倾注法和表面 涂布法 (1 ， 23-28， 31 )。谑膜过谑技术，最近也发展起来 (4 1)。
菌落计数时所用营养培养基只能支持任一水样中部分细 菌的生长，而且还因培养基的不同而异。厌氧菌，即使在倾 注平板中也一般不长。此外，微生物还会结团和成链，所以
实际上形成的菌落数明显地低于该样中真正存在的活微生物 数。不同的微生物有不同的生长最佳温度，所以一般总要把
同一样品做两块平板，一块在35~370C 培养 1~2 天，另一块 在 20~22 "C培养 3 天。培养时间的长短会影响菌落数目。因
而重要的是严格坚持相同的作法，使结果始终有可比性。对 于瓶装水，通常是把平板在 35 "C或 37 0C 培养 3 天 (23) 。计数
是以 1 毫升原水样中的菌落形成单位
(colony-forming
units = CFU)
数表示，并说明所用培养基、温度、培养时
间和方法。简言之，可将此法概括如下 z 将样品做 10 倍系列 稀释，其数目取决于该水的性质和历史。每稀释度各取 1 毫
升加于无菌平皿中，共做两套，将融化后冷却至 44~46 "C的 营养琼脂加于各平皿，每个 15 毫升，立即将样品与培养基混
一 29 -
匀。琼脂固结后翻转平板在规定温度中培养。清数每块平板 的菌落数，以几块平板的菌落数的几何均数乘以该稀释度的 倒数即得原样品中的细菌数。如果平板上的菌落数在 30--300
个范围之外，虽然仍可计数，但该结果只能做为估计予以记 录。
1.6.7
病原菌的检查
虽然还没有把直接查找特殊病原菌列入水的细菌学检查 常规，但偶尔会需要检查肠道病原体，例如，流行期间或评
价新水源时。此时如果做大水样量检查，又使用对特定肠道 病原体有选择性的培养基，则成功的机会很大。检查步骤可
以包括下列全部或某几步 z 浓缩样品中的微生物，接种于营 养肉汤，用选择性琼脂培养基做继代培养，最后做可疑菌落 的生化与血清学检查。只靠单一的方法不如尽可能使用几种 方法，这总会有机会把病原菌检查出来邸， 24 ， 26-28)。特别是
检查沙门氏菌属时尤为如此，因为没有任何单独的方法适合 于全部血清型。
1. 6.7.1
样品的浓缩
使用的方法很大程度上取决于特殊物质的量。低由度水 样可以通过滤膜过滤。各种原水的浊度较大，可利用硅藻 土 (23 ， 24) 或滤深筒式臆器 (42) 以协助过滤，从而允许处理较大
量样品。另一种方法是可以采用纱布垫法 E233 ，特别在病原
菌数量少或非连续出现的时候。
1. 6.7.2
沙门氏菌
样品浓缩后可能需要用缓冲蛋白陈水做前增菌，然后再 用含四焦硫酸铺、亚晒酸盐、氯化娱或孔雀绿的肉汤增菌。
生民后转种于诸如煌绿、亚硫酸锡、木胶糖赖氨酸去氧胆酸 30 一
铺 (XLD) 琼脂、去氧胆酸铀拘橡酸盐、或麦康开琼脂之 类的培养基上，并对可疑菌落做生化学及血清学检查。生化
筛选试验应包括三糖铁琼脂，昭|喋形成，脱竣基酶，以及自 半乳糖苦酶活性。血清学试验应包括用多价抗-0 、抗 -H 、 以及抗-Vi 血清作凝集反应。一定要优先摒弃有自发凝集的 菌株。怀疑伤寒沙门氏菌时最好用亚晒酸盐 F 培养基。目前 用多管法计算沙门氏菌的数目。 志贺氏菌
1. 6.7.3
因为大肠菌群和大多数普通变形菌菌株对志贺氏菌拮
抗，最好选择能减少来自那些拮抗者的挥发性物质和副产物 的选择性培养基 (23)。可以使用 pH 调到 8.0 的营养肉汤
(此 pH 大肠菌群不喜生长〉。用一种自发细胞毒性培养基 也能成功取得志贺氏菌增菌，该培养基是在膜酶消化酷蛋白 大豆汤的基础上加 1mM/ 升 4-氯-2-环戊基苯基自-D-半乳 糖毗喃糖苦， 0.25% 乳糖，并以拘棒酸缓冲至 pH6.2(43) 。
于 35 "C培养 6.......18 小时。培养 6.......18 小时后在 XLD 琼脂
上划线培养。可疑菌落做生化筛选试验，然后用志贺氏抗血 清(多价及型特异性血清〉做可疑菌落的证实。 霍乱及非霍乱弧菌
1. 6.7.4
初增菌使用险性蛋白陈水或牛胆盐亚暗酸盐蛋白脏水，
继代培养用的选择培养基为硫代硫酸拘棒酸胆盐静、糖琼脂或 牛胆盐亚暗酸盐明胶琼脂 (44) 。把可疑培养物培养于 K
liger
氏铁琼脂。然后将这些培养物进一步做尿素酶氧化酶活性的 筛选，对于尿素阴性和氧化酶阳性的菌株应交送者关实验室
进一步做生化试验和血清学分型。
1. 6.7.5
肠道致病性犬肠埃希氏菌
使用检查水中粪便大肠菌群的技术。可疑菌落的证实与 31 -
EC 同，如果有流行病学证据证明，可将继代培养物送交有
关实验室做血清学分型，必要时做产肠毒素能力试验。
1. 6.7.6 0
小肠结肠炎耶尔森氏菌
发现改良远藤琼脂是最适宜的通用培养基，由于 25 和 35 C 培养的菌落形态学特征都明显。菌落是深红色的，而
且培养 72 小时后十分明确。麦康开琼脂，只要是用 25 "C培 养，生长也同样好 (45) 。所有可疑分离物都应以鼠李糖、棉
子糖、以及蜜二糖在 25 和 35 "C做生化筛选。如果有流行病 学证据证明，应将继代培养物送交有关实验室做血清学分型 和抗菌素敏感试验。
1. 6.7.7
胎儿弯曲杆菌
以滤膜过洁、法配以含万古霉素、多粘菌素和三甲氧韦二 胶嘈腔的血琼脂，可成功地分离这种病原体 (46) 。在降低厌
氧罐中氧气分压条件下 42---43 C 培养 3 天，逐日检查非 j容 0
血性灰色粘菌落(直径 1---2 毫米〉。可疑菌落以革兰氏染色呈 典型的弯曲的 "S" 型，并且氧化酶和接触酶反应均应为阳
性，有动力， 36 0C 好气性条件下不能生长。应将继代培养物 送交有关实验室做进一步生化试验。对于散发流行时的分离
物作血清学分型是不切实际的，因为本菌为多相性，不论用 共同抗原或不同血清型的组合都不行。 〈吴联熙译〉
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33. Boizot , G. E. An examination of the modified Eijkman method applied to pure coliform cultures obtained from waters in Singapore. Journal
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hygiene
(Cambrid ge) , 41 : ;:;66~569 (1 941). 34. Evison , L. M. & J:Lnes , A. A comparison of the distribution of intestinal bacteria in British and East African water sources. Journal 35. Stanfield , G. et a l. Isolation 1978. 36. Clausen , E. M. et a l. Fecal streptococCÌ: indicators of po lI ution. In: Hoadl 町， A. W. & Dutka. B.
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applied
bacteriology , 36: 103~118 (1 973).
0/ /aecal streptococci /rom sewage. Stevenage , Water Research Centre ,
J.,
ed. ,
Bacterial indicators 0/ potential health hazard s associated with water. Philadelphia , American Society for Testing and Materials , 1977 , pp. 247~264.
37. Summary report
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a Working Group on
BJcteriological Examination
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Water. Copenhagen ,
WH') Regional Office for Europe , 1975. 38. Bonde , G. J. Bacterial indication on water pollution. In: Droop , M. R. & Jannasch , H. W. , ed. , Advances in
aquatic microbiology , vo l. 1, London , Academic Press , 1977 , 381 pp. 39. Angelotti , R. et a l. Quantitation of Clostridium
per/ringens in foods. Applied microbiology , 10: 193-199 (1 962). 37 一
40. Cabelli , V. J. Clostridiurn perfringens as a
lV ater
quality indicator. In , Hoadley , A. W. & Dutka , B. J.. ed. , Bacterial indicators of potential health hazards associated with t且later ， Philadel phia , American Society for Testing and Materials , 1977 , pp. 65-79. 41. Taylor , R. H. & Geldreich , E. E. A new membrane filter procedure for bacterial counts in potable water and swimming pool samples. Journal of the
Arnerican Water Works Association , 71 :402-405 (1 979) .
42. Levin , M. A. et a 1. Quantitative large-volume sampling technique. Applied rnicrobiology , 28:515521 (1 974). 43. Park , C. E. et a 1. Improved procedure of selective enrichment of Shigella in the presence of
Escherichia coli by use of 4-chloro-2cyclopenty lpheny 1 beta-D-galactopyranoside , Canad i an
journal of rnicrobiology , 23:563-566 (1977). 44. Guidelines for the laboratory diagnosis of cholera. Geneva , World Health Organization , 1974. 45. Highsmith , A. K. et al. Is olation of Yersinia
enterocolitica from well water and growth in distilled water. Ap plied and environrnental
rnicrobiology , 34:745-750 (1 977). 46. Skirrow , M. B. Carnpylobacter enteritis: A "new" disease. British rned ical jourflal , 2 '9-11 (1 977).
-
38 一
2. 饮用水的病毒学质量 2.1 概述
和介水传播传染病有关的病毒，主要是在肠道中繁磕并 在受感染者粪便中大量排泄的那些病毒 ω。据报告，每克粪便 中病毒浓度高达 10 8 病毒单位。尽管肠道病毒离开活的宿主
不会繁殖，但在水环境中有显著存活能力，在数天或数月内 可能仍是活的 (2)。病毒进入水环境的主要途径是污水排放。 用现代有效方法，病毒在污水中的数量变动很大，曾发现最 高达 10 6 单位/升 (3) 。任何一天当中，从城市污水都能分离到
一百余种已知肠道病毒中的多种，优势类型是当时在人群 中流行的类型。还没有能分离污水中可能存在的每一种病毒
的方法。还没有一种方法能分离出污水中可能存在的所有病 毒。污水处理可以减少病毒浓度 90.........99.9% ，其实际程度主
要取决于处理的性质和程度。即使是三级处理也不能水流完 全不含病毒 (4) 。由于污水与受体水混合把病毒带到下游，存
活时间的变化取决于温度和其他几种未充分肯定的因素。因
而者来病毒要在被污水污染的水中存在下去。曾有记录，水 处理厂的进水口处病毒的数量高达 49 病毒单位/升 (5) 。
2.2
接触途径
一般相信，接触肠道病毒的主要途侄是与受感染人的直 接接触，或是通过粪便污染物体的接触。但是由于病毒的存
-
39 一
活时间长，也由于感染剂量低，因而接触和尔后的感染会以 不明显的方式出现，包括喝进被污染的饮用水.
给水由于污水污染所引起的病毒性肝炎和胃肠炎的爆发 流行，在流行病学上曾被充分证明 (6) 。相反，病毒通过符合 饮用质量的饮用水的低水平传播，曾被怀疑是在人群中维持 地方性肠道病毒病的原因(7)，但一直未被证实。
在许多发展中地区，水源可能被严重污染而水处理方法
却又不够先进和可靠。由于这些因素以及大量处于危险之中
的人，一定要把饮用水看成是肠道病毒通过环境传播的一个 非常重要的潜在媒介。像其他细菌性传染病一样，肠道病毒
在一般直接接触方式之外，也由受污染的食品和气溶胶传播。
2.3
对健康的影响
肠道病毒能引起广泛多样的症候群，包括发摩，发烧， 胃肠炎，心肌炎，脑膜炎，呼吸道疾病和肝炎。一般来说，
通常是无症状感染，较重症表现极少。但是，一旦饮用水被 污水污染，有两种病会以流行规模出现一一胃肠炎和传染性 肝炎。除这两种传染病外，没有多少流行病学证据表明处理 充分的饮用水与病毒传染病的传播有关。
病毒性胃肠炎可能与各种因子有关。其中一些只是最近 才被鉴定出来阳，从腹泻病人粪便中发现直径为 27--35 毫
微米的小颗粒。其中大多数还未作出化学定性或实验室培 养.病毒性胃肠炎在各年龄组易感人均可发生，一般 24.......72 小时内出现呕吐、恶心和腹泻。最严重的是婴幼儿或高龄老 人，因易于出现脱水和电解质平衡失调，如不及时纠正则有
生命危险。 轻型肝炎只需休息 1.......2 周并限制活动即可，但重症者 40 一
有时可因肝脏损害而致死或造成慢性肝病的结局。重症肝 炎，年龄越大者越难耐受，因而中年以后的病死率激增。原 有恶性肿瘤和肝硬化的人当中死亡率较高 (9) 。
2.4
建议的依据
理论上一个病毒粒子有能力发动感染过程，实际上以志 愿者的研究表明，在某种人工条件下可以用组织培养检出的 1 单位脊髓灰质炎疫苗就能启动感染 (10) 。但吃下去病毒也 子(例如饮用水中的〉并不一定能保证必然与敏感细胞发生
接触。事实上对于大多数人来说，最小感染量似乎比 1 病毒 单位大，最好是用暴露一定病毒剂量人群中某一比例受感染
的统计学概率来表达。但是，涉及到可靠的公共卫生常例， 很难不同意美国国家科学院安全饮水委员会的这个结论 z
"传染性病毒在饮用水中的存在是对公共卫生的潜在威胁， 而至今还没有有效的依据赖以建立出厂水中病毒污染的无害 浓度叫 1 1)。然而这一结论在实际和经济上距离多远依然是个 问题，显然还需要大量的进一步工作。由于缺乏水病毒污染
程度与疾病的相关资料，所以没有据以规定实用病毒学标准 水平的科学依据 (12) 。因此，坚持主张符合水处理标准是控
制病毒借饮用水传播的最合理途径，而这一标准是积多年之 经验在防止介水病毒病的显性感染病例上发现是有效的。这 些处理方法应当符合曾以实验证明在去除或灭活人工加入的
病毒上有效的标准。至今，经常检查饮用水中有无粪便污染
指示菌仍然是饮用水微生物学安全性常规化验的唯一实用而 又经济的办法。
-41-
2.5
病毒检验方法
近年来，从水样中浓聚病毒的方法发展很快，但其可靠
性、检出限度和精确度尚未很好建立。对于未处理水源水有几 种方法可以使用，例如在((国际卫生组织水污染控制分析手 册 }}CI3) 上记载的方法。对于出厂水， ((水与污水检验的标准 方法泸 14) 中记载的方法主张用大容量样品的病毒浓聚方法。
因为水样的病毒评价至少需 2 周才有结果，这段时间中该被 检水早已被分配并被吃掉了。必须强调的是，饮用水的病毒 检验不应排斥细菌学监测或其他质量控制措施，例如对出厂 水的卫生监督和物理化学测定，而是应当看做它的扩大。非 常希望在做病毒检验的同时还做大肠菌群检验，需要时加大 水量，如果查出有大肠菌群，不必等待病毒学结果便应采取
治理措施。 一般来说，污水中粪便细菌的数量比病毒多得多。从而
产生这种希望，适当的粪便污染指示菌也能对各种水体中有 无病毒起指示作用，大约也就不必直接检查病毒污染了。然
而通，常评价饮用给水安全所用的指示菌，其对环境因素以及 对水与污水处理过程的抵抗力比病毒的低。其结果，肠道病 毒可能存在于细菌污染迹象不大或没有的水中。 有些城市正在考虑污水回收再供生活使用的方案，而另
一些城市的饮用供水则是从受很大比例污水污染的地表水源 取水。不论哪种情况，一定要认真估计到病毒穿过水处理过
程(包括处理前贮存和消毒〉的危险。应进一步考虑用肠道 细菌的噬菌体做为肠肩毒潜在存在指示菌的可能性。与检查 肠病毒相比，细菌噬菌体试验的速度和经济使这一建议很吸 引人 E152.
_. 42 -
2.6
对阳性结果的解释和评价
病毒学工作中样品没有病毒或只含有极少病毒时，采样 和实验室检查中意外污染的危险是非常现实的问题。故应特 别注意，并最好将饮水样品检验工作与其他病毒材料分开进
行。应当建立实验室质量控制体系，并鼓励参与工作的实验 室将其方法标惟化。 按规定，病毒学检查时应经常从同一地点分别采取两份 样品。其中一份应尽可能立即检验，第二份则贮存于冷处。
假如第一份样品结果阳性，则应考虑有无意外污染的可能， 如果认为这种考虑是必要的，则应对第二份样品做病毒测
定。肯定是病毒后，则应尽可能做种型鉴定。对于阳性结果，
应协同公共卫生当局评价其对健康的可能影响。 参考文献 1. WHO Technical Report Series , No. 639 , 1979 (Human
vi ruses in water , wastewater and soi 1. report of a WHO Scientific Group). 2. Akin , E. W. et a l. Enteric viruses in ground and surface waters , a review of' their occurrence and survival. In: Proceedings
0/
the 13th Water Quality
Con/erence , University of Il linois , 197 1. 3. Buras , N. Concentration of enteric viruses in wastewater and effluent , a two-year survey. Water
research , 10:295~298 (1976). 4. Miele , R. P. Pomona virus study-/ina l. reþort. Los ‘
Ange!es , Sanitation Districts of Los Angeles County , 1977. 5. Slade , J. S. Enteroviruses in partially purified water. 43 一
Journal 01 the Institute 01 Water Engineers and Scienlists , 31 :219~225 (1 977). 6. Craun , G. F. & McCabe , L. 1. Review of the causes of w 且 terborne
disease outbreaks. lournal 01 the 65:74~84
American Water Works Association , (1 973) .
7. Berg , G. Virus transmission by the water vehicle. 1. Viruses. H ealth laborator y science , 3'86~89 (1 966).
8. Madeley , C. R. Viruses in the stool8. lournal 01
clinical pathology , 32: 1~10 (1979). 9. National Research Council. Proceedings 01 a
Symþosium on Viral Hepatitis , Washington , DC , National Academy Press , 1975. 10. Plotki 口，
S. A.
& Katz , M. Minimal infective doses of
viruses for man by the oral route. In: Berg , G. , ed.
Transmission 01 viruses by the water route , New York , Interscience Publishers , 1967. 1 1. National Research Council. Drinking water and
health , vo l. 1 , Washington , DC , National Academy Press , 1977. 12. Gamble , D. R. Viruses in drinking-water: Reconsideration of evidence for postulated health hazard and proposals for virological standards of purity. Lancet , 1 :425~8 (1 979). 13. Suess , M. J. , ed. Examination 01 water lor pollution
control , vol. 3 , Oxford , Pergamon Press , 1982. 14. Standard methods lor the examination 01 water and wastewate 勺 15th
ed. , Washington , DC , American
Public Ee J! th Association , 1980. 15. Kott , Y. et a l. Bacteriophages as viral pollution ;ndicators. Water research , 81165~171 (1 974).
第二章生物学指标
1. 原生动物 1.1 概述
对人致病的肠道原生动物中有三种可介饮用水传播 z 即 溶组织内阿米巴，贾第虫和结肠小袋虫。这些生物分别是阿 米巴病(阿米巴性荆疾人贾第虫病和小袋虫病的病原体， 这些病都育与饮用水有关的流行 CI-4)。许多通常自由生存的
阿米巴，例如纳氏阿米巴、哈民阿米巴和棘阿米巴，可能是
介水病原因子，常引起致死性疾病。然而关于这些生物的介
水传染几乎总是和水的娱乐性接触而不是与饮用有关。 溶组织阿米巴全世界都有分布，其存在有滋养体和包囊 阶段两种形式。感染是由于吃了大小为 10.......20 微米〈平均
12 微米〉的包囊而发生的。因为溶组织内阿米巴主要是灵
长类的寄生虫，人是感染的贮存所。病疾病人只排出滋养 体，在外环境中它对干燥以及温度和盐类浓度变化敏感，而
且大部分或全部处于这种活泼的阿米巴样阶段的寄生虫都会 被胃液破坏 (5) 。随后的慢性病例以及带虫者都排泄包囊，则 为更重要的传染源。全世界的多次调查表明，溶组织内阿米 巴庸的流行率为 0.8.......50% (6) 。流行期间的带虫率估计高达
63%(3) 。平均每名带虫者每天排出的包囊数估计为 1. 5X
10 7 J 污水中包囊的密度，如以流行期间带虫率为 50% 计， 估计每升有 5000 个包囊 (3) 。
贾第虫全世界都有分布，是以滋养体和包囊形式存在的 46 一
鞭毛虫。除人外，还曾在许多种哺乳动物 (7 ， 8)和鹦鹉川中发
现过贾第虫。感染给人的贾第虫，曾称为兰伯氏贾第虫，肠
兰伯民鞭毛虫，或肠贾第虫。除小鼠中出现的鼠贾第虫外， 各种动物中发生的各种贾第虫之间在特征、形态或其他方面 都无法区别。过去认为此虫有高度的宿主专一性，如今已有 疑问 (10) ，最近研究 (8)表明其他动物可能起到对人的感染宿
主的作用。和 j容组织内阿米巴病一样，本病也是吃包囊发 生的。滋养体，只有在急性水样泻的粪便材料中才能查到， 相信在动物宿主体外环境中它不会存活很长时间。包囊呈卵
形，长 8~12 微米，宽 7~10 微米。全世界曾进行过多次调 查，人贾第虫感染的流行率为 2.4~67 .5%(6) 。此病在儿童
中比在成人中更多。无症状带虫状态很多，但与显性病例之
比并未确定。曾计算过一个受感染成人每天排出的包囊平均 为 2.1 X 10 8 至 7.1 X 10 8 。在人群流行率为 10~25% 时，未 经处理的生活污水中包囊的密度，估计为每升 10000~ 24000 个 (1 1)。河狸 (Castor canadensis) ，至少在美国
一个居民点饮用水流行中已被怀疑为污染源 (12) 。
结肠小袋虫是一种全世界广泛分布的有纤毛生物，其 滋养体和包囊阶段都能使人感染 m。包囊呈球形至卵圆形，
直径 40~60 微米，色调呈带黄至带绿色，包囊壁有双层膜。
人的感染一般是由吃受病猪粪便材料污染的食品或水而引起 的 (5) 。其他宿主，包括较少几种灵长类，狗与鼠罕见。结肠
小袋虫在猪十分普遍，调查表明流行率达到 ~100%(7) o '二人 的流行相当少，全世界 12 次调查为 0.77% 。人可能有无症 状带虫感染。
致病性纳氏阿米巴，已知是原发性阿米巴性脑膜脑炎最 常见的致病因子。纳氏属的阿米巴有滋养体、鞭毛虫及包囊
-
47 一
三种存在形式(1 3) 。滋养体一般呈弹子状或梨子形，大小为
8"""'14 微米。大多数有一个中心核，但偶尔也出现双核或多
核。靠简单的二等分裂繁殖。将培养基用蒸馆水槽释，会使 滋养体变成鞭毛虫阶段。鞭毛虫运动活泼，呈梨子形，前端 有 2--4 根鞭毛。包囊呈环状，直径约 8 ......， 12 微米。致病性 和非致病性纳氏虫，将培养的阿米巴给小鼠鼻内灌注即可加
以区别。
1. 2
接触途径
凡来自病人含有活的原生动物的粪便材料，不论以何种
机制到达人口，都可出现致病性肠道原生动物向人的传播。
1. 2.1
饮用水
由于人是溶组织内阿米巴感染的主要宿主，给水受到生
活污水的污染可导致该生物介饮用水的传播。追踪到污水污 染饮用水的流行曾有报告 (3) 。热带地区介水传播的危险要大
一些，其带虫率常超过 50% ，相形之下温带地区在全人口中
的流行率一般不到 10% 。包囊在 0 "C 71<中能生存几个月， 30 "C时 3 天， 45 "C 30 分， 50 "C时 5 分钟 E52 ，并且在已知病原体 中是对氯最有抵抗力的一种 (14) 。贾第虫病的介水流行，首先
是美国报告的。在美国，介水流行中鉴定过致病因子的只有 50% 左右。 1972 ......， 1977 年间，在致病因子得到过定性的流
行中，兰伯民贾第虫是最常见的病原 F
自 1965 年以来，贾
第虫病的介水传播在美国共报告 23 起(1)。去苏联的旅游者 中发生的流行，其传播媒也怀疑为饮用7]< (15 ， 16)。大多数流
行都与未经处理或只受消毒的饮用水有关 (2)。有人指出具感
染量很小， 10 个包囊盛在明胶胶囊使人吃了就能引起感 48-
染 07 ， 18) 。此种生物在环境中存活以及对消毒剂的抵抗力的
资料，迄今尚不完整。但有迹象表明，冻结之后还能活下来
的包囊只有百分之几，而这种包囊在 80 0C 的饮用水中却可以 维持生命达 77 天 (19) 。初步研究表明，贾第虫对氯的抵抗力 处于溶组织阿米巴和肠道病毒之间 04 ， 20) 。考虑到贾第虫在
人以及各种家畜和野生动物中的分布，以及这种生物并不像
从前相信的那样有特殊的宿主，大概贾第虫在讨论的三种肠 道原生动物中是最有介饮用水传播危险的。
关于小袋虫病唯一的一次介水传播发生于 1971 年密克 罗尼西亚的特鲁克群岛 (4) 。报告的结论是，这次流行大体上
是由于给水受粪便污染所致，当时一场破坏性台风摧毁了猪 圈和不牢固的汲水设施。 问米巴性脑膜脑炎的传播记录很少见于用生活给水沐浴 的人中，曾有一例发生于用水洗鼻子之后。
1.2.2
食晶
溶组织内阿米巴包囊可通过食品传播，而生蔬菜可能是
传染源 (7) 。尽管该包囊在手上极少存活超过 10 分钟，除非 藏在指甲之下，但食品操作人员如是带虫者则会是本病的重
要传染源。从委内瑞拉的一次研究中得到了这种接触途径的
一个直接的证据，当时对食品操作者的洛组织内阿米巴病进 行了检查和治疗，使 3 年中每年阿米巴荆疾的发病率由
36.84%。降低至 0.61%0 (7)。另一次在中国的研究关系到吃用 手掰过的冷面包引起的传播，其结论是 z 经食品操作者传播 大概是比其他传播途径更为重要。食品也会受家蝇污染，巳 查明，蝇粪能传播此生物 (5) 。
只报告过一次贾第虫病的介食品流行 (2 1)，据信制作该 49 一
食品的病人是传染源。虽然草莓 (22)和蔬菜 (23) 上都查出过贾
第虫包囊，但介食品传播在贾第虫病流行病学上的意义还不
清楚。 Levine(7)指出，小袋虫包囊在猪粪中只要没有干透便 可存活数周，再者，食品一且受到病猪或病人粪便材抖污染
就会发生介食品传播，这似乎合乎逻碍。
1.2.3
空气
肠道原生动物病原体通过空气直接传播给人的似乎不会 有，因为它们对干燥灭活敏感。尽管向肺和胸膜扩散是未治 疗的阿米巴性肝肤肿病人的并发病，但还无人报告过原发性
呼吸道感染。即使吸入污染空气继而咽下粘液吸附的虫体能
造成感染，恐怕这种途径的重要性也比人与人、介食品或介 水传播要小几个数量级。
1.2.4
其他接触遮径
性传播的贾第虫病和溶组织网米巴病曾有过报道 (24 ， 25) ，
特别是在同性恋者当中。 Freeman(5) 说过，虽然游泳池还
不肯定是阿米巴病的，但却是溶组织阿米巴感染的一个潜在
来源。娱乐用水做为贾第虫病传播途径的意义尚未肯定，但 大概不大能和别的接触途径相比。小袋虫病除一次介饮用水
流行外，唯一的另一次流行发生在精神病医院，其原因是个 人卫生不良以及病人之间的色情行为造成的人与人之间的传 播 E42.
1.2.1
接触途径的相对意义
三种肠道原生动物病原体中，全世界流行最广的是溶组 AU
织内阿米巴。在美国，人对人传播是最常见的传播方式 (26) 。
在世界的其他部分，由有病的食品加工人员引起的食物污染
是最重要的传播途径的。介水流行的危险是存在的，而且发 展中国家很可能由于介饮用水传播才保持地方流行性感染水 平，虽则绝不是说这一点已十分清楚。 儿童和成人之间贾第虫病的传播，在个人卫生→惯很好
的地方是少见的。然而学龄前儿童之间感染的传播，在日托 托儿所 (27)和类似机构中大概是常见的。 Burke(23) 曾概括
说，许多国家的地方性感染发病率与社会经济水平呈反比， 并发现卫生状况处于标准线以下的地方流行得最广。在美
国，介水传播显然是重要的传播途径，但关于其他途径有关 的证据还不足。 人的小袋虫病发病率不高，而直接和猪接触似乎是此病
的主要传播途径。受猪粪污染的食物和水都在传播此生物的 危险。 几乎阿米巴脑膜脑炎的所有病例都由娱乐用水而不是由 生活用水引起的，但发展中国家的流行病学证据非常少。
1. 3
对健康影响
虽然大部分溶组织内阿米巴感染后没有症状或只引起轻 微症状，还是有死亡 (26) 。在临床上从轻型腹泻到暴发血性病
疾的症状，都诊断为胃肠炎。肝肤肿是最常见的转移合并
症。致病性似取决于株毒性和宿主因素，包括个人营养状态 和有关细菌群系 (7) 。
贾第虫病症状的范围，从轻型可自愈的肠炎到慢性使人 衰弱的腹泻和无症;快感染都有 (28) 。急性、亚急性;相慢性阶 段的感染都有记载 EmN312.
-
51-
贾第虫病引起的死亡，不论人或其他动物的都没有报告 过，只有鹦鹉可能是例外 (9) 。
小袋虫病可能以急性血性病疾的形式出现 E53 ，但也有人 出现无症状带虫阶段 (7) 。人对感染的抵抗力很高，即使发病 一般也很轻而且可自愈 (0 。
福氏纳氏虫阿米巴是一种在土壤和淡水中广泛分布的阿 米巴鞭毛虫 (32) 。近十年来已查明不论热带或温带国家中致 死性脑膜脑炎都有上升趋势，记录在案的已达约 100 例 E333 ，
一般都是曾在天然水中游泳过的人。侵入门户是鼻咽腔，阿
米巴从这里穿过嗅觉粘膜和筛板而进入脑内。对于这种生物 的认识仍然不完整，但它显然以自由生活为主，可能在温水
贮存期间繁撞，虽然人的感染与接触可能感染者的关系极 少，但临床病例的预后极坏，而且难以治疗 C3430
1. 4
监测
不主张搞监测计划，因为没有定量或定性的标准方法。
关于浓集和检出语组织内阿米巳和贾第虫包囊的实验方法是 者的 (11 ， 35) ，但主张只在有地方性流行和具体的流行病学研
究时使用。目前者的方法各有缺点 ，
j在集技术没有重现性;
浓集样品中的虫体很难鉴定，至少贾第虫时是如此，检出包 囊的生活力和来源也无法断定 (11 )。此外，关于采样频率无
法推荐。
当饮用水受致病性肠道原生动物污染而造成疾病流行 时，水的煮沸可以有效控制贾第虫 C192 ，溶组织内阿米巴和
结肠小袋虫，使其灭活。应试图进行污染源的鉴定和消除。 应进行卫生监测以鉴定和娇正处理系统和配水系统的故障。
应收集流行病学材料与有关该处理厂和配水系统物理化学资 一 52 一
料的相关关系。已有关于介水疾病的调查大纲 (36) 。为预防
或控制今后流行，收集这类资料是必要的。
1;5
建议的依据
肠道致病性原生动物做为一个群，在人和各种家畜和野 兽等受感染个体的粪便中大量出现。人受兰伯氏贾第虫的感 染量至少少到经口吃进 10 个包囊，据推测，贾第虫和溶组织 内阿米巴都是只一个活包囊就能使敏感宿主开始感染。 7 ， 37)。
还没育可供饮用水中肠道原生动物检出、鉴定和计数的简单 方法 (1 1)。有鉴于此，水源防护，充分处理，以及保证配水 管网水的完善是预防感染的最有效途径。 在处理过的水中，大肠菌群不是贾第虫和 j容组织内阿米
巴的好指示菌，因为这类原生动物对消毒灭活有很强的抵抗 力。在未消过毒的水中，指示菌的存在可以指示致病性原生 动物的存在。早期研究表明，硅藻土过滤可以有效地去除溶 组织内阿米巴和贾第虫的包囊，并表明，以粒状滤料也能有 效地去除海组织内阿米巴 C38) 。硅藻土去除贾第虫包囊可达
99.998% ，近来的研究ι39)表明，使用粒状漉料并注意混凝剂 量、过洁、速度、浊度以及反洗方法，则可能去除贾第虫包囊 99.99% 。
如果要从地表水和无防护的地下水中去除原生动物，建 议把大肠菌群不适于做这群生物的指示菌、缺乏原生动物的
标准检验方法、原生动物对消毒的抵抗力强、以及在去除致
病性肠道原生动物上正确运用过滤的巳证实的效果等等都做 为考虑混凝、沉淀和过憾的依据。
一 53 一
2. 蠕虫 饮用水中曾查出过多种多样的蠕虫卵及其拗虫，要得到 安全的饮用水，显然所有对人有传染性的一切都应消除。然 而绝大部分蠕虫主要不是介水传播的，因而在常规上水的蠕
虫监测既无可行性又无必要性。与给水较有直接关系的蠕虫 有两类:完全靠吃入已感染的挠足类中间宿主才能传播的
(第一组〉和其尾拗直接传染给人的〈第二组〉。其余大部分
品种均可归入第三类(第三组〉。有关系的蠕虫，按正式分类 划分分别属于两个动物门 z 即线形动物门 (Nemathelmi (Pla tyhelminthes)
nthes)
fl.[J圆虫类，
和扁形动物门
即扁虫类。扁形动物门中有吸虫纲〈吸虫〉和多节练虫纲(练
虫〉。有介饮用水传播危险的蠕虫名单见下表。
2.1 2.1.1 概述
第一组〈龙线虫、叠宫练虫〉
第一组包括在水生挠足类甲壳动物中发育的线虫，并要 求为人饮用的水含有中间宿主甲壳动物。这一组中最重要的
成员是麦迪纳龙线虫，即几内亚蠕虫，是人的一种寄生丝 虫。雌虫在组织深部成熟，尔后游走并寄居于皮下。许多拗 虫在雌虫体内发育，引起皮肤服肿和渍烂。雌虫将其脱出的
子宫在此处暴露，一旦感到布水时即将拗虫排出。如果这种
杆形拗虫进入的水中含有能吞食幼虫的剑水蚤、真剑水蚤、中 ← 51 一
有介饮用水传播危险的蠕虫 类别
名称 大片吸虫 (Faociola
分组 a 品
w 皿冒EEEEEE
吸虫纲
gigantica) hepatica)
(Trematoda)
肝片吸虫 (Faaciola
布氏姜片虫 (Faaciolopsie buski) 埃及血吸虫 (Schistosoma
haematohium) intercalatum) japonicum) 且
刚果血吸虫 日本血吸虫
(Schistosoma (Schistosoma
EEWEZEEEEE
曼氏血吸虫 (Schistosoma manloni) 据公河血吸虫 (Schistosoma (Echinococcul (Echinococcue
mekongi) granuloBus) multilo-
多节缘虫纲
细粒棘球缘虫 多脏棘球缘虫
(CreBtoda)
cula ril) 矮小膜壳综虫 (Hymenole peis nana) 多头缘虫 (Multice
EETAI-
ps .p.)
曼氏叠宫缘虫 (Spirometra mauoni) 类是民叠霄缘虫 (Spirometra 多育叠宫缘虫 (Spirometra 寨民叠宫缘虫 (Spirometra 猪肉综虫 (Taenia
maneonoides)
proliferum) theile ri)
lo 1i um) hraoilienae) doude冒血冒
钱虫钢
巴西钩口线虫
(Ancylostoma
(Nematoda)
十二指腾钩口线虫 (Ancylostoma
nale) 人翩虫 (ABcari.
lumbricoideB) medinenBi.)
麦地那龙线虫
(Dracunculus
I
冒且冒
镜虫 (Enterobiu.
vermicularis) americanuB) ItercoraliB)
美洲板口线虫 (Necator 粪类圆线虫
E
冒 EE 皿
(Strongyloidee
弓姻虫
(Toxocara
.PP.) trichiura) Btenocephala)
w 皿
毛首鞭虫
(TrichuriB
w 且
狭头钩虫
(Uncinaria
·注意，除龙线虫外，其它线虫实际上极少主要集饮用水传播。
-
55 一
刽水蚤和大剑水蚤时，则生活环就能连续下去。助虫在甲壳
动物体内发育成第 3 阶段助虫，一旦被人吃下去就能使人感 染 C40
其次，没有防护的水井和水池 t~ 管道给水尤能传播龙线 虫。这种病不均匀地分布于非洲，尤其是西非萨赫勒，西 亚，以及远东阿富汗，印度和苏联南部，以及印度尼西亚。
人是主要的最终宿主。 叠宫练虫属的综虫，虽然人极少感染，但也有一个在水
生挠足虫体内的阶段。成虫在猫的小肠中发现。叠宫绪虫卵 随粪便排出，在水中孵化产生钩毛助，被挠足虫吞食后发育 成前尾拗阶段 (41 ， 42) 。如果人吃掉该挠足虫，则拗虫在组织
中进一步发育成全尾助或裂头助阶段。在亚洲的一些地区， 常用做伤口擦药的另一种中间宿主的肉也能将此拗虫传染给
人。几种叠宫绪虫散在分布于从美国到乌拉圭之间的美洲大 陆、东亚 (44 ， 45) 、肯尼亚和坦桑尼亚联合共和国 (43) 。
2. 1. 2
接触途径
含有感染挠足类的饮用水是龙线虫感染的唯一来源，而 龙线虫是只靠供应安全饮用水就能根除的唯一动物寄生虫。 龙线虫是偶然的人体寄生虫，除饮用水之外还有其他接触途
径。
2. 1. 3
对健康的影响
几内亚蠕虫病，即麦地那龙线虫病〈龙线虫病〉是萨赫
勒和印度次大陆尤其是农业地区的一种致残性主要疾病。感 染的疼痛和距离虫最近关节的关节炎使病人数周内不能活
动。由于此虫的成熟恰好与多雨的种植季节一致，劳动力的
-
56-
丧失影响生产力减少。伴随雌虫游动的深刻变态反应，包括 红斑、尊麻彦和强烈搔痒，同时可包括呕吐和腹泻等全身症 状。7]<.泡会发生化服，试图拔除时虫体的破渍可发生溃痛。 纤维化有时是一种结局，而虫体异位也能导致其他内脏中形
成肤肿。龙线虫病是公共卫生上十分重要的地方病，局部地 区发病率可超过 30% 。
裂头拗病是更少见的传染病。此时皮下的拗中可造成损 伤而引起水肿和炎症，特别在该寄生虫趋于死亡之时。如果
侵犯到眼睛，则后果严重，这是特别发生于东南亚的一种并 发症。
2. 1. 4
监测
龙线虫病的问题出在无管道的小型给水〈例如阶梯水井 和蓄水池) ，而这种给水往往无法做定期监测。这种情况的调
查涉及到人感染的流行率和使用浮游生物网或其他容器采集 挠足虫，然后做显微镜检查去发现寄生性微丝拗。因为预防 是通过水源防护〈见后〉达到的，监测的最好方法是确定水
源有无防护。监测叠富练虫既不可行也无必要。
2. 1. 5
建议的依据
含有一只幼虫的一只感染挠足虫就有能力使人感染裂头 拗或龙线虫，尽管该蠕虫侵袭力取决于吃入感染性幼虫的数
目及其性别。因为一个成熟的成年雌性几内亚蠕虫就能引起 严重疾病，应将其感染性阶段自饮用水消除。正因为这是龙 线虫向人传播的唯一途径，所以是重要的。从抨状拗到达挠
足虫的途径来看，是从汲水者肢体冲洗落入井水所致，显然
水源防护是最好的预防方法.使用高于地面的井台并从这种 En· η'
井打水一般即可，但最好加用井盖并接上水泵。在应急状态
下，向井中投放双硫磷颗粒使达到控制昆虫幼虫所需剂量， 便可杀死受感染的挠足虫 E462.
2.2
第二组〈血吸虫、钩虫、 板口线虫〉
2.2.1
总述
第二组包括一群性质不同的吸虫和线虫，其感染性幼虫
都有能力穿透人的皮肤和粘膜。因而能通过饮用水传播，但 用这种水洗涤和沐浴时害处更大。主要的种属是血吸虫。
传染给人的血吸虫属于三个主要品种 z 埃及血吸虫，侵 袭内脏静脉丛，主要发生于非洲和西亚，曼氏血吸虫，发生 于非洲，部分南美和中美洲，以及加勒比海的几个岛屿，而 日本血吸虫则发生于中国、印尼、菲律宾、以及东亚的其他部
分。曼氏血吸虫和日本血吸虫都是在门静脉系统中发现的。
刚果血吸虫发生于中非西部，而-种与日本血吸虫类似的泪 公河血吸虫则发生于东南亚酒公河盆地。目前已知，每一主
要血吸虫种，实际上都有不同的株，其地理分布和宿主均有 差异。 成虫寿命长且雌雄异体。有些虫卵穿过血管壁和组织而
从尿〈埃及血吸虫〉或粪便(其他血吸虫〉排出。虫卵抵达 淡水孵化释放出毛拗，毛助穿入适当的水生螺宿主并在一个
月或更多时间内经受发育和繁殖，此后尾助，即能感染人的 幼虫，便被排入水中，大概以其分歧的尾部游动。尾助刚好
可以用肉眼蕾到。与感染水体接触时，尾拗迅速穿入人的皮 肤并在人体内移动和成熟，完成其求育环. 58 一
非人类血吸虫和其他有关吸虫的尾助，也企图穿入人的
皮肤但在皮中死亡，从而产生红斑和强烈剌激，是称之为血 吸虫性皮炎。日本血吸虫感染许多种家畜和野生动物和人。 人类钩虫〈十二指肠钩虫和美洲板口线虫)部分布于热 带和亚热带，虫卵在土壤中孵化并发育至第三阶段拗，此拗
能穿入皮肤使人反复感染。家畜的钩虫也能侵袭人，但其幼 虫在皮肤中游动从而产生局部症状。
2.2.2
接触途径
血吸虫病是日常生活、沐浴或洗涤时使用传染性水而感 染的。吃进去的尾拗能穿入口腔粘膜，但这是次要的侵入途 径 (47) 。与饮用水的相关性在于用饮用水洗涤(如果易于获
得) ，而安全水的效益只有减少与从前所用感染源的接触才 能显露出来。
至于有管道的未处理地表水传播血吸虫病的可能性很
小，大部分传播都来自无管道的水源，如水塘、水井，以及 宗教沐浴用的水池。 血吸虫性皮炎是娱乐用水和职业用水而不是使用饮用水
的恶果。已知钩虫幼虫在饮用水中有感染力，这是一个可以 估计到的但不是主要的传播途径剧，492.
2.2.8
对健康的影响
人类血吸虫病发病极广，全世界受感染的人达 2 亿，并 时有死亡。病理主要是宿主对未排出卵的反应。原发损伤主 要在肝脏、小肠和膀胧周围，但最严重的续发症是后尿道的 二次损伤、膀胧癌、肝硬化及其血液动力学续发症。
钩虫主要导致缺铁性贫血，而本组中其他蠕虫则产生皮 一 59 一
肤损害。
2.2.4
监测
检出水中的血吸虫尾拗还是一个不适于常规监测的研究 方法。可以用过熄法从水中浓集尾拗，也可以将适当的小白
鼠漫泡在水中待发育一定时间后进行解剖的方法检出。滤 纸、玻璃纤维惊器或醋酸脂滤膜上的尾拗常丧失原有形态， 所以做种属鉴定不大可靠，尽管设计过许多鉴别染色法。可 行的方法是寻找适当的脊椎动物宿主，并测定具感染率。
2.2.5
建议的依据
因为一个尾拗就能感染，因此应从饮用水消除尾拗，而 且居有安全线。由于没有常规监测方法，如果怀疑某地区饮
用水有较大危险，可靠的办法是放在预防措施上。尾拗自由 活动的寿命不超过 48 小时，所以把水贮存 48 小时即可表示 安全 (50) 。似乎贮存 24 小时会大大减少感染力。慢砂撼，只
要正常运转就能去除大部分尾助，而且加氯消毒 1 小时后游 离余氯水平为 0.5 毫克/升即能杀死人类血吸虫尾拗 (5 1)。比
较合理的方法是启用不含有宿主螺聊的水源并不使受粪便污 染。
2.3
第三组(捆虫、鞭虫、 膜壳综虫、棘球拗〉
类圆线虫、挠虫、片吸虫、
2.3.1
总述
大多数蠕虫的虫卵抵抗力都强，或包囊对人有感染力。 60 一
如果它有机会进入饮用水并被吃掉，人就会受到感染。 55 页
表中本组名单列入所有各伸，其他传播方式如食品和直接的 粪-口途径都比饮用水途径重要得多，从实际效果来看，粪便
处理是非常重要的，而仅仅注意饮用水将不会使世界大部分 地区的蠕虫病得到显著减少。 传播极广的肠道蠕虫(人蝴虫和毛首鞭虫〉都产生抵抗
力强的虫卵，其特征在于其胚胎一定要在外环境中发育到它 有感染力为止。虫卵一旦被人吃掉，幼虫即在肠道中出现并 经历复杂的迁移，然后回到肠中成熟并以每天近 20 万个的速 度产卵，卵随粪便排滥，因为很重，很快即沉入水底。
不太普遍的是粪类圆线虫和烧虫。粪类圆线虫的幼虫是 有感染力的形态，烧虫则更适于直接的粪口传播，因其虫卵 枯而少弹性。此外还有一些能在人体有限发育的动物线虫。 片吸虫和姜片虫属的肝吸虫能感染人和其他哺乳动物，
它在螺聊中发育并形成尾助，然后在水生植物上形成包囊，
通过吃或啃这种植物即可使人感染。包囊有可能进入饮用 53 尸 UK A雏 俨、
是吃了一般从狗来的虫卵而感染的，都有由饮用水传播的危 险。
2.3.2
高相对密集的。尽管蠕虫卵特别是一般捆虫和鞭虫的虫卵不 时地进入水中，但饮用水绝不是主要传播方式。
寸/
膜壳练虫属的人练虫有直接的生活环 F 而棘球拗属，人
这些蠕虫全都有粪-口传播，而感染性阶段倾向于是相当
。
接触途径
-
61 一
2.3.3
对健康影晌
肠道蠕虫产生的症状极广。许多感染是亚临床型的，个 别是致死的。似乎大多产生低度慢性、在个人而非集体意义 上难以估量的作用。营养从宿主转向喂育虫体是显著的。肠 道蠕虫以其极高流行率 (52)弥补了它们在许多受感染个体产
生的有限病理改变。
2.3.4
监测
在流行时或进行研究时，可借过滤将蠕虫卵和幼虫从饮
用水中浓集出来，大部分虫卵可用显微镜进行鉴定，然而这 些方法不适于常规使用，而水污染的频度也不宜于这样做。
2.3.5
建议的依据
单独一个成熟卵、成熟幼虫或形成包囊的尾拗都能引起
感染。因此应从饮用水中消除它们，而最好的方法是保护水 源免受粪便污染。如果它一旦进入原水，大部分会被过滤特 别是被慢砂谑所去除，然而却全都是对氯有相当高抵抗力 的，尤其是烟虫 (53) 。
一 62 一
3.
自由生活生物 3.1 概述
给水中一般认为重要的自由生活生物有，浮游生物和大 型无脊推动物，而本节将细注其中几群。浮游生物系由主要 在水体中以悬浮状态生活的显微镜生物或小生物所组成。浮
游植物包括自由生活的细菌、真菌和藻类。藻类是有叶绿素 的生物，有各种色泽。它们是自营性单细胞或成团的，有运 动或无运动。真菌和细菌大多为异营性。浮游动物的组成有
自由生活的原生动物、轮虫类、枝角目、挠足类，蠕虫以及 处于发育的早期阶段或短短的旅行期间的某些水生昆虫的幼
虫和鱼。大型无脊椎动物包括较大的底栖动物，如水生昆虫 幼虫、甲壳类和腹足类动物。
给水中浮游生物是重要的，因为它干扰水处理过程、产 生对人有毒性物质，包藏对人的病原体，并促成加氯消毒期 间转化成卤化有机化合物的有机物质。大无脊椎动物可影响 配水系统的效力和用户对水的接受。
藻类可能由于在未处理过的地表水以及盛处理过(出厂〉 的露天贮水池出现而发生问题。除这些位置外，浮游动物还 能在大口井中出现。与给水问题有关的大型无脊椎动物的出 现，仅限于配水系统。
地表给水中藻类的聚积和品种组成受天然环境条件的组 合所支配 E542. 地表水中藻类的最大数量可由每毫升中只有
一 63 一
几个到多达百万计。地表水中的藻类最常以硅藻占优势，其次
为绿球藻和蓝绿藻。冬季一般以羽状硅藻占优势，而夏季则 中枢硅藻、绿球藻和蓝绿藻最多。晚夏或全年温度保持在 25 "C以上的地方，蓝绿藻常占藻类总体的大部分 (55-58)。关 于给水中藻类的广泛研究已由 Palmer 发表 (59) 。
在给水分配系统中已发现种类极多的自由生活生物，其 中包括真菌、藻类、原生动物、轮虫类、蠕虫类、水蚤类、 虾类和其他甲壳类、弹尾类、蝴幼虫、淡水蚌、水螺。关于 它们出现的原始报告，是以听到用户抱怨水龙头上出现了活 物之后的调查为依据的。最近的系统研究巳揭示出许多管网
中有生物存在，甚至有些管网是意想不到会有这种东西的。现 在相信，大部分系统会含有某种动物，而 Collingwood 已编辑出一份善于侵袭某些主干水管的生物名单 (60)。
3.2
对健康影响
有肯定的证据证明，某些藻类在给水中产生某些有毒物
质，具有对公共卫生十分不利的影响。有毒品种多出于二群 主要藻类之中 z 黄绿藻 (Xanthophyta) 和蓝绿藻 (Cya nophyta) 。有毒蓝绿藻在淡水给水中最为重要 (61 阳6430
藻类释放的有毒物质可以穿过处理设施。在实验室实验 中，明矶混凝、过洁、和加氯消毒起不到消除藻类毒素的作 用。即使使用相当于水处理时剂量的活性炭来处理也未成 功 (65) 。
由于藻类出现的公共卫生问题是十分有限的。高浓度的 蓝绿藻与人类胃肠炎流行之间的相关关系，在印度 (69) 、菲 律宾 (68)和美国 (66 ， 67) 已有报告。已从几种蓝绿藻分离出类似
革兰氏阴性菌内毒素样的物质 E702.
-
64 一
美国宾夕法尼亚州塞维克莱市发生的波及约 5000 人的 介水胃肠炎流行是迄今最严重的一次 (71 ， 72)。它与出厂水开
放式贮水池中最常见的丝状蓝绿藻 (Schizothrix cola) 的泛滥有关。
calci-
藻类及其细胞外毒素可以代表地表水给水中大部分有机 物， 76) 。
并可能是氯消毒时产生三卤甲皖前体的重要来源、 (73帽
曾报告主要在温带国家水中出现过的自由生活动物，虽 还不知道它能引起什么疾病，但有人指出在实验室条件下它 能吞食并保护致病性生物 (77)。就地贮存生活用水的地方一
定要加盖，防止蚊类病媒在贮水容器中草生。
3.3
干扰处理或扰害配水系统的
其他影响一一臭与味，色度与浊度 藻类在给水中会在感观上令人不快，并且增加耗氯量的 结果会干扰水的处理，造成昧与臭的问题并堵塞滤器剧， 78)。 不论水生植物或动物的生长，都会构成昧与臭的天然来
源。在地表水中构成主要问题的是藻类，而动物类可以在地
下水、贮水地或管线中繁殖。有几种藻类分泌油脂，这些油 脂或是代谢活动中或是死细胞崩溃时释放出来的。这些油脂 使水发生典型的昧和臭 (79 ， 80) 。
当贮水池开放或处理不充分时，色度和浊度会成为无管 道而处理过的水中的一个问题 E8130 此外，藻类及其他生物
的生长，可由于培塞 i虑器而干扰水处理系统的维护和运 转邸'832.
-
65 一
3.4 物时， 84-89) 。
监测
有几种适当方法可供采集和分析给水中的自由生活生
3.4.1
采样
藻类及其他微生物会因经受快速而暂短的改变，其数量 和品种组成逐日都有显著不同 (90)。如果微生物量和品种组
成等资料能够有效地用于改进水处理过程、控制臭和昧，并 检出出厂水中微生物有害浓度的话，则应经常测定。
3.4.2
生物量
不同藻类的大小、形状和体积差别很大，单靠计数无法
精确地估价每一分类单元提供的有机物质(生物量〉的量。 生物量的估计通常靠测量细胞表面积、细胞体积，或该藻类 的叶绿素含量来进行邸， 84 ， 85)。
3.4.3
肉毒素
堂阿米巴包囊溶解试验对内毒素是相当简单的，因而使 用者日多 (70 ， 72 ， 9 1)。但此试验无法鉴别该内毒素来自细菌抑
或来自藻类，内毒素浓度和对健康影响之间尚未建立起量的
相关，而试验结果的意义和解释尚有待确定。
3.5
建议的依据
生物在给水中出现会造成对健康不利的影响，感观〈美
学〉问题，令人厌恶的昧和臭，且能干扰水处理。温暖季节 在配水系统中成批出现的生物，且然就目前所知它们对健康 一 66 一
没有什么不利影响，但出于感观的理由希望尽少在用户水龙 头出现。现行的技艺状态不允许建立最大容许限度，但建议
凡有可能的地方都要把自由生活生物从饮用水中去除。这可 以通过水源防护，实现处理良好常规，定期而系统地洗刷管
线以及监测水质等达到。 (吴联熙译〉
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investigate waterborne illness , 1st ed. Ames , IA , IAMFES , 1979. 37. Beaver , P. C. et a 1. Experimental Entamoeba
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removal of Giardia cysts and cyst models. In: lakubowski , W. & Hoff , 1. C. , ed. Waterborne
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giardiasis , Cincinnati , US
Environmental Protection Agency , 1979 (EPA-600/939. Logsdon , G. S. et al. Alternative filtration methods for removal of Giardia cysts and cyst models.
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73:111~118
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4 1. Mueller , J. F. The laboratory propagation of
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Transactions 01 the American M icroscopical Society , 78:245~255 (1 959). 42. Mueller ,1. F. Host-parasite relationships as illustrated hy the cestode S pirometra mansonoides. In. McCauly , 1.丑， ed. Host-parasite relationships , Corvallis , Oregon State University Press , 1966 (Proceedings of the 26th Annual Biology Colloquium , Corvallis , Oregon , 23~24
April 1965).
43. Opuni , E. K. & Muller , R. L. Studies on Spirometra
theileri (Baer , 1925) n. comb. 1. Identification and hiology in the laboratory. J ournal 48:15~23
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45. Faust , E. C. et a I. Craig and Faus t' s clinical
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World Health Organization , 42:563-576 (1 970). 47. Maldonado , 1. F. & Perkins , K. W. Schistosomiasi :s in America. Barcelona , Editorial Cientifico-médica , 1967. 48. Okamoto , K. An experimental study of the migration route and development of Ancylostoma duodenale in pups after oral infection. !ournal o/. the Kyofo
Prelectural Medical University , 7Q:135-152(96 1). 49. Higo , A. An experimentaI study on the migration route and the development of Ancylostoma duodenale in pups after cutaneous infection. !ournal of the
Kyoto Prelectural Medical University , 70:851874 (1 962). 50. Jones , M. F. & Brady , F. 1. Survival-of Schistosoma
japonicum cercariae at various temperatures in several types of water. N ational 1 nstitute 01
H ealth bulletin , 189: 131-136 (1974). 51. Coles , G. C , & Mann , H. Schistosomiasis and water works practice in Uganda. East African medical
journal , 48: 40-43 (1 971). 52. Sto l!, N. R. This wormy world. Journal of
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aquatic communities and environmental stresses , Interstate Commission on the Potomac River Basin , Rockville , MD , 1978. 57. Williams , L. G. & Scott , C. Principal diatoms of major waterways of the United States. Limnology and
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58. Williams , L. B. Possible relationships between plankton-diatom species numbers and water quality estimates. Ecology , 45:809~823 (1 964).
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Algae and man , New York , Plenum Press , 1964 , pp. 74-
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62. Schwimmer , M. & Schwimmer , D. Algae and medicine.In. Jackson , D. F. , ed. Algae and man , New York , Plenum Press , 1964 , pp. 368~412.
63. Schwimmer , M. & Schwimmer , D. Medical aspects of phycology.Ina Jackson , D. F. , ed. Algae , man and the
environment , Syracuse , Syracuse University Press , 1968 , pp. 279-358. 64. Moore , R. E. Toxins from blue-green algae.
BioScience , 27 a797-802 (1 977). 65. Ingram , W. M. & Prescott , B. W. Toxic fresb water algae. The American midland naturalist , 52:75-87 (1 954) .
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public health , 21:198-200 (1 93 1).
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public health , 21 :1227-1235 (1 931). 68. Dean , A. G. & Jones , T. C. Seasonal gastroenteritis and malabsorption at an American military base in tbe Pbilippines. American journal 95 a 111-127 (1972). 69. Gupta , R. S.
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epidemiology ,
& Dashora , M.
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American Water W orks Association , 72: 170~176 73. Miller , S. Drinking water and its treatment (An interview with Professor Sontheimer , University of Karlsruhe). Environmental science and technology , 14:510~514
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American journal of tropical medicine and hygiene. 9'136~142
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Symposium on Biological Indicators of Water Quality , Newcastle , October 1~15 ， 1978 , University of Newcastle , 1978. 8 1. Fogg , G. E. Extracellular products of algae in freshwater. Archiv fUr Hydrobiologie , Beiheft.
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the response of plankton and periphyton to
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Applied and enuironmental microbiology , 37 :928 931 (1 979).
,..."
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78-
第三章与健康有关的 无机成分
1. 硝 1.1 1.1.1 来源
概述
耐存在于整个自然界的环境中，常见的为畔的硫化物以 及肺与很多金属(铜、钻、铅、悻等〉的化合物(1) 0 U申在地
壳中的平均含量约为 2 毫克/公斤 (2) 。
且然肺具有各种各样
的化合价，又以有机碑和无机碑两种形式存在，但是碑在环 境中的含量通常以总肺量计算 (3)。在某些局部地区，由于商
业上生产和应用肺化物，致使该地区环境中碑含量明显高于
自然本底水平。
1. 1.2
水中存在的呻
由于很多碑化物是水溶性的，所以可能污染水体。关于 醉在水中的化学式尚不完全了解，已证明有三价和五价两种 F 还有一些有机碑 C 1)。已发现在新西兰地热水中含大量碑。虽
然碑是普遍存在的，但是水中的大部分碑来源于工业废水， 除了在有些泉水中天然存在高浓度呻以外，通常在工业高度
发达的地区浓度最高 ω.
-
80-
1 ，， 2 .. .2. 1 饮用水
接触途径
大多数给水含碑量很低，远远低于 10 微克/升Cl -3) 。在
特殊情况下，井水遭到严重污染时，水中肺含量可达数千毫 克/升。 3 。关于给水中畔的形态或种类几乎未凡报道。
1.2.2
盒物
一般而言，大多数食物中碑含量低于 1 毫克/公斤干 重(1)。海鱼的碑含量可能高于该水平，贝类动物中呻含量可 t 能明显地超过 50 毫克/公斤 (2， 5) ，海产品中的呻主要是以有
机畔的形式存在的(1)。 值 (3 ， 8) 。
从膳食中摄入畔的量平均约为 30 微
克/公斤体重/日向 6， 73@ 二些国家已采纳了这一较高的估计
1.2.3
空气 弓，、
在非城区和非工业区空气中肺含量很低，一般低于 0.01 微克/立方米(1)。在城区，特别是在工业城市中， 空气中呻
含量超过 1 微克/立方米 El30 典型的空气接触的为 0.2 微克/ 立方米，而且主要是无机碑。
1.2.4 1. 2.4.1
其他接触途径 职业性接触
在职业接触中，特别是在冶炼工作中，可能接触含硝烟 尘 F 碑浓度可能超过 1 毫克/立方米(1)。 曾有报告，冶炼厂 附近的土壤中碑含量高达 380 毫克/公斤 (9) 。
-
81 一
1. 2.4.2
烟草
由于减少了使用含碑喷雾剂，所以虽然烟草中含有一些 畔，但是含量很低 EI ， m-
1. 2.4.3
其他
由于一些药品含有肿，对于某些人来说，随药物服入网
碑量可能很高 F 而对于一般的居民来说，通过这种途径接触 的畔，与从食物、水和空气中接触相比，是微不足道的。智、
1. 2.5
不同接触途径的相对意义
仅就总肺摄入量而论，通常从水中摄入低量畔，与总摄
入量相比所占比例很小，但是当饮水中肿含量为 0.05 毫克/ 升时，从饮水摄入的肺量可能等于或超过非职业接触者从膳 食中的总摄入量。 还无法估计从各种途径吸收的呻量，因为对于碑在食物
和水中的特殊形态了解甚少，而这正是决定肺吸收比例的主 要因素 (2 ， 6) 。
1.3
代谢
畔的形式影响它的吸收。元素碑很难被吸收，但一些三 价和五价的无机肺化物却容易被吸收 (2) ，同样，有机碑通常
也易被吸收 (6)。至于不同形式肿被吸收的精确数字尚没有报
告。通常接触畔以后呻进入白液 ω ，继之主要在肝、肌肉、 肾、脾和皮肤中出现(1)，也有少量呻在脑、心、子宫、甲状 腺和膜以及头发和指甲中出现 (10) 。呻可穿过胎盘 E130 肺的
生物半减期从大约 10 小时到几天(1)。
没有资料表明，碑的
蓄积随年龄而增加。人体能将无机呻转化成畔的甲基化合物 和二甲基化合物(1)，但是还不了解这种生物转化的全部机
-
82 一
理 (5) 。呻化物主要通过尿排泄 0 ， 2 ， 5)。三价碑能与蔬基反应，
从而抑制很多酶的活性 s 这种反应被认为是碑化物产生毒性 作用的原因 (5) 。
1. 4
对健康的影响
还没有确切的证据表明，任何形式的肺对人是必需的， 然而已知某些有机呻化物作为一种生长剌激素对于动物是有 益的 (2) 。
碑化物的毒性取决于其理化性质、进入机体的途径、接 触的剂量和时间、能与碑相互作用的元素在膳食中的含量以 及接触者的年龄和性另IJ (2) 。无机碑的毒性比有机呻大，三价
无机碑比五价无机呻更危险。有人建议，当发现水中碑含量 为 0.05 毫克/升时，应确定这种碑的化学价和化学式。 急性肺中毒累及中枢神经系统，可引起昏迷，当剂量达
到 70.-.....-180 毫克时可导致死亡。1)。急性碑中毒可严重损害 胃肠道、神经系统、呼吸道和皮肤 (2)。慢性碑中毒表现为一
般的肌无力、食欲减退、恶心，引起眼、鼻和喉粘膜炎症以
及皮肤损伤，也观察到神经症状，甚至致命器宫发生恶性肿 瘤 (2) 。当剂量低到 3.-.....-6 毫克/日，但接触足够长的时间，亦
可出现中毒(1 2) 。曾报道了一个死亡病例，这个人饮用含耐量
高到 7.6 毫克/升的井水达 2.5 年之久 (3) 。甚至在智利水中 耐含量为 0.6 毫克/升也曾引起一些婴儿死亡 (14-16) ，尽管关 于这件事还有些疑问 (2) 。虽然还没有证据表明，水中肺含量
为 0.05 毫克/升对健康有害，但是从饮用该浓度水的人发中 可测出被吸收的肺 (2)。已发现，在中国台湾省凹， 18) 某些村
庄中的皮肤癌与饮用含碑浓度为 0.5 毫克/升的井水有关， 但是对于水中畔的实际含量还有些怀疑 (2) 。世界卫生组织专 一 83 -
家组和美国环保局巳对患皮肤癌的危险性进行了评价(1 9)' 。
世界卫生组织专家组使用线性非阔值模式，估算了终生饮用 含肺浓度为 0.2 毫克/升的饮水患皮肤癌的危险性为 5% (该
模式假设对阔的接触在低浓度和高浓度时其代谢相同〉。 当职业性的大量接触呻时，巳观察到各种各样的表现， 诸如色素沉着增多、角化病和肺癌 (9) 。
参考文献
1. Commission of the European Communities. Trace
metalsa
expos臼 re
and health effects. Oxford ,
Pergamon Press , 1979.
2. Guidelines for Canadian drinking water qua/ity , 1978. Quebec , Ministry of Supply and Services , 1979 (Supporting documentation).
3. Quality criteria for water. Washington , DC , US Environmental Protection Agency , 1976.
4. Some metals and metallic compounds. Lyon , International Agency for Research on Cancer , 1980 (IARC Monographs on the evaluation of the carcinogenic risk of cbemicals to humans , vol. 23). 5. National Research Council. Drinking water and
health. National Academy of Sciences , Washington , DC , 1977. ' 6. Underwood , E ‘J. Trace elements in human and animal
nutrition. New York , Academic Press , 1977. 7. WHO Technical Report Series , No. 532 , 1973 (Trace e lements in human nutrition) 基准文件，
, pp.
49-50.
"Greathou.e , D. G. 饮水中呻的最大可接受限值，为世界卫生组织准备的 1980 (未发表)。
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84 一
8. Toxicology
01
mefals , vo l. 1 1. Washington , DC , US
Environmental Protection Agency , 1977 (Environmental Health Effects Research Series). 9. National Research Counci l. Arsenic. Medical and
biological effects
01
environmental pollutants.
Washington , DC , National Academy Press , 1977. 10. Mealey , J. et al. Radioarsenic in plasma , urine , normal tissues , and intracranial neoplasms. Archives
01
neurology and psychiatry , 81 :310 (1 959). 1 1. Vallee , B. L. et a l. Arsenic toxicology and biochemistry. A. M. A. archives
01
industrial
health , 21 :132 (1 960). 12. Lisella , F. S. et al. Health aspects of arsenicals in the environment. Journal 157 (1 972). 13. Wyllie , J. Investigation of source of arsenic in well water. Canad i an journal (1 937).
01
environmental health , 30:
01
public healt h , 28: 128
14. Rosenberg , H. G. Systemic arterial disease with myocardial iufarction. Report on two infants
Circulation , 47: 270 (1973). 15. Rosenberg , H. G. Systemic arterial disease and chronic arsenicism in infants. Archives (1 974) .
01
pathology , 97: 360
16. Zaldivar , R. Arsenic contamination of drinking water and foodstuffs causing endemic chronic poisoning.
Beitriige zur Pathologie , 151 1 384 (1 974). 17. Tseng , W. P. et a 1. Prevalence of skin cancer in an endemic area of chronic arsenicism in Taiwan. J ournal
01
the National Cancer lnstitute , 40:453 (1968). 85 一
18. Yeh. S. Skin cancer in chronic arsenicism. H uman
path%gy. 41469 (1 973).
19. Arsenic. Geneva. World Hea It h Organization. 1981 (Environmental Hea It h Criteria 18).
-
86-
2. 石棉 2.1 ?.1. 1 来嚣 石棉是对于舵纹石或闪右类呈纤维状的硅酸盐矿的一个 统称。具有作为石棉特征的 6 种矿物是温石棉、青石棉、青 闪石、透闪石、阳起石和铁石棉。温石棉是蛇纹石类的纤维 成分 s 其他则属于闪石类。这些各种各样的石棉是由 40........60% 的二氧化硅以及铁、模和其他金属的氧化物组成。 由于含有石棉的矿溶解及工业废水的排放，致使石棉进
概述
入到自然界水中。有些迹象表明，大气污染同样可以增加石 棉在自然界水中的含量(1)。沉积、再悬浮、迁移和化学反应
均影响水中石棉纤维的移动、丰度和命运。至于从石棉纤维 进入水中至其消失的这段时间究竟有多长，尚不了解 (2) 。
在配水系统中使用石棉-水泥管〈含 170 克石棉/公斤，
其中 80% 为温石棉， 20% 为青石棉〉亦可增加石棉在饮用水 中的含量。
2.1.2
水中存在的右棉
在家庭的给水中经常能够发现石棉，人们认为石棉在河 水平日湖水中者代表性的本底水平约为 1 X 10 8 个纤维/升。 2.
尽管报道的数值波动于小于 1 X 10 8 到 1 X 10 7 个纤维/升的范
,,
围，但是这个数值的明显变动取决于是否靠近工业污染。据 报告，在渥太华河未处理的水中石棉含量为 9.5 X 10 6 个纤 维/升 (1)。 一般来说，普通的沙滤能从供水中去除大约 90% 的石 棉 ω 。去除石棉最有效的方法是在过滤后用铁盐和聚合电解
质进行化学絮凝。根据在加拿大对给水中石棉浓度的调查，
发现大约在 5% 的加拿大公共给水中石棉含量超过 1 X 10 7 个 纤维/升，大约 0.6% 的给水中石棉含量超过 1 维/升 E43.
x 10' 个纤维/ X
升。在一些石棉矿区，饮用水中石棉含量高达 2
10 8 个纤
2.2 2.2.1 饮用水
接触途径
由于报道的结水中石棉含量变动范围很大及纤维聚集成
堆所造成的计数误差也大，因此难以精确地估计平均每天从 饮用水中摄入的石棉量。根据加拿大全国调查结果，以及假 设每天水的摄入量为 2 升，对于 95% 的人口来说，从饮水中
摄入的石棉大约低于 0.0001 毫克/日。
2.2.2
食物
鉴于缺少实用可靠的分析方法，故对于石棉污染固体食
物的程度尚没有进行很好的研究。凡含有固体粒子、灰尘和 泥土的食物，几乎肯定会有石棉纤维。食物中的石棉同样可 能来源于水或者加工制备食物过程中使用的含杂质的滑石 〈滑石作为口香糖、米糕外表的粉及用模子加工食物时的抗
粘剂 )E52. 88
石棉同样还可随被作为喷雾杀虫剂的载体不纯的
一
硅酸盐〈诸如滑石、皂石或叶蜡石〉进入食物 (6)。石棉纤维
是一种极好的撞料，曾被广泛地应用于饮食工业以澄清饮料 和其他液体。有时，由于使用石棉滤器致使最终产品中石棉 的含量增加。已发现，在一些英国啤酒中石棉的含量为 0.151X10 8 个纤维/升 (7)，而在加拿大的啤酒中石棉的含量 为 4.3.......6.6 X 10 8 个纤维/升，在汽水中为 1. 7.......12.2x 10 6 个纤维/升 ω 。 但是，目前在食品工业中已减少了对于石棉
滤器的应用。
2.2.3
空气
石棉存在于空气中，这是由于自然作用(诸如岩石风化〉 或工业排放的结果。在美国，石棉在空气中的典型浓度波动 在 1
X 10- 5 毫克/立方米之间 (8 ， 9) 。据估计， 1 x 10- 8 毫克的温石棉可能含有高达 1 X 10 6 个纤维 (8) 。 假设石棉在空气中的浓度低于 3 X 10- 5 毫克/立方米(在
X 10- 5 到
10
美国城区记录的空气中石棉浓度通常低于该值 (10)) ，每天的 呼吸量为 20 立方米，那么每天从空气中摄入的石棉将少于
0.0006 毫克。
2.2.4
不同接触途径的相对意义
显然，接触石棉的程度有相当差异，这取决于是否靠近
工业和天然的石棉污染源。但是，一般来说，从空气中摄入 的石棉量多于从水中的摄入量。当然应当注意，空气中纤维
大小的分布可能与水中的纤维有相当的不同。关于石棉在食 物中的含量几乎还没有定量的材抖.
-
89 一
2.3 2.3.1 盒入
吸收和分布
石棉纤维可随食物、饮料和饮用水被食入，很大一部分
被吸入的石棉由于粘膜纤毛的作用在呼吸道被清除，继之也 被吞咽下去 (1 1)。关于食入后石棉纤维的结局说法不一，一些 作者认为，石棉纤维可穿过胃肠道壁，进入其他组织 (12-15).
但另一些学者主张，还没有证据能确切地说明这点 E11 ， 16， 1730 把温石棉纤维混悬液直接注入麻醉大鼠的胃中，看到石棉纤 维能透过肠壁，转移到血液、脾、网膜、脑和其他组织中 (13) 。
但在另一个得到很好控制的实验中，
给予大鼠摄入含有 5%
温石棉的饲料达 21 个月，结果没有看到石棉纤维造成的损 害或穿过肠壁转移的证据 (11) 。进一步的工作是确定在正常
情况下石棉纤维穿过人胃肠道壁的量及已吸收的纤维量是否 足以引起有害的局部作用或全身作用 E123.
2.3.2
眼入
动物实验、病理学观察及体格检查多方面研究表明，由 于沉降作用和阻截作用，石棉能沉积于呼吸道内 (18) 。沉降
作用主要取决于纤维的直径大小，而阻截作用则取决于纤维
的长度。 吸入小量石棉后，短纤维被巨噬细胞所吞瞌，继之由于
粘膜纤毛作用被清除，长纤维通常被包围，继之断裂，在 18 个月内消失 (6) 。而另一些石棉纤维被包围，并作为石棉体
或含铁体存留下来。石棉从肺中的清除取决于纤维(形状、
大小、分布和化学类型)、能影响肺炮巨噬细胞活性的任何 90 一
因素以及接触浓度(1 6) 。
动物实验表明，已吸入的石棉纤维可从肺或气管转移到 其它组织 (16) 。在接触青石棉的荷兰猪淋巴结内发现石棉纤 维。可能石棉纤维能通过淋巴系统。在吸入直闪石的豚鼠甲 状腺内亦同样看到石棉体。在间皮瘤患者的肺、 i 淋巴结和腹 膜组织中发现石棉，在脾和小肠组织中也看到石棉体 (19) 。
关于组织中石棉的浓度仅有很少的资料，很需要确定发 生疾病前在各种组织中必须蓄积的纤维数量 (20) 。在一项研
究中报道，在一般病人肺中仅发现很少的石棉纤维 F 在接触 石棉的间皮瘤患者肺中含中等量石棉纤维 s 在患严重石棉沉 着症的病人肺组织中含有数千个纤维 (2 1)。
2.4 2.4.1 2.4. 1. 1 食入
对健康的影口向
动物实验研究
给大鼠喂以温石棉，在结肠上皮和固有层的很多部位看 到石棉纤维 (12) 。
给 32 只 Wistar SPF 大鼠每天喂以含有 100 毫克意
大利滑石粉或加拿大温石棉的麦乳精粉，每周 5 天，在 6 个 多月期间共喂以 100 天，而 16 只对照组动物仅给予麦乳精。 结果在喂以温石棉和滑石粉的每组动物中均出现一例胃平滑 肌肉瘤，但对照组中未出现 (22) 。在大鼠实验中，终生每天食
用含有温石棉的惊纸〈剂量约为 25 毫克/公斤体重)，结果 在 8 到 14 个月出现恶性肿瘤 (23) 。
2.4. 1. 2
流行病学研究
尽管已经进行了很多的流行病学研究，但是仅有一篇流
-
91 一
行病学的研究揭示，在饮水中石棉的含量和消化道癌症之间 刚刚具有相关性 (20 。把在加里福尼亚 5 个海湾区县的 721
个人口调查系统的癌症发生率的资料与饮水中温石棉纤维的
浓度进行了相关性检定，饮水中温石棉纤维含量的范围是从 未检出到 36
x 10 0 个纤维/升。根据低接触区癌症发生率与高
接触区发生率进行比较的资料，适当考虑到社会经济因素， 表明由于接触石棉而引起男性胃癌的发生率为 10% 。但尚需
进一步证实。 对魁北克省 22 个城市的调查，未发现超过本底水平的 癌症死亡率与饮水中存在石棉有关 (25) 。被调查地区包括
Thetford 矿和石棉矿，该石棉矿已被开采了近一个世纪。
最近美国的流行病学家调查了明尼苏达州 Duluth
地区的
胃肠道癌死亡率，自 1955 年以来公共供水中石棉浓度高达
100 x 10 0 个纤维/升，但却没有证据表明胃肠道癌死亡率升 高 (26吨的。职业性吸入石棉发生癌症的潜伏期为 20 ......， 40 年， 所以欲从 Duluth 地区的研究中得出任何肯定结论以前， 尚需更全面的材料。在佛罗里达州 Pensacola 地区进行的
一项研究中，没有看到癌症死亡率与水中石棉浓度之间的关 系，这个地区水中石棉浓度高到 38
x 10 0
个纤维/升 (2 4)。同
样，在华盛顿州 Puget Sound 地区调查的初步结果证明，
在饮水中石棉浓度和结肠、胃、肾或整个泌尿道癌之间没有 相关。在这个地区的一项精细的病例对照研究正在进行 中
如前所述，在配水系统中使用石棉-水泥管是石棉泻染
饮用水的一个潜在性的来源。一般来说，可以断定，由于使 用石棉-水泥管而造成饮水中石棉的含量对人体健康尚未构 成危害 (30时32) 。对康涅狄格 35 年中癌症发病率的调查，未表 一 92
-
明使用石棉-水泥管和胃肠道癌症发病率之间有相关 (33) 。 Meigs 等 (34)发表了对石棉-水泥管和饮水与癌症发病率关
系的进一步研究，获得了明显的阴性结果。
2.4.2
眼入
职业性接触空气中的石棉，可引起肺纤维化〈石棉沉着 症〉、胸膜锦化、支气管肺癌、胸膜和腹膜的恶性间皮瘤及 胃肠道癌。虽然现有资料还不能得出结论，但是有迹象表明， 职业性吸入石棉和卵巢及喉的恶性肿瘤之间存在着一定关 系 (35-38) 。石棉沉着症被形容为是一种"末端细支气管和肺
泡的慢性炎症反应，并伴有明显的纤维化，进而引起变形以 致肺也腔闭合" (39) 。其潜伏期为 7 ......， 20 年。通常人们相信， 各种类型的石棉都能引起石棉沉着症 (40 ， 4 1)，从石棉引起癌 症的病因角度来看，纤维类型不那么重要 (41 ， 42) 。
已经多次注意到，职业性接触空气中的石棉与支气管癌 发生率之间具有明显的相关 (43-47) 0 14 ......， 50% 的石棉沉着症 吸烟和石棉在 患者死于枝气管癌 (40-4 1)。有证据明确提示，
引起支气管癌的过程中有协同作用 (40 ， 46 ， 48 ， 49) 。
联邦德国、英国、南非、美国和其他国家均已注意到， 石棉引起胸膜和腹膜恶性间皮瘤发生率增高 (4Q) 。但在加拿
大几乎没有发现间皮瘤 (43)。从开始接触石棉到出现肿瘤之 间的潜伏期为 20 ......， 40 年 (50) 。
看来，长期吸入石棉和胃肠道癌发生率之间可能有 关 (47 ， 5 1)。据报道，魁北克省温石棉矿工人的胃肠道癌占全
部恶性肿瘤的二分之一，而且对于接触石棉浓度高于 400 x
] 0 8 粒子/立方英尺 ( 14000 X 10 8 粒子/立方米〉的工人来说， 所占比例更高 E41 ， 482. 在有头这个问题的评论中 E522 ，仅有一 93
一
篇文章不支持关于职业性接触空气中的石棉引起消化系统癌 症发生率升高的结论。 参考文献
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32. Olson , H. L. Asbestos. in potable water supplies ,
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N ature , 251'423 (1 974). 40. Parkes , W. R. Occupational lung disorders. London , Butterworths , 1974. 41. Harington , 1. S. The biological effects of mineral fihres , especially ashestos , as seen from in vitro and
in vivo experiments. Annales d'anatomie pathologi que , 21: 155 (1 976). 42. Anderson , H. A. et a I. Ashestos disease resulting from household exposure to occupational dusts. Chest , 66:318 (1 974). 43. McDonald , 1. C. Cancer in chrysotile mines and mills. In. Biological ellects 01 asbestos , Lyon , International Agency for Research on Cancer , 1972 (IARC Scientific Puhlications , No. 8). 44. Wehster , I. Malignancy in relation to crocidolite and amosite.In , Biological ellects 01 asbestos , Lyon , International Agency for Research on Cancer , 1972 (IARC Scientific Puhlications , No. 8). 45. Meurman , L. D. et a l. Mortality and morbidity of employees of anthophyllite asbestos mines of Finland. In. Biological ellects 01 asbestos. Lyon , International Agency for Research on Cancer , 1972 (l ARC Scientific Puhlications. No. 8). 一 98 一
46. Newhouse , M. L. Cancer among workers in the
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textile industry. In. Biological effects of asbestos , Lyon , International Agency for Research on Cancer , 1972 (l ARC Scientific Publications , No. 8). 47. Se !i koff , 1. 1. et a l. Cancer risk of insulation workers in the United States. In. Biological effects of
asbestos , Lyon , International Agency for Research on Cancer , 1972 (l ARC Scientific Publications , No. 8). 48. Selikoff , 1. 1. et al. Asbestos exposure , smoking and neoplasia. lournal of the American Medical
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J. Air pollution and asbestos
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50. Asbestos. Lyon , International Agency for Research on Cancer , 1977 (l ARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vol. 14) pp.62-65. 51. Selikoff , 1.
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J. et al. Asbestos exposure and neoplasia. the American Medical Association , 188:
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一 99 一
3. 3.1 8.1.1 来源
锁
概述
锁在地壳中的浓度为 0.5 克/公斤，其最普遍的形式是 重晶石(硫酸钢)，另一种不大普遍的形式为碳酸顿(毒重
石〉。在大多数土壤中含有微量的坝。顿化物用于石油钻井、 生产油漆、加工柴油燃料、造纸、橡胶、漆布及类似产品、 制造陶瓷袖料和搪瓷制品以及用于医学诊断。
3.1.2
水中存在的银
大多数水中都含有一些钮，但是一般来说浓度远低于 0.1 毫克/升。， 2) ，虽然在地热咸水地区的某些地下水中含顿
量可能高到 10 毫克/升 E330 顿的来源通常是天然的矿物质， 尽管硫酸钢〈是钢存在的主要形式〉仅微溶于水，但是若水
中含有某些通常的阳离子时，可明显地提高顿在水中的溶解 度。关于舰在自来水中的化学形态尚不清楚 (4 ， 5) 。
3.2 8.2.1 饮用水
接触途径
关于银在全世界自来水中的含量知道得较少。但是对 -100 一
100 个美国城市的水质调查中发现，顿的浓度范围为 0.002~ 0.38 毫克/升，均值为 0.043 毫克/升 (6); 在分析的 2595 个 水样申，顿含量超过 1 毫克/升的水样低于 0.1%ω。苏联制 订的饮用水中顿的最高容许浓度为 0.1 毫克/升 E73.
8.2.2 盒" 很多食品中都含有微量顿(1)，巴西坚果含有非常丰富前 钮，其含量高达几千微克/克(1)。 已发表的从膳食中摄入量
的资料很少。发表的一些数值表明，其范围是从低于 0.1 毫 克/日到接近 2 毫克/日 (8)。估计在美国从膳食中顿的摄入量 约为 0.06~ 1. 2 毫克/日 (9 ， 10) ，而加拿大为 0.4 ，......1. 2 毫克/ 日 (5) 。
3.2.3
空气
公开发表的这方面资料极少。但是据报告，美国城市空 气中平均顿的浓度为 0.005 毫微克/立方米空气(1)。
3.2.4 3.2.4.1
其他
职业接触
在工作环境中可检出坝，根据顿对健康的影响，美国制
订在工业环境中可溶性顿化物的阔限值为 0.5 毫克/立方 米 E43.
3.2.4.2
吸烟
在烟草中可检出微量坝，但是由于在吸烟时吸人的银极 少，所以通过吸烟的方式接触顿是不重要的。
一 101 一
3.2.5
不同接触途径的相对意义
为说明相对摄入量，计算了两个简单的接触实例(当水 中顿含量高时)。
假设每天水的摄入量为 2 升，其中顿浓度为 0.05 毫克/ 升，那么每天从水中摄入顿的量应是 0.1 毫克。如每天吸入
空气为 20 立方米，顿在大气中的含量为 0.005 毫微克/立方 米，那生通过吸入的方式每天接触顿为 0.1 毫微克/日，可见
这个量与通过食物或水的途径摄入顿的量比较是可忽略不计 的。假设每天从食物中摄入坝为 0.5~ 1. 2 毫克，则每天从 各种途径织的总摄入量是 0.6 .....，1. 3 毫克。关于儿童对顿的 接触还没有可靠的材料。 鉴于织在饮用水中的含量不可能超过 0.05 毫克/升，故
从水中摄入的顿量与正常总摄入量的比率不可能高于大约
15% 。但是，当水中顿含量高〈超过 1 毫克/升〉时，则这 个比率可高于%。 难以精确定量顿的吸收。食物中的顿不易被吸收 E83 ，但
是如果水中顿是以可谁性盐的形式被摄入，则吸收的比率很
高。因此通过饮水的途径吸收元素钢可能具有特别重要的意 义。
3.3
代谢
还没有证据表明，顿是人体的必需营养素。不可溶性坝， 诸如(在医学上使用的)硫酸坝，很难被吸收，毒性甚低。 而可溶性顿却很容易迅速地被吸收 E43，摄入这种形式的顿能
被吸收 50% 或更多些(1)。通常食物中坝仅能被吸收一小部 分 (8)。被吸收的钢大部分存在于骨中正82，但在其他一些组织， 一 102 一
诸如肾、肝和心中，同样也能检出顿 (8)。
顿的代谢途径与钙相似(4)。顿的排泄比钙更迅速，大约 被吸收顿的四分之一在 24 小时内被排出"， 52.
3.4
对健康的影响
经饮水的动物实验表明，饮水中醋酸顿浓度为 5 毫克/升 〈以顿计〉时，甚至长时期饮用，也未出现毒性 (11)。在大鼠
的 4.......13 周的实验中，顿的总剂量达 250 毫克时，未引起有 害作用 (12) 。
摄入大量的可溶性顿盐出现急性毒性，顿(以氯化顿的 形式〉对成人的致死剂量约为 550.......600 毫克 E430 高剂量引 起所有肌肉〈包括心肌和胃肠道肌肉〉呈现强烈的长时间的 兴奋状态 (4)。一个研究小组报道，引起急性中毒的剂量为 200.......500 毫克(1)。在另一项研究中表明，一次剂量为 125 毫 克是急性作用的阔剂量 (13)。
在某些工业中，吸入含有顿的粉尘出现良性的尘肺 ω 。
近来流行病学研究指出，当顿在水中的浓度达 10 毫克/升时， 与心血管病死亡率有统计学上的相关，但是由于被研究地区 的居民在变化，所以应该谨慎地对待这个发现 (3)。据报告在 中国有顿中毒的流行 (14)。有人认为摄"趴"病是与长期摄入 含氧化顿为 250 克/公斤的食盐有关 E142.
参考文献
1. Toxicology 01 metals. vol. 11. Washington , DC , US Environmental Protection Agency , 1977(Environmental Health Effects Research Series). 2. Quality criteria lor wat.r. Washington , DC , US 一 103-
Environmental Protection Agency , 1976. . 3. Brenniman , G. R. et al. Cardiovascular disease death rates in communities with elevated levels of barium in drinkin'g water. Envi ronmental research , 20: 318 (1 979) .
4. Nationa l. Research Counci l. Drinking water and
health. Waehington , DC , National Academy of Sciences , 1977.
5. Guidelines /or Canadian drinking water qua/ity. 1978. Quebec , Ministry of Supply and Services , 1979 (Supporting documentation). 6. Durfor , C. N. & Becker , E. Selected data on public supplies of the 100 largest cities in the United States ,
1962. Journal
0/
the American Water Works
Associati on , 56'237 (1964). 7. Krasovsky. G. N. et a l. (A proposed maximum a Jl owable concentration for barium in water.)
Gigiena i sanitarija , (6):86
(1 980) (i n Russian).
8. Underwood , E. 1. Trace elements in human and animal
nutrition. New York. Academic Press , 1977. 9. Gormica , A. Inorganic elements in food used in hospital menus. Journal
0/
the American Dietetic
Association. 56:397 and barium. Journal (1 972) .
(1 970).
10. Schroeder , H. A. et a l. Trace metaIs in man , strontium
0/
chronic diseases , 25 =4 91
1 1. Schroeder , H. A. & Mitchener , M. Lifetime studies in rats: effects of aluminum , barium , beryllium and
nutrition , 105:420 (1 975). 12. Guidance /or the issuance 0/ variances and tungstell. Journal
0/
-104 一
exemptions. Washington , DC , US Environmental Protection Agency , Office of Drinking Water , 1979. 13. Browning , E. Barium. Ina Toxicity 01 industrial
metals , London , Butterworths , 1969. 14. Polson , C. J. & Tattersall , R. N. Barium. In , Clinical
toxicology , London , Pitman Medical , 1969.
-106-
4. 被 4.1 概述
4.1.1
来源
镀通常存在于长石矿中，作为长石矿的一个组分，还可 以绿柱石矿的形式存在于少数局限的矿床中(1)。坡的开采量
很少。环境中的镀主要来源于矿物燃料的燃'鹿，尽管这种污 染通常是轻微的 (2) 。由于键的重量轻、张力大，所以这种金
属作为特殊的合金成分被用于空间运载工具、 X 射线窗和 些电器元件等 (3)。
4.1.2
水中存在的镀
岩石风化、大气降尘及在工业上和城市排放的废水都可 使镀进入水中。但是在淡水中坡的浓度很低，通常低于 1 微 克/升 ω。
4.2
接触途径
4.2.1
饮用水
对饮水中镀含量的调查工作很少。在一项美国的调查中
表明，饮水中坡的浓度波动在 0.01 ........，1. 2 微克/升的范围， 平均浓度为 0.2 微克/升。2. 一 106 一
4.2.2
食物
关于键在食物中的含量报道较少。美国的资料表明，通 常镀从膳食中的摄入量约为 100 微克/日 (6) 。据一项英国的 研究估计，平均坡的摄入量可能低于 15 微克/日 (7)。从澳大
利亚的新南威尔士收集的各种各样食物中的镀含量为 0.010.12 毫克/公斤 E830
4.2.3
空气
美国的材料表明，在大气中偶然检出坡，而且通常浓度 很低 (4) ，其范围一般为 0.3........3 毫微克/立方米空气 E63.
4.2.4 4.2.4.1
其他
职业接触
接触坡的工业过程可能包括开采镀矿、航空器材、合金
的机械加工、电镀以及原子能工业等。在工作场所接触键是 通过吸入和经皮的途径。据报告，在没有通风的条件下，工 厂中坡的浓度可高达 23 微克/立方米 (3)。最近美国车间空气 中坡的标准为 2 微克/立方米 E93。-
4.2.5
不同接触途径的相对意义
与从食物和水中摄入的镀相比，从空气中吸入的镀对机
体的总负荷是微不足道的。因此下面的计算仅考虑到食物和 水的作用。
4.3
代谢
键的经皮吸收是微不足道的，因为它能与某些上皮成分 一 107-
C a )每日膳食 z 水中皱
每日摄入 10 微克皱〈成人) 水中摄 λ 量/总量 食物 总量
每周皱的摄 λ 量(微克) 水
的浓度 1 微克/升
(%) 17 28
14 28
70 70
84 98
2 徽克/升
Cb) 每日膳食 z 水中触
每目摄入 100 微克锥(成人〉 水中摄 λ 量/总量
每周敏的摄 λ 量(微克) 水 食物 总量
的浓度 1 微克/升
(%) 2
14 28
700 700
714 728
2 徽克/升
4
结合。已有报道，在存留不溶性镀化物的皮肤局部出现镀皮 炎、肉芽肿溃瘤，此外，还有发生结膜炎的报道 (10) 。镀化
物经消化道不易被吸收，因为在生理性 pH 的条件下镀容易 生成不溶性的化合物 E33 ， 通常仅能吸收摄入量的 1% 以
下。， 1 1)。吸入的可 i容性坡化物经肺吸收入血，主要以正磷酸
盐肢体的形式被运输，最终以不溶性氢氧化物的形式蓄积于 骨中。对于不溶性的镀化物来说，被吸入后可无期限地存留 于肺中 (4 ， 6) 。巳吸收的坡经尿排泄或蓄积于肾和骨中 (3)。键 的排泄是相当快的 (11 ， 12) 。
4.4
对健康的影响
现已证明吸入镀对人类有害 (1 1)。急性吸入中毒可严重 地损害健康，引起鼻炎、咽炎、肺炎和肺水肿 (11) 。
坡经胃肠道很难被吸收，所以经口摄入毒性低。在现有 文献中尚未见到关于人经口中毒的报道。 -108 一
吸入、气管内注入和静脉注入坡，可引起实验动物出现
肿瘤。虽然还缺乏人接触键引起癌症的确切证据，但是有些 证据是值得注意的(4)。在一些诱发肿瘤的实验研究中，给动 物饲以镀含量为 500 毫克/公斤的词料，未能得出结论。在 另两项研究中，给大鼠和小鼠终生饮用含镀浓度为 5 毫克/升
的饮水，与对照组相比，未引起肿瘤发生率具有统计学意义 的明显增加 (13 ， 10 。在另一项研究中，给大鼠喂以镀含量为
5 、 50 和 500 毫克/公斤的词料，作者得出结论 z 无证据表明， 摄入镀具有致癌作用 (15)。
美国的流行病学研究未表明，摄入坡和人类癌症之闹具 有明显的相关。但是国际癌症研究机构 (16) 的结论是...金属
坡和某些镀化物对三种实验动物致癌是有充分证据的。…根 据已得到的实验室资料和人的资料，应该认为镀对人是一种
可疑的致癌剂"。这个结论主要是根据吸入坡的资料。
仅仅苏联颁布了水中坡的标准，即水中坡的最高容许浓 度为 0.2 微克/升 (1) 。
参考文献
1. Sashina , L. A. (Experimental data to substanliate the maximum permissible concentration of beryllium in water bodies.J Gigiena i sanitarija. (2)'10 (1 965) (in Russia 口) •
2. Tepper , L. B. Beryllium. Inl Lee , D. H. K. , ed.
Metallic contaminants and human health , New York , Academic Press , 1972. 3. Oak Ridge National Laboratory. Revi ews
0/ the
environmental e//ects
0/ pollutants VI. Beryllium. -109 一
Ciucinnati , Health Effects Research Laboratory , 1978
(E P A-600/1-78-028) .
4. Ambient water quality criteria 1980 (EPA-440/5-80-024).
10 俨 be 俨 yllium.
Washington , DC , US Environmental Protection Agency , 5. Kopp , 1. F. & Kroner , R. C. Trace metals in waters
01
the U"ite d States. Cincinnati , US Department of ,
the Interior , 1967.
6. Toxicology
01
metals , vo!. 11. Washington , DC , US
Environmental Protection Agency , 1977 (E nvironmental Health Effects Research Series). 、7.
Hamilton , E. I. & Minsky , M. 1. Abundance of chemical elements in man's diet and possible relations with environmental factors. Science
Ih 、
01
the
tota! envi ronment , "375 (1 973). 8. Meehan , W. R. & Smythe , L. E. Occurrence of beryllium as a trace element in environmental materials. Environmental science ançl technology , 839 (1 967).
,:
9. TLVs-Threshold limit va!ues lor chemical substances and physical agents in the workroom environment with intended changes lor 1979. Cincinnati , American Conference of Governmental Industrial Hygienists , 1979. 10. Nichimura , M. Clinical and experimental studies on acute bèryllium disease. Nagoya journal
0/
medical
science , 28.17 (1 966). 1 1. National Research Council. Drinking water and
health. Washington , DC , National Academy of Sciences , 1977. 12. Stockinger , H. E. The toxicology 一 110 →
0/
beryllium.
Washington , DC , US Department of Health , Education and Welfare , 1972 (Publication 2173). 13. Schroeder , H. A. & Mitchell , M. Lifetime studies in rats; effects of aluminum , barium , beryllium and tungsten. Journal
0/ nutrition , 105:420
(1 975).
14. Schroeder , H. A. & Mitchell , M. Lifetime effects of mercury , methylmercury and nlne other trace metals on mice. J ournal
0/
nutrition~
1'05: 452 (1 975).
15. Morgareidge , K. et a l. Chronic /eeding studies with
ber ;y llium sulphate in rats. Pittsburgh , Food and Drug Research Laboratories , Inc. , 1975 (Final report to the Aluminum Company of America , 15219).
16. Some metals and metallic compounds. Lyon , International Agency for Research on Cancer , 1980 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vol. 23) , p. 190.
-111 一
5. 铺 5.1 5.1.1 来源
概述
虽然微量的金属铺均匀分布于地壳中，但在世界的特定
地区存在着铺矿。特别是所有的悻矿都含有少量的铺。为数 不多的铺矿为硫铺矿〈铺闪铮矿) ，即疏化铺，但是它通常
与闵铮矿〈硫化铮〉矿床共生。于十九世纪末锚的生产才逐 渐开始。通常铺是作为开采铸的副产品而生产的。在本世纪
榻的应用日益增多，然而仅在近 20 年才引起人们更大的兴 趣。铺已经污染环境，在空气、食物、土壤、植物以及水中
都能发现它。铺主要应用于制造合金和焊料、金属电镀、颜 料、塑抖稳定剂及电池等。
&.1. 2
水中存在的帽
锢在水中的溶解度是受铺的来源和水中酸度的影响。地
面水中含铺量可能大于数毫克/升，这是由于工业排放废水 或从填埋区的渗漏或从加污泥的土壤的渗漏所致(1)。非污染 水含铺量一般低于 1 微克/升 (2-4) 。
锢在公共给水中的含量通常很低，因为一般来说在原水 中仅含有极微量的锚，即使这个含量稍升高些，大量的铺也 能在常规的水处理过程中被去除 E52. 。"
自来水中含铺量较高，与电镀设备、含银焊料及马口铁 管材有关 El2.
5.2 1.2.1 饮用水
接触途径
饮用水中通常含有很低浓度的铺，约为 1 微克/升或更 低些 (1 ， 6帽835 但是据报道，有时可高达 5 微克/升 E 口，在偶
然情况下，甚至可达 10 微克/升 (9) 。在一些区域井水中可能
含有较高浓度的铺(1)。在低 pH 的软水中可能铺的含量较 高，这是由于对含铺的管道系统具有较强腐蚀的缘故。水样
中铺的含量可能与水和管道接触的时间有关，因此在一天中 不同时间从同一水管中取水时铺的浓度是变化的。如果不收 采大量的水样，很难确定铺的平均摄入量。 假设水的摄入量为 2 升/日，那么估计通过水摄入的锦
为低于 1 微克/日到超过 10 微克/日。当然这个估计是假设 水中所有的铺都能被摄入。但是，很可能在制备一些饮料(例
如茶〉的过程中，并非全部的铺都能被摄入，因为其中一些 铺可能附着于容器的表面等。
5.2.2
盒幅
在大多数食物中都含有微量的铺。在污染土壤(由于工 业污染和使用污泥作为肥料〉中生长的庄稼或用污染水灌溉
生长的庄稼都能使其铺含量增加，就像在受污染的牧场放牧
的动物肉中铺的含量可能升高一样。动物的肾和肝对铺具有
富集作用，所以吃了这些食物的人可能比不吃这些食物的人 摄入较多量的铺(1)。贝壳类动物同样能够富集这种金属(1)。
一 113 一
此外，食物中的铺还可能来源于磷肥。 已发表了一篇关于世界范围内食物中铺含量的综述 (1)。
大多数食物含铺量低于 0.1 毫克/公斤(湿重)。一般从膳食 中铺的摄入量为 15.......60 微克/日 (1 ， 10) 。通常在日本的食物中
铺的含量较高 E92 ，在日本铺的自然本底最高，在污染区铺的
污染水平也最高。关于锢在食物中的化学形态尚不了解，而 这个资料对于确切地确定在食入铺时对这种金属吸收的程度
是有价值的。
5.2.3
空气
通常大气中铺的浓度很低(1l) J 长期的平均浓度变化范 围可能从低于 0.001 微克/立方米至 0.5 微克/立方米，这取 决于工业化的程度和是否存在排放铺的工厂 (10) 。已发表了
关于在世界范围内空气中铺含量的综述 ω 。据估计，一般居 民通常吸入的铺含量低于 0.05 微克/日(1 0)。在一些特殊的
污染区，估计空气中铺的最高含量可达 3.5 微克/日 ω ，甚 至在一些城市大气中的铺含量为农村本底水平的 30 倍，然
而这与食物来源比较仍是很少的。
含铺颗粒能沉积和存留在肺中，其程度随颗粒的大小而 异。据估计，在大气存在的颗粒的 25% 可沉积于肺中(1)。
5.2.4
眼烟
在烟草中含有铺，通常一支香烟含锯量为 1-2 微克， 因为铺在高温时挥发，所以吸烟时能吸入一些铺。对不同国
家从各种香烟中吸入的铺量进行了估计，通常吸 20 支香烟 能吸入 2-4 微觅俑，大约 50% 的铺沉积于肺中(1)。
一 114-
5.2.5
职业接触
工厂工人可能吸入榻的浓度，从几个微克到几千个微克/ 立方米空气 (10' ，但是最高接触浓度多发生在数年前。
5.2.6
不同接触途径的相对意义
从 5.2.1.......5.2.5 节的材料看出，对于不同的个人或人 群来说，通过水、食物、空气和职业性等途径接触锚的量变 动很大。世界卫生组织专家委员会∞推荐，榻的摄入量不能 超过 0.5 毫克/人。为了提供一些特殊情况下的摄入量，在 此提供几个简单的例子。
当水中铺浓度低时 (1 微克/升) ，成人从水中摄入的铺 占总摄入量的比例低于 5.......10% (根据从食物中摄入榻的量
为 15.......60 微克/日，从空气中吸入榻的量为 0.05 微克/日〉。 已有一些资料报道了在食入和吸入铺时的吸收情况，因 此应该考虑铺的吸收比例。在下表中列出了对于成人吸收锚 的估计量，但对于儿童却没有可靠材料。 (a) 每日从食物摄入 20 微克铺〈吸收 6%) ， 从空
气中吸入 0.05 微克铺/目，其中肺存留 25% ，吸收 64% ，饮用水中含有 1 微克铺/升或 5 微克铺/升 〈吸收 6%) 水中铺 每周锅的吸收量(微克〉 从水中的吸收量
的浓度 1 微克/升
水 。 .8
空气
食物
总量
总吸收量<%)
0.1 0.1
8.4 8.4
9.3 12.7
9 33
5 微克/升
4.2
如果每天吸烟 20 支，则从水中的锅吸收量与总吸收量的比卒，将分别减少 到 5% 和 21%.
一 115 一
( b) 每日从食物中摄入 50 微克铺(吸收 6%) 自从空 气中吸入 0.05 微克铺/日，其中肺存留 25% ，吸收
64%.
饮用水中含有 1 微克铺/升或 5 微克铺/升 (吸收 6%)
水中锚的
每周铺的吸收量(微克)
从水中的吸收量
浓度 1 微克/升
水 。 .8
空气
食物
总量总吸收量(%)
0.1 0.1
2 1. 0
2 1. 9 25.3
4
5 徽克/升
4.2
21. 0
17
如果每天吸烟 20 支，则从水中的锚吸收量与总吸收量的此事，将分别减少 到 3% 和 13% 。
5.3
代谢
铺很易从消化道和肺吸收。铺经消化道吸收受很多因素 的影响，诸如年龄、钙、铁、铮、蛋白质缺乏口， 7 ， 10) 及被摄 入锚的化学式等。正如铅那样 E123 ，胃的状态可影响铺的吸
收 P 空胃时吸收的隔量比胃充盈时吸收的要多。标记铺在人体 的实验表明，口服量的 4. 7~7% (均值为 6%) 被吸收(1， 10) 。
食物因素，例如铁、钙和蛋白质缺乏，都可增加胃肠道的吸 收率 (13) 。 已发现，铁缺乏的妇女能吸收铺摄入量的 20% 。
经肺的吸收取决于含铺粒子的大小和溶解度，而且也受呼吸
的深度和频率的影响。对于 0.1 微米的粒子，大约 50% 存留 在肺中，而 2 微米的粒子，则被肺存留 20% (1 0) 。根据在普
通空气中和香烟的烟中含铺粒子大小分布的资料估计，一般 来说分别有 25% 和 50% 的粒子被存留 E1030
被吸收的铺能进入血液，并蓄积于人体的一些部 位口， 1430 肝和肾是榻的储存器官(约 50% 的铺蓄积于这两
种器宫中) (1) 0 铺在很大程度上与低分子量蛋白结合，称为
金属硫蛋白 O l) J 并认为这种金属结合蛋白参与铺的转运和 吸收 (15) 。铺在体内的生物半减期很长(1 3 .-.， 38 年) ，因此其
蓄积随年龄而增加 EIK 人们发现，胎盘对于铺是一个有效的 屏障，所以新生儿体内实际上没有铺〈大约仅为 1 微克 0 ，巧，
但相比之下，一个 50 岁的非职业接触者体内却可能储存 10 到 50 毫克蠕 (5) 。接触大量铺的工厂工人体内可储存 1000 毫 克以上的俑(1)。不吸烟的普通居民血铺含量通常低于 20 微 克/升。， 10) 。血铺反应近期的接触，而不是身体内的负荷量。 铺的排泄一般来说很慢，主要通过尿排泄(1)。就群体而言，
尿铺通常被认为是反应机体负荷量的一项良好指标。铺可与 其它金属，特别是铮，相互作用，所以铺能影响铮在体内的 相对分布。
5.4
对健康的影响
当食物被电镀容器中的铺污染时会发生急性铺中毒，表 现为严重的胃肠道不适口， 5) 。但是还未确定铺对人的急性经 口致死量，估计为几百毫克(1 6) 。已经证实，铺对大量接触
铺氧化物烟尘的工人产生健康危害(1 0) 。巳有报道出现枝气 管炎、肺气肿、贫血和肾结石 (5) 。肾皮质通常被认为是人蓄 积榻的关键器官 (8 ， 10 ， 17) 。铺中毒对肾损伤的主要表现为蛋
白尿、葡萄糖尿和氨基酸尿 (7) 。当接触高剂量铺时，正像在
日本爆发的痛痛病(骨痛病〉那样，结果在那些最严重接触 的人〈特别是老年妇女〉中出现不可逆性肾损害(1)。当摄入
饮水中所含的低浓度铺时，未见任何有害作用的报道，然而 由于使用含隔电镀材料的自动出售机而造成饮料受到锚的污 染，引起了对儿童的急性作用(1)。
已有很多夫于铺的动物实验研究。实验研究证明，经口 一 117 →
长期给予低剂量铺引起高血压，而注入高剂量后出现致畸、 致突变及致癌作用(1)。动物短期饮用含铺浓度为 10 毫克/升 的水，对于胃肠革吸收铁有部分抑制作用(1) 0
Krasovsky'
的工作表明了铺的一般毒性和对性腺的毒作用 E1830
有些证据似乎表明，摄入铺和人的高血压有关，但目前 尚不能结论 (19) 。
铺可能对人有致癌作用的证据是不充分的 (7 ， 20) ，虽然长
期重度的职业性接触可增加患前列腺癌的危险性。 1 ， 22) 。根 据世界卫生组织研究组的意见 E153，关于铺致癌性的流行病
学研究尚不能定论，因为大部分的研究仅调查了很小数量的 工人。基于某些理由，还不能把铺可能有致癌作用应用于确
定职业接触限量。铺的接触水平和血铺浓度之间的关系也不 足以精确地确定血铺的生物学限量。但是世界卫生组织研究 组(1 5)建议全血中铺的浓度为 10 微克/升作为暂时性的无害 作用水平。
据估计， 3 毫克铺是人的一次经口无作用水平 ω 。在 1972 年 4 月召开了第十六次联合国粮农组织和世界卫生组
织关于食品添加剂和食品污染的专家联席会议，建议成人对 榻的暂时耐受摄入量不应超过 400~500 微克/周，即 57--71 微克/日 (9) 。
对于长期接触铺的无作用水平作了估计 (1) 。根据流行病
学研究，建议阔作用剂量为 200 微克/日，这个值相当于每 天吸收 12 微克〈假设经口吸收为 6%) 。如果肾皮质的临界 浓度在 200"""250 毫克/公斤，则计算得出一个人摄入 248 微 克/日〈接近于 200 微克/日的阔值〉的铺，在 50 岁时就可 达到临界肾浓度(1)。正如上述报告 (9) 提及，每日摄入量为
57---71 微克，那么肾皮质浓度就可能达到 1I面界浓度 200 毫 一 118 一
克/公斤的四分之一。
参考文献
1. Commission of the European Communities. Criteria
Cdose effects relationships) for cadmium. Oxford. Pergamon Press , 1978. 2. Hiatt , V. & Ju 日，
J. E. The environmental impact of
cadmium: an overyiew. 1 nternati onal journal of
environmental studies , 7:277 (1 975). 3. Fleischer , M. et a l. Environmental impact of cadmium; a review by the panel on hazardous trace substances. Envi ronmental health perspectives , 7: 253 (1 974).
4. Frib~rg. L. et al. Cadmium in the environmen/. 2nd ed. Clev_land , CRC Press , !nc. , 1974. 5. National Research Counc iI. Drinking water and
health. Washington , DC , National Academy of Sciences , 1977.
6. Guidelines for Canadian drinking-water quality , 1978. Quebec , Supply and Services , 1979 Csupporting documentation) . 7. Ambient water quality criteria lor cadmium.
Washington , DC , US Environmental Protection Ageney , Office of Water Regulations and Standards , Criteria and Standards Division , 1980. 8. WHO Technical Report Series , No. 505 , 1972 CEvaluat 归 n
01
certain food additives and the
contaminants: mercury , lead and cadmium). 9. The hazards to health of persistent substances in
water: annexes to a report of a WHO Working Group. Copenhagen , WHO Regional Office for Europe , 1972 , 10 Commission of the European Communities. Trace
metalsl exposure and health effects. Oxford , Pergamon Press , 1979.
11. Fnvironmental health criteria for cadmium. Genevu , World Health Organization , 1979 (interim report). 12. Chamberlain , A. C. et a l. Report R. 9198. Harwell , UK Atomic Energy Research Establishment , 1978. \3. Flanagan , P. R. et a l. Inereased dietary cadmium absorption in mice and h uman subjects with iron deficiency. Gastroenterology , 74 :841 (1 978). 14. Underwood , E. 1. Trace elemenls in human and
animal nutrition. New Y or k , Academic Press , 19 7'1 . 15. WHO Technical Report S 汀 ies ，
No. 647 , 1980
(Recommended health-based limits in occupational exposure to heavy metalsl report of a WHO Study Group) . 16. Gleason , M. Clinical toxicology of commercial
products. Baltimore , Williams & Williams , 1969. 17. Toxicology of metals , vo l. 11. Washington , DC , US Environmental Protection Agency , 1977(Environmental Health Effecte Research Series). 18. Krasovsky , G. N. et a l. Toxic and gonadotropic effects of cadmium and boron relative to standards for these substances in drinking water. Environmental
health perspectives , 13 :69 (1 976). 19. National Institute for Occupational Safety and Health. Criteria fur a recommended standard. Occupational exposure 10 cadmium. Washington , -120-
DC , US Department of Health , Education ana Welfare , 1977.
20. Cadmium. nicke l. some epoxides. misce 1/ aneous
industrial chemicals and general considerations on volatile anaeslhetics. Lyon , International Agency for Research on Cancer , 1976 (I ARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo I. 11). 2 1. Kipling , M. D. & Waterhouse , 1. A. H. Cadmium and prostatic carcinoma. Lancet , 1 :730 (1 967). 22. Leman , R. A. et a I. Cancer mortality among cadmium production workers. Annals 01 the New Yorh
Academy 01 Sciences , 271 :273 (1 976).
-12] -
6. 锚 6.1 6. 1. 1 来源
概述
大多数岩石和土壤中都含有少量的铭。最常见的矿石为 铅铁矿，在该矿石中的锚是以三价的形式存在的，仅在少数 国家中这种矿具有开采价值。六价铅同样也存在于自然界，
但量很少。自然状态的锚是以极不溶解的形式存在的，但是 经风化、氧化和细菌的作用可转变为较易洛的形式。在土壤 中大多数较易溶解的锚，特别是六价锚，是由于工业废物污
染的结果。某些污染是因在田地上使用了污泥所致。空气、 水和食物受到污染是人类使用铭所致，微量的自然界的铭存
在于食物中。在生物介质中存在三价锚和六价锚，但仅三价 铅是稳定的，因为六价锚易被各种有机物还原 (1) 。
锚主要应用于铝合金、锚电镀、氧化剂、腐蚀抑制剂、 制造恪化物(诸如颜料〉以及纺织、陶瓷、玻璃和摄影工业。
6. 1. 2
水中存在的错
一般来说，由于锚的溶解度小，所以在水中的浓度通常 很低 (9.7 微克/升川口，但是也有水被污染的例子，有些很
严重，含有锚化物的水被排到河流里 P 锚可能以三价或六价
的形式存在，可能是作为可溶性的盐类或是作为不榕性颗粒 一 122 一
形式存在，而且常常以化学复合物的形式存在。在天然水中 恪的化学价是受水的酸度影响的 (3) 。在 pH 中性时，三价 锚转变成不溶性的氢氧化物 (4) 。在原水中总铅含量通常为
10 微克/升或更低，除严重污染的情况以外，很少超过 25 微克/升 (1吨3 。在天然情况下存在高水平锚时可能与水的 硬度高有关 (5) 。
进入公共给水中的出厂水中铅含量，通常与水源水中恪 含量相似或稍低。
6.2 6.2.1 饮水
接触途径
E 常情况下饮用水中含有很低浓度的恪(通常为 5 微克/ 升或更低) (2) 。在自来水中发现锚浓度高达 20 微克/升可能
是很少见的。自来水中铭的浓度可能和水与含铭设备接触时 间呈函数关系，因此由同一水管在不同时间取出的水中锚浓 度可能不同。
根据水的摄入量为 2 升/日，可以估计每天从饮水中摄
入的恪从实际上低于 10 微克/日到偶然的条件下大约为
40 微克/日。很明显，在氯化的水中很少有三价锚，所以设 想大部分水中的锚是以六价的形式存在的(1)。
6.2.2
盒物
食物中的含铭量变化很大，其范围为 20.......590 微克/公 斤。由于引用的数值变化很大和相互矛盾，所以不可能对不
同的食物进行恰当的比较 z
已经发表了关于 45 种食物中锚
含量的一篇综述 (5) 。某些海产品含有高达 0.02.......0.21 毫克/ •
123-
公斤的锚 (8) 。据报告葡萄酒中含有锚，其浓度高达 60 微克/
升 (3) 。食物中的锋是以三价和六价的形式存在的。使用电镀
或不锈钢的容器制各食品时可能造成某些食物的污染。目前 尚缺乏从膳食中锚的总摄入量的资料，据估计在美国从膳食 中摄取锚的量为 5........500 微克/日口， 4 ，的，这个范围大概包括
了全世界的绝大多数食物。从膳食中铅摄入量的平均值大约 为 100........300 微克/日。实际上对于儿童从食物中的摄入量还
未见报道。
8.2.3
空气
关于空气中铭浓度的资料很有限。根据已有的报告，在 城市空气中铭的平均浓度约为 0.02 微克/立方米 E53 ，但在重 工业区的浓度却高于该值的 20 多倍 E53@
大多数空气中的锚是以微细颗粒的形式存在的。当这些 粒子被吸入时，其中半数能存留于呼吸道中。根据每天呼吸
量为 22.8 立方米， 50% 能存留在肺炮中，故每天能储留在 肺中的量大约为 0.2 微克。
6.2.4
吸烟
香烟含有微量恪，每支香烟含1. 4 微克恪 E33 ，其中一些 锚被吸入，并被人体所吸收。虽然确定吸烟者准确的接触量
是非常困难的，但是由于仅少量的铭被吸收，而且大约一半 存留在肺中。所以可以估计，若每天吸 20 支香烟，则每天 存留在肺内的错最多为数微克。
6.2.5
职业接触
在很多工厂环境中，特别是电镀厂和焊接的地方，空气 一 124 一
中恪浓度可能比周围环境要高得多。据记载，空气中锚浓度 高达几百微克/立方米。
6.2.6
吞入尘土
关于灰尘中锚的含量了解甚少，但是一般来说，即使在 儿童育机会吃灰尘的地方，这个途径也不可能是很重要的接 触来源。
6.2.7
不同接触途径的相对意义
从 6.2.1 到 6.2.6 节的讨论中看出，对于不同个人和群 体来说，接触水、食物、空气和环境中锚的机会都有很宽的
范围。为阐明从水中摄入的锚占总摄入量的比例，在此仅列 出几个实例。
6.2.7.1
对于鳝吸收的估计
这个估计是根据从食物和水中摄入时锚被吸收 10% ，从 空气中吸入锋时能被吸收和存留 50% ，并进一步假设一个人 每天摄入 2 升水，每天呼吸空气 20.0 立方米，而空气中锚浓
度为 0.02 微克/立方米。 还没有可靠的资料可用来评价儿童和其他敏感人群对锚 的吸收。
( a )从膳食中络的摄入量 I 100 微克/目 水中筋的 浓度 水 每周锚的吸收量(微克} 从水中吸收量
空气
食物
总量
总吸收量(.%)
20 徽克/升 50徽克/升 100 徽克/升
28 70 140
2 2 2
70 70 70
100 142 212
28 49 66
一 125 一
(b) 从膳食中锋的摄入量: 300 微克/目 水中筋的 每周锚的吸收量(微克) 从水中吸收量
浓度 20 微克/升 50 徽克/升
水
空气 qaM 句 404
食物
总量
J总吸收量(%) qapSAυ 'An4
28 70
210 210 210
240 282 352
100 微克/升
，"嘻
140
6.3
代谢
恪通过胃肠道和呼吸道吸收。但是每个系统的吸收量不 同，这取决于恪的化学式。 2. 对于人类来说，三价锚是这种 元素的基本化学式，而六价恪是有害的。
关于铭从消化道吸收的量，在文献上的报告是不一致的， 精确的数值还不知道。三价锚难以吸收，对三价铅盐的吸收 率为 0.1.-... 1. 2% ，而 25% 的葡萄糖耐量因子〈它是对于正
常葡萄糖耐量所必需的一种恪的复合物〉能被吸收。食物中 天然的恪复合物似乎比简单的错盐更易吸收(1)。看来食物中
的铅至少能被吸收 10%(10) 。给大鼠饮用六价锚含量为 25 毫
克/升的水一年，大鼠组织中锋含量大约比饮用相同剂量三 价锚的动物高 9 倍(11)。因此，推测对于水中六价锚的吸收
率至少是三价锚的 9 倍。水中含有高浓度的恪通常是六价的 恪。关于通过呼吸道的吸收率尚没有定量的材料。但是估计 它取决于颗位的大小和溶解度口， 10) 。有理由估计，吸入锚的
吸收率为被吸入量的 50% 。 分布在入组织内的铝合量低，而且不稳定。除肺脏以外，
在组织中的锚含量随年龄而降低(1)。人体内的铅最主要储存
于皮肤、肌肉和脂肪，组织中的错含量水平与性别、年龄和 -126 …
地区的地理条件呈函数关系。飞机体内的平衡机制〈包括肝 或肠的转运机制〉可以防止三价锚的过量蓄积 (12) 。
锚排泄很慢，主要经尿排泄，也通过粪排泄。
6.4
对健康的影响
铅可能对于葡萄糖和类脂质代谢以及对于几个系统中氨 基酸的利用是必需的。同样对于预防人类的轻度糖尿病和动 脉粥样硬化也是重要的(1)。水中的锚对人体的有害作用与六
价恪有关s 而三价锚却被认为对人是必需的，它实际上是无 毒的，迄今尚未见三价锚具有局部或全身作用的报道。甚至
在世界上动脉粥样硬化患病率低或根本没有这种病的地区的 居民身体组织中锚含量，可能高于在动脉粥样硬化流行区居 民组织中的锚含量 (13) 。
六价铭剂量为 10 毫克/公斤体重能引起人肝坏死、肾炎 和死亡，较低剂量可引起胃肠粘膜剌激 E1430
在大鼠和家兔的实验中看到，当饮用水中六价格含量高 于 5 毫克/升时出现毒作用 E33， 而在其他的研究中当剂量高
到 25 毫克/升时尚未出现有害作用。研究六价铅和胆固醇对 家兔发生动脉粥样硬化影响的实验结果，与铭能抑制实验性 动脉粥样硬化发生的假说是一致的 (15) 。喂以低铭饲料的大 鼠，血清中胆固醇含量也较高(1)。
高剂量的六价锚被认为是对人消化道的致癌因子 EL1830
已有确切的证据表明，接触高剂量锚工人患肺癌的危险性升 高邸， 5) 。对铭酸盐〈六价〉染料工厂的两项调查研究表明，
该厂肺癌的危险性与发生肺癌的危险性很高的制造厂相 似 E72 ，从事加工重铭酸盐〈六价〉或三氧化锚(六价〉的工
人发生肺癌的危险性最大。巳有关于在其他方面使用锚的工 一 127-
厂工人(电镀路工人〉患有前列腺癌和下简窦癌的报告，但
是根据目前的资料还不能估价除肺以外其他部位癌症的危险 性。接触不同洛解度的恪化物〈六价〉混合物(如生产恪酸
盐的工厂)对人类的危险性最大。根据流行病学资料，对于 金属锚、三价铅和六价锚致癌危险性的比较或对于可洛性格
化物和不可溶性锚化物致癌危险性的比较都不能做出评 价 (7) 。还有报告表明，工业接触对健康的其他方面亦有影响，
六价铅能引起皮肤和鼻粘膜溃窃及皮炎(通过皮肤接触) (4) 。
但是接触六价恪对健康影响的阔值是很不清楚的。国际肿瘤
研究机构的专题报告指出 z "还没有证据说明目前非职业性 接触锚的水平对健康具有不利的影响 "EWE参考文献
1. Towill , L. E. et a l. Reviews of the environmental
effects of pollutants , 111. Chromium. Cincinnati , US Department of Commerce , National Technical Information Service. 1978 (PB-282-796). 2. National Research Counci l. C hromi um. W ashington , DC , National Academy of Sciences , 1974.
3. Guidelines for Canadian drinking water quality. 1978. Quebec , Ministry of Supply and Services , 1979 (supporting documentation). 4. National Research Counci l. Drinking water and
health. Washington , DC , National Academy of Sciences , 1977. 5. Commission of the European Communities. Trace
metals. exposure and health effects. O.xf ord , Pergamon Prelø , 1979. 一 128 一
6. Kopp , 1. F. & Kroner , R. C Trace metals in waters
0/
the United States. Cincinnat i. US Department of
the Interior , 1967.
7. Some metals and metallic compounds. Lyon , International Agency for Research on Cancer , 1980 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo l. 23) , p. 303. 8. Teherani , D. K. et a l. Determination of heavy metals and selenium in fish from Upper Austrian waters. 11. Lead , cadmium , scandium , chromium , cobalt , iron. zinc and selenium. Berichte der Oesterreichischen
Studiengesellscha/t /Ur Atomenergie (1 97 7) (SGAE No. 2797 , pp. 1-21 (Chemical abstracts , 88 , No. 49150e)). 9. Underwood , E. 1. Trace elements in human and
animal nutrition. New York. Academic Press , 1977. 10. Friberg , L. et al. Chromium. In: Handbook on the
toxicology
0/ metals ,
Amsterdam. Elsevier/North-
Holland Biomedical Press. 1979. 1 1. Mackenzie , R. D. et a l. Chronic toxicity studies. 11. Hexavalent and trivalent chromium administered in drinking water to rats. A. M. A. archiues
0/
industrial health. 181232 (1 958). 12. Schroeder , H. A. et a l. Abnormal trace metals in man-chromium. !ournal (1 962) •
0/
chronic diseases , 15'941
13. Schroeder , H. A. The Tole of chromium in mammalian nutrition. American journal 21 1230 (1 968).
0/
clinical nutrition ,
14. Kaufman. D. B. et al. Acute potassium dichromate -129 一
poisoning in man. American journal
0/
diseases
0/
children , 119 374 (1970). 1
15. Novakova , S. et a I. (The content of hexavalent chromium in water sources and the effect on the development of experimental arteriosclerosis in warm blooded animals.) Gigiena i sanitara"j a.39(5)1 78-80 (1 974) (in Russian). 16. Teleky. L. Krebs bei Chromarbeiten. Deutsche
medizinische Wochenschri/t. 62 1 1353 (1936). 17. Some inorganic and organometallic compounds. Lyon. International Agency for Research on Cancer. 1973 (l ARC Monographs on the evaluation of the carcinogenic risk ot chemicals to humans. vol. 2). p. 100.
一 130 一
7. 氟化物 7.1 概述
7. 1. 1
来源
只要有生命和工业的地方就存在氨化物。鲁在化物可分无 机事在化物和有机氟化物两种，后者通常被划归为腊类。氧化 物，特别是作为代谢的中间体，参与生命的过程。霞化物最 普通的形式包括氟化氢〈在溶液中为氢氨酸人易恪于水的 霞化物盐类和金属-氟化物复合物 E130 氧化物离子能与重金 属离子结合，生成复合物，其中某些复合物是很稳定的 (2)。
氨化物盐类水解，生成氢篝酸(1)。 氨化物应用于很多的工业过程中，例如生产丙烯腊、己
二腊和异丁烯酸甲醋。氟化物同样还应用于提炼金银，生产 钢，电镀及制备在化学合成中的一些中间体 (2) 。在这些过程
中可能造成空气和水的污染。偶尔使用氨化物杀灭害虫也可 能是水受污染的原因。
7. 1. 2
水中存在的氧化物
氢氟酸在水中离解出氧离子 (1 ， 2) ，它的离解作用取决于 pH ，当 pH 在 8.2 以上时以离子形式为主 (2) 。 当 pH 为
8.5 或以上时氧化物将转变成毒性低得多的氯酸盐 (3) 。除非
严重污染时，主要是工业废水排入河流或其它污染来源引起
的污染情况下，一般来说，在原水中氟化物浓度较低(低于 0.1 毫克/升〉 ω 。处理金属工业，炼焦，生产煤气和各种各 样的化学制品都可能是氧化物污染水的主要来源 (4) 。在中性
或碱性条件下，由于饮水氯化而产生的游离性余氯，可使出 厂水中氨化物浓度减少到很低的水平 (3) 。水的氯化作用 (pH
大于 8.5 时〉使氧化物转变成无毒的氨酸盐 (3 ， 5) ，最终可能
分解为二氧化碳和氮气。
7.2 7.2.1 饮用水
接触途径
还没有关于饮水中氧化物浓度的全面资料 p 但一般来说，
这个浓度远低于最大可接受水平。
7.2.2
食物
大多数食物含有微量的氨化物。一些植物来源的食物
(例如杏仁〉含有的氨化物比天然本底高。在生活污染水中 的鱼里发现霞化物。 3 。靠化物遇热分解，因此烹调的食物含 氧化物的量很低阳。通常对每天从食物中摄入的氟化物量不
甚了解，但一般来说日摄入量是低的。已经确定，允许每天 从熏蒸食品中摄入的氟化物量为 0.05 毫克/公斤体重 (7) 。
7.2.3 的。
空气
具体的数值尚没有发表，但是通常认为，这个值是很低
-132 一
7.2.4
其他
已育报告表明，在某些工业环境里接触霞化物的剂量可 能较高 (8) 。在这种情况下，工业接触可能是主要的接触途径。
7.2.5
不同接触途径的相对'义
由于没有充分的资料来确定从食物〈除饮水以外，食物 是氟化物的主要天然来源〉中的摄入量，所以不可能可靠地 估计从食物和水中的摄入量所占的比例。
7.3
代谢
霞离子易被动物体吸收，故迅速地表现出高毒性。氨化 物抑制颈动脉体和主动脉体细胞的氧化过程，致使厌氧产物
(例如乳酸〉蓄积，剌激呼吸。这个反应是由于氟化物与细 胞色素氧化酶中的含铁辅基结合，进而阻断了向分子氧上提 供电子。因此细胞对氧的吸收受到抑制，即细胞不能进行氧 化作用，所以停止了对于细胞能量的基本供应。当葡萄糖不
绝氧化时，神经元使葡萄糖以更快的速度转变成乳酸。在动 物实验中看到，即使给予小剂量的氟化物，脑乳酸含量增加，
还可引起动物昏迷、惊厥伴有不可逆性的脑损伤，尽管对机 体的其他部分并没者损害 (9) 。接触低剂量霞化物对人类不会
玫死，因为人类具有一个有效的解毒系统，能使戴化物转变 成硫氧离子，这种离子在低浓度时是无毒的 (2) 。
7.4
对健康的影响
一次剂量 50~60 毫克通常引起人的死亡 (3) 。一般认为，
每天接触 2.9 ........， 4.7 毫克的氨化物对人是没有损害的，因为
人体内有高效的解毒系统，通过硫氨酸酶和硫代硫酸盐酶系 统，能使人体内的氨离子转变成相对无毒的硫氯化物(1)。接
触较高剂量的氨化物可引起死亡。根据计算，对于人每天允 许摄入量为 8.4 毫克氟化物阳。
.增文献
1. Quality criteria /or water. Washington , DC , U~ Environmental Protection Agency , 1976 (E P A-440/9-76023).
2. Guidelines /or Canadian drinking water quality , 1978. Quebec , Ministry of Supply and Services , 1979 (supporting documentation).
3. National interim primary drinking water regulations. Washington , DC , US Environmental Protection Agency , 1976. 4. Gotts , R. M. et a l. Treatment
0/ industrial wastes
c
,'
municipal water po l1 ution control plants. Ontario , Ontario Water Resources Commission , 1966 (Proceedinge , Ontario Industrial Waste Conference) , p. 15 1. 5. Cyanides.1na Kirk , R. E. & Othmer , D. F. , ed.
Encyclopedia
0/ chemical technology , 2nd
ed. New
York , John Wiley and Sons , 1965. vol. 6 , p. 574. 6. Leduc , G. et al. The use of sodium cyanide as a fish eradicant in some Quebec lakes. N aturaliste canadien , 100: 1 (1 973).
7. Internationa! standards /or 队 US 一 134-
drinking ω ater.
3rd ed.
Geneva. World Health Organization , 197 1. Env ironmental Protection Agency. Water quality
criteria. availability. Federal register , 44:43667 (1 979) .
9. Passmore , R. & Robson , J. S. A companion to medical
studies. vo l. 2. Oxford , Blackwell Scientific Publications , 1970 , pp. 15-25.
-135-
8. 氟化物 8.1 8. 1. 1 来摞 氟化物是一种极普遍的元素，在地壳中的含量约为 0.3 克/公斤(1)。它是以氟化物的形式存在于一些矿物中，最普遍
概述
的是氟石、冰晶石和氟磷灰石 z 很多石头也含有氟化物。氟
化物在工业上用于生产铝，并且通常存在于磷酸盐肥料、 砖、瓦、陶瓷中，也用于冶金工业 (2) 。此外，还常常将其加 入药物制品中，包括牙膏和维生素的补充品 (2) 。由于工业中
如此广泛地使用含氟物质，所以在环境中普遍地存在着氟化
物的污染，在植物、食物和水中都含有微量的氟化物。
8. 1. 2
水中存在的氟化物
在很多水中存在微量氟化物，氟化物浓度较高通常发生
在地下水中。在富含氟化物矿〈如氟磷灰石〉的地区，井水 中氟化物含量可能高达 10 毫克/升或以上口， 4) 。据报告，最
高的天然含量达 2800 毫克/升(1)。而大部分水的含量低于 1
毫克/升 ω 。此外，有时可因工业排放致使氟化物进入河流。
-136 ......
8.2 8.2.1 饮用水
接触途径
除进行加氟处理以外，在自来水中氟化物的浓度与水源 水类似。一般而言，在未加氟的结水中氟化物的含量通常低 于 1 毫克/升，但是这取决于水源的类型和状况，在个别情 况下可高达 10 毫克/升。 30 迄今在世界的大部分地区已经将
这种水源鉴别出来。在给水中加氟的地区，在正常情况下，
氟化物的浓度为 0.6....... 1. 7 毫克/升，其决定因素是周围大气 的温度 (5) 。因此以每人每日水的摄入量为 2 升计，则从加氟
处理的饮用水中每日摄入氟化物的范围为1. 2.......3. 4 毫克，在 其他地区每日摄入量的范围从低于 1 毫克直至在例外的情况
下也许高达 20 毫克。
8.2.2
空气
在空气中含有氟化物，主要来源于工业排放。由于工业
活动的类型不同，其浓度变动很大。但是，据估计一般来说 通过空气接触(相当于空气中氟浓度低于 1 微克/立方米 (2)) 与食入的氟量相比是无足轻重的例。 8.2.8 盒栩
实际上，所有食物至少都含有微量的氟。所有蔬菜也含 有一些氟，这是从土壤和水中吸收的。某些食物可能含量很 高，特别是鱼、某些蔬菜和茶。， 2) 。例如，一些鱼的含氟量
可高达 100 毫克/公斤，而茶可高达此浓度的两倍以上，大多 数其他食物很少超过 10 毫克/公斤 ω 。使用氟化给水的食品 一 137 一
加工厂所制备的食品中氟化物含量可能往往是原食物中含氟
量的两倍。一些国家估计了每日从食物中摄入的氟化物量， 成人为 0.2-3.1 毫克 E12，在美国 1'""-'3 岁年龄组儿童的摄入
量约为 0.5 毫克/日 "2.
8.2.4 8.2.4.1
其他途径
职业性接触
已知许多工业加工过程可向车间空气中排放含氟化合 物。一些文献报道了人体接触的实例，特别是冶炼铝和生产 玻璃时(1)。在这种条件下(即空气中浓度达数毫克/立方米) , 职业接触在人体的总接触中就会占有主要部分。但是，随着 工作条件的改善，这一接触途径就可能显得不那么重要了。
几乎没有关于从吸烟中接触氟化物的资料，当然，与其 他接触途径相比，通过吸烟接触不是氟化物的重要来源。
8.2.4.2
一般人群的接触
各种产品，如牙膏、牙粉、口腔清洗剂、口香糖、维生 素补充品和药物，可能含有额外加入的可溶性氟化物〈主要 是无机氟形式)，通常加入牙膏中的浓度约为 1 克/公斤(1)。 研究表明，通过这种途径可能吸收大量的氟化物 F 现已证明， 每一次刷牙可能吸收约 50 微克(1)。局部应用氟化物洛液可 能增加吸收量的。口腔清洗剂可提供高达 2 毫克的氟化
物(1)。现已有一系列不同的含氟制剂，包括被用作为抗踊剂 的含氟化铀的片剂。定期使用这种片剂，每天可供给大约 1 毫克的氟化物 E130
8.2.5
不同接触途径的相对意义
下表说明加氟饮水对成人从食物中摄入氟化物总量的影 一 138 一
响。 从食物中摄入的氟化物为 1 毫克/日，饮用水中含
氟化物 0.5 、1. 0 和1. 5 毫克/升 F 水的摄入量为 2 升/日 水中氟化 物的浓度
从水中的摄 λ 量 水
每周氟化物的摄 λ量(毫克} 空气* 食物
总摄 λ量 总量
(%) 50 67 75
0.5 毫克/升 1. 0 毫克/升 1. 5 毫克/升
7.0 14.0 2 1. 0
0.0 0.0 0.0
7.0 7.0 7.0
14.0 2 1. 0 28.0
*从2i':气中的摄 λ 量可忽略不计(未考虑从其他途径的接触)。
本文未提供从食物和水中吸收氟化物的估计量，因为一 般而言，吸收效率很高 E83 ，上述的摄入量与相应的吸收量是 很接近的。
8.3
代谢
从水中摄入的氟化物几乎全被吸收(1)，食物中的氟化物 不象水中氟化物那样能被完全吸收，然而吸收率仍然很高，
但是从某些食物〈如鱼和某些肉类〉仅能吸收约 25% 的氟化 物 (8) 。
被吸收的氟化物迅速分布全身。主要存留于骨悟，小量 存留于牙齿中 (7) 。骨中氟化物的含量不断增加直至 55 岁 (9) 。
在高剂量时，氟化物干扰糖类、脂类、蛋白质、维生素和矿 物质的代谢(1)。急性氟化物中毒的许多症状是因其与钙结合 作用的结果 (6) 。氟化物主要通过尿排泄，其排泄速度受很多
因素的影响，这些因素包括一般健康状况及氟化物的接触
虫。)，而存留速度随年龄的增长而降低，为了实用的目的，
可认为大多数成年人是处于"平衡状态 "ω 。在这种稳走状 态下，存于体内的氟化物存留于钙化组织中，大部分其他的 氟化物存在于血浆中，因此可以排泄。骨髓的存留和肾的排 泄是防止中毒量氟化物蓄积的两种主要途径。
8.4
对健康的影响
已经肯定证实，氟是某些动物种属的必需元素 (7); 特别
是相当小剂量的氟能提高生育力和促进生长速度的。 一旦氟化物掺入牙齿，就能降低语琅质在酸性条件下的
溶解度，从而防止发生鹉齿。有充分的证据表明，水中的氟 化物能明显降低儿童和成人的踊齿。 3 。虽然氟化物的浓度达
1. 5""2.0 毫克/升时者时发生斑袖齿，甚至影响美观 E63，但 是当浓度约为 1 毫克/升时踊齿发生率降低。长期饮用氟化
物浓度为 1 毫克/升的水，慢性肾脏病或烦渴症的病人可能 出现斑轴齿 E62，但是仅仅发生在牙齿正处于钙化过程中的
。 ""7 岁儿童。当水中氟化物放度高于 3""6 毫克/升时(还取 决于其他的摄入来源〉出现氟骨症。长期每天摄入 20""40 毫克氟化物或更高， (即当水中浓度超过 10 毫克/升)时， 出现氟骨症所造成的残废 (6) 。已经公认，就氟化给水而言，
水中氟化物含量为 1 毫克/升是安全的，因此建议其在水中 的限量为 1 毫克/升左右。， 102， 确切的浓度限量取决于气
温。
高剂量氟化物对人具有急性毒性，病理学改变包括出血 性胃肪炎、急性中毒性肾炎及不同程度的肝脏和心肌损 伤 (1)。氟化铀的急性致死量约为 5 克，即约 2 克氟化物 (2) 。
在重污染区动物从环境中接触氟化物，可以看到各种严重的 中毒症状 (7) 。接触高浓度氟化物对人的慢性作用主要是斑抽 一 140-
齿租氟中毒(即骨的结构受到影响) ，有时甚至引起令人吃惊 的严重残废(1)。也观察到对肾脏的慢性作用，但通常见于患 肾脏病的人 ω 。其他比较特殊的问题是接触高浓度氟化物对
甲状腺的影响。 3 。中毒的最初症状是呕吐、腹痛、恶心、腹 泻、甚至抽搞 (6) 。
对天然水中含氟量高的区域所进行的流行病学调查表 明，很少呈现有害作用 (6); 在水中含氟量特别高的区域则出 现斑袖齿和氟骨症。 3 。在氟化物浓度适宜的区域曾友现两名 糖尿病患儿出现明显的氟作用(即令人厌恶的斑袖齿) (6) 。
有的研究似乎表明，伸舌样白痴和癌症可能与水中氟化 物浓度高有关。伸舌样白痴可能与接触氟化物有关的想法来 自于一项很局限的研究，据记录登记的伸舌样白痴的患病率
与水中氟化物浓度有关Cl1)，但是该研究受到了英国皇家内 科学院的严厉批判 (6) 。近 30 年来进行了各种流行病学调查
以评价癌症和水中氟化物浓度之间是否有关，结果仅在少数 调查中两者呈正相关，但是这些报道已经受到了抨击 (2) 。目
前一般认为，没有可接受的证据表明水中氟化物是人类的致 癌物口， 6 ， 12 ， 13) 。
已经提出了是否存在对氟化物的敏感性问题 ω 。总之，
已经打消了某些人对氟化物敏感的说法。例如，数百万的饮 茶者未呈现特殊的敏感性，且然他们从茶叶漫出物中摄入了
大量氟化物。当然不能完全排除敏感性或某些特异性反应的 可能性 E63，尽管从现有证据中预计其发生率是很低的。目前 已有一些实验表明，氟化物具者致突变和致畸作用，可能还
与先天性缺陷有关。对于这些方面己进行了全面综述，但是 证据尚不够充分 E62.
嘈i
参考文献
1. Fluorides and human health. Geneva. World Health Organization , 1970 (Monograph Series , No. 59).
2. Guidelines /or Canadian drinking water quality. 1978. Quebec , Ministry of Supply and Services , 1979 (supporting documentation).
3. Fluorides. Washington , DC , National Academy of Sciellces , 197 1. 4. Bulusu , K. R. et a l. Fluorides in water , de fJ uoridation methods and their limitations. J ournal
0/
the
Institution
0/
Enginee 俨 s
<I ndia) , 60 d γ inking
(1 979).
5. International standards /or
water. Geneva ,
World Health Organization , 197 1. 6. National Research Council. Drinking water and
health. Washington , DC , National Academy of Sciences , 1977. 7. Underwood , E. J. Trace elements in human and
animal nutrition. New York , Academic Press , 1977. 8. Newburn , E. & Zipkin , I. Fluoride metabolism.
Fluoride and dental caries. Springfield , IL , Charles Thomas , 1976. 9. Jackson , D. & Weidmann , S. M. Fluorine in human bone related to age and the water supply of different regions. Journal 451 (1 958).
0/
palhology and bacteriology , 76:
10. European standards /or drinking water. Geneva , World Health Organizatioll , 1970. 11. Rapaport , 1. Nouvellea recherches sur le mongolisme , -142 一
A propos du rôle pathogénique du fluor. Bu l! etin de
l'Académie nationale de M édecine. 143:367 compounds , and inorganic Iluorides use r/ in
(1 959).
12. Some aromatic amines , allthraquillones and nitrJso drinking-water and dental preparations. Lyon , International Agency for Research on Cancer , 1982 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo l. 27). 13. Clemmesen , J. The alleged association between artificial fluoridation of water supplies and cancer. a review. Bulletin
01
the World Health
Organization , 61: 871-883 (1983).
9. 硬度· 9.1 概述
水的硬度不是一种特异性的成分，而是一种不稳定的复 杂的阳离子和阴离子的混合物。硬度主要取决于钙和镖，虽
然硬度也受银、顿和其他多价离子的影响。硬度通常是以每 升中相应量碳酸钙的毫克数来表示的。这是整个这篇文章所
采用的单位。一些国家也使用几种其他的单位。传统上认为， 硬度是水与肥皂反应能力的测量。它通常被分为"碳酸钙"
〈暂时〉型的硬度和"非碳酸钙"(永久〉型的硬度。
9.1.1
来源
钙和侯是很多矿物中普遍的元素，水中钙和模最普通的
来源是石灰石，其中包括白圭(碳酸钙〉。钙和钱存在于许多 的工业制品中，并且它们是一般的食物成分。
对水中总硬度起较小作用的包括一些多价离子，如铮、 缸、铝、悟、、顿和铁，它们是从矿物(如闪铮矿、 armangl
te 、铝土矿、菱惚矿、碳酸银矿和磷酸菱铁矿)中溶解出来 的。
9.1.2
水中存在的硬度
虽然大多数钙化物不易溶解于纯水中，但是在存在二氧 *关于硬度对水质其他方面的-些影响见第五章第 5 节 (380 页}。
-144 一
化碳的条件下易使其溶解度升高，水中含钙高达 100 毫克/ 升的水源是极普遍的 CH43; 但水中含钙超过 200 毫克/升是
少见的(1 -4) 。含模的很多盐是易溶解的，水中含模高达 10
毫克/升的水源是普遍的。-4)。水中含镇量高至Jj 100 毫克/升 的水源是少见的(1阳42，而钙的硬度通常是主要的。
水的缓冲能力通常被描述为碱度，它与水的哽度密切相 关。因此像氢氧化物、碳酸氢盐和碳酸盐等阴离子对水的缓 冲能力影响很大，而磷酸盐和硅酸盐影响较小。弱酸的分子 型同样也能产生影响。
9.2 9.2.1 饮用水
接触途径
虽然在水处理厂有时对水进行人工软化处理，但是一般
来说原水的硬度和送往住户水管中饮水的硬度是相似的。硬 度变化的范围可能从小于 10 毫克/升到 500 毫克/升以上 ω， 水源水的硬度可能是小于 50 毫克/升，水硬度高于 500 毫克/ 升在大多数国家是相当罕见的口， 3 ， 5) 。
9.2.2
食物
实际上所有食物都含有钙和镜。典型膳食每天提供大约 1000 毫克钙阳和 200~400 毫克镇(1， 7) 。摄入钙和模的主要
来源通常是食物。特别是乳制品富含钙 E23 ，而钱倾向于与肉 类和植物来源的食物有关 (6) 。
9.2.3
空气、职业性接触和吸烟
虽然这些对人来说都是接触的途径，但是与食物相比都 一 145-
是微不足道的。
9.2.4
不同接触途径的相对意义
只有食物和水才是重要的接触途径，因此只考虑这些来 源。对于可能的情况估计如下(所有的估计均是针对成人 的h
(a) JJ...食物中旬的摄入量: 1000 毫克/目 水所( 中占% r h ，、，， 锅，『
从量
λ例
rr 民 LVUM
*中钙的浓度
每周钙的摄 λ量(毫克) 水 食物 总量
25 毫克/升 100 毫克/升
P 。
350 1400 2800
7000 7000 7000
7350 8400 9800
'in4 "'03
200 毫克/升
从水中钙的典型摄 λ 量约为总摄 λ 量的 5"-20% 。
(b 水中锐的浓度
)从食物中键的摄入量:
200 毫克/目 从水中摄 λ 量所占比例
每周偿的摄 λ 量(毫克) 水
食物
总量
(%) 9 33 50
10 毫克/升 50 毫克/升 100 毫克/升
140 700 1400
1400 1400 1400
1540 2100 2800
(c 水中锐的浓度
)从食物中镜的摄入量:
400 毫克/目 从水中摄 λ 量所占比例
每周鳞的摄 λ 量(毫克) 水
食物
总量
(%) 5 20 33
10 毫克/升 50 毫克/升 100 毫克/升
140 700 1400
2800 2800 2800
2940 3500 4200
从水中簇的典型摄 λ 量约为总摄 λ 量的 5"-20% •
… 146
•
虽然已有一些关于儿童从膳食中摄入量的资料，但是尚
未进行计算，因为这个可以根据不同接触途径摄入量的相对 比例对于儿童和成人大致相似这一点估计出来的。 对于钙和键的吸收量还未做单独的计算。虽然从食物中
吸收钙和钱分别为 30% 和 35% ，但是几乎还没有从自来水 中吸收钙的可靠资料，因此仅能对吸收量做粗略的估计。
9.3
对健康的影响
有些暗示性的证据表明，饮用硬度极高的水也许会导致 尿结石患病率增高。这可能解释在苏联局部地区一个小量的 人群中出现的尿结石，在那里自来水中含钙量为 300-500 毫克/升 (8) ，在动物实验中也得到了证实， 30 "C高温中，饮用硬度很高的水 让动物生活在 (200-400 毫克钙/升) (9) 。
但是在饮水中含钙量高到 500 毫克/升必定是罕见的。 这样看来，似乎还没有确凿的证据说明水的硬度能引起
对健康的危害 EZU 相反却有一些研究结果暗示，水的硬度能 预防疾病。
9.3.1
水的硬度和心血管攘病
1957 年在日本证明，中风死亡率与来源于河水的饮水中 酸度密切相关(4)。此后在世界各地进行的很多研究证朋，水 的硬度与心血管疾病之间具有统计学上明显的负相关 (3 ， 10) 。
在大多数的研究中表明钙浓度具有最明显的相关，然而在一 些加拿大的研究中指明镜的含量有很大的相关系数 (4 月。但 是在一些小规模的研究中还不能确定这种相关 (11-13) 。关于
"水作用"对心血管病影响的大小及混杂因素影响的程度尚
不能肯定。混杂因素是指诸如气温、降雨、纬度、经度、社 -147 一
会经济因素、城市类型和空气污染等。然而，当考虑到很多 混杂因素的变量，就能证明具有统计学上明显的相关 (2 ， 1 4) 0
于 1975 年曾召开了该课题的一个特殊的国际科学讨论会 ω 。 最近在英国 253 个市镇进行了一项大规模的回顾性调查研 究 E143，在考虑到气候条件和一些社会因素后表明，当水硬度 高达约 170 毫克/升(以 CaCO a 表示〉时中凤和贫血性心
脏病的死亡率与水的硬度明显相关。在分析很多水的因素中，
与水的硬度和水钙含量的相关性最明显，虽然与水硬度有关 的很多其他水的指标从统计学上与心血管病也有明显的相 关。这项回顾性调查是很有限的，正在继续研究 F 特别寄予
希望的是在选择的人群组对心血管病的危险因素进行回顾性 研究以评价产泛的水指标的意义 (14) 。 为说明这种相关提出了几种假说，但是目前还没有明确
的证据表明包括水的硬度或其主要成分钙和镜。两个最经常 引用的假说是: (a) 硬水中的成分能预防疾病 J (b) 软水中
的物质(例如从管材中逸出的金属)能促进疾病。预防性假 说中常常认为，食物供给足够量的钙和侯，虽然在有些情况 下食物中有可能缺乏镜 (7) 。但是存在着的其他因素，例如 鲤、锚、矶、砂也可能有预防作用 (15 ， 1 的。有迹象表明，从管
材可能逸出铅和铺，可能促进疾病发生，但是还没有确凿的 证据表明这两种金属参与这一作用 (3) 。
9.3.2
水的硬度和其他疾病
几个研究结果暗示，多种多样的其他病可能与水的硬度 有关，这些疾病包括一些神经系统缺损、无脑畸形、死胎和 各种癌症 ELM-2030 虽然其中某些发现已在不同国家得到了
证明白马但是对于它的意义仍有相当的怀疑。这种相关可能 一 148 一
仅仅反映了疾病的类型，而这些类型能用社会、气候和各种 环境因素，而不能用水的硬度来解释。 参考文献
1. National Research Counci I. Drinking water and
health. Washington , DC , National Academy of Sciences , 1977. 2. Guidelines lor Canadian drinking water quality. Quebec , l\1 inistry of Supply and Services , 1979 (supporting documentation). 3. Amuris , R. et al" ed. Ha r_ dness 01 drinking ω ater
and public health. Oxford , Pergamon Press , 1975 (Scientific colloquium , Luxembourg , 1975 儿
4, Marier , 1. R. et a I. Water hardness , human health ,
and the importance 01 magnesium. Ottawa , Canada , National Research Council. 1979 , ,
5. Quality criteria lor water. Washington , DC , US
Environmental Protection Agency , 1976. 6. WHO Technical Report Series , No. 532 , 1973 (Trace
elements in human nutritiona report of a WHO Expert Committee) , 7. Neri_, L , C. & Johansen , H. L. Water hardness and cardiovascular morta 1i ty￥ Annals 01 the New York
Academy 01 Sciences , 304:203 (1 978). 8. Bokina. A , 1. et al , (Hygienic assessment of drinking wate.r hardnees as a factor favouring the development of uro 1i thiasis.) Gigiena i sanitari ja , 30 (6): 3 (1 965) (in Russian) , 9 , Bokina , A , I. & Yurieva. V, K. (Shift8 of certain 一 149 一
- biochemical indices in persons after long-term use of hard drinking water.J Gigiena i sanitarija , 31 (í2):33 (1 9tiò) (i n Russian).
10. Kobayashi , J. On the geographical relationship between the chemical nature of river water and death-rate from apoplexy. (1 957) . Be 俨 ichte
d es Ohara
Instituts /ar Landwirtscha/iliche Biologie , 2:12 1 1. Allwright , S. P. A. et a I. Mortality and water harrlness in three matched communities in Los Angeles. Lancet ,
2:860 (1 974). 12. Bierenbaum , M. L. et al. Possible toxic water factor in coronary heart disease. Lancet ,
,:
1008 (1 975). th 龟 water
13. Meyers , D. Ischaemic heart disease and factor. A variable relationship. British
jo 盯 nal
0/
preuentiue and social medicine , 29'98 (1975). 14. Pocock , S. J. et a l. British regional heart studYa geographic variations in cardiovascular mortality , and the role of water quality. British medical journa! ,
28011243 (1980). 15. Voors , A. W. Lithium in the drinking water and atherosclerotic heart deatha epidemiological argument for a protective effect. American journal
0/
epid emiology , 92 a 164 (1 970). 16. Schwartz , K. Silicon , fibre , and atherosclerosis.
Lancet , "454 (1 977). 17. Stocks , P. lncidence of congenital malformations in the regions of England and Wales. British jou 俨 nal
0/
preuentiue medicine , 24 167 (1970). 18. Hart , J. T. The distribution of mortality from 一 150-
coronary heart disease In South Wales. Journal (1 970) .
0/
the Royal College 0/ General Practitioners , 19:258 19. Fedrick , J. Ancncephalul and the local water 8upply.
Nature , 227 1 177
(l~宁ω.
20. Lowe , C. R. et al. Malformations of the central nervous system and softness of local water supplies.
British medicol :journal. 2 1 357 (1971).
-151 一
10. 铅 10.1 10. 1. 1 来翻 铅是地壳的一种天然组成成分，其平均浓度大约为 16 毫克/公斤 (1) 。铅存在于一些矿物中，其中主要的矿是方铅 矿〈硫化铅) J 在大多数国家都储存着某种铅矿。铅已被广泛 地应用很多世纪了，许多地方已经出现了环境污染，这是由
概述
于采矿、熔炼或使用铅制品所致。所以铅存在于空气、食 物、水、土壤、灰尘和雪中。环境中的铅几乎完全以无机铅
的形式存在，而少量的有机铅是由于使用含铅汽油和自然界 的烧基化过程的缘故，所谓烧基化过程就是生成甲基铅化物
的过程 ω 。铅具有广泛的用途，包括制造酸性蓄电池、用于
汽油的炕基铅化合物、焊揭、染料、弹药、培缝和电缆辅。 目前已经停止和劝阻应用铅作为屋顶材料和管道材料〈包括 用于饮用水的管道材料〉。
10. 1. 2 水中存在的铅 据估计， 在全世界的湖水和河水中铅的天然含量为 1"" 10 微克/升口， 3) 。虽然在受污染的地区，特别是有工业污染源
的地区水中铅含量较高，但是这种情况还是较少见的，因为 有一些控制铅含量的天然机制.在水进入管网前的出广〈即 一 152 一
处理后〉水中通常铅浓度低于水源水，因为大多数的常规水 处理过程能把部分铅去除(4)。但是饮水中铅的水平也可能较 高，这是由于水从街道到家庭流经了铅管和〈或〉使用了内 衬铅的储存罐 (4 ， 5) 。当水具有腐蚀性、软水或 pH 低时，铅
浓度可能特别高。这些情况甚至可使铅浓度达最高水平 (1 ， 4 ， 6) 。目前在全世界已经广泛地不使用铅管了， 水中含高浓度的铅口， 4) 。
但是在某些国
家的一些镇和城市中仍然广泛地使用铅，因此偶尔造成自来
10.2 10.2.1 饮用水
接触途径
在大多数国家，家庭自来水中铅浓度很低，通常远低于 10........20 微克/升。但在某些地方，水中铅浓度可能很高 σ 例 如，在苏格兰，水极软， pH 低，普遍使用铅管和内衬铅的 储水罐。因此，据估计， 在苏格兰 10% 以上的家庭大约为 500 万居民中第一次流出的水含铅量高于 300 微克/升 E730
据记载，在世界上一些地方饮水中铅含量超过 2000 微克/ 升 E的。但是很难根据水中的铅浓度精确地确定平均接触水 平，因为在龙头水中铅浓度变动很大。该浓度主要取决于水
在铅管中或在家庭给水设备中停留的时间。甚至在同一的供
水区，家与家之间水中的铅浓度也有相当的不同，这是因为 水管长度、用水类型及水的储存类型不同所致。 根据每天水的摄入量为 2 升，计算表明，每天从水中摄 入的铅为 10""""20 微克到 1 毫克或以上。这个估计是基于假 设全部铅都能被摄入。然而制备一些饮料(例如茶〉时 E93 ，
不是水中所有的铅都出现在制备的饮料中。另一方面，自来
水同样用于烹调和制备食物，这就为从家庭自来水中摄入铅 增加了另外的机会。
10.2.2
盒"
铅存在于各种各样的食物中，其含量的变化取决于食物 的种类。例如，如果在生产罐头时使用铅焊惕，则罐头食品 中可能含有很高的铅 E3， 10， 1130 很多新鲜蔬菜、谷类和水果含
铅量低，这是由于从生长的土壤中吸收的铅有限以及空气中
的铅沉积于地表面的缘故。铅还存在于牛奶、奶制品和葡萄 酒中(1 0) 。
由于各种食物中铅含量变动很大，所以不可能精确地计 算铅的摄入量。估计通常每天摄入量的范围从少于 100 微克 到多于 500 微克铅 (1 ， 5 ， 8 ， 10 ， 12) ，对于成人全世界的平均值大
约为 200 微克/日。估计近年来铅在食物中的含量逐渐降 低吼 10) 。由于女人吃饭少于男人，所以女人从食物中摄入的 铅量较低。估计 1-5 岁儿童摄入的铅约为 90 微克/日 (10) 。
食物中另外的铅可能来自于烹调容器的污染，诸如被焊接的
锅和一些袖制的陶器炊具。某些铅则来自于制备食品时使用
的自来水。一般来说，摄入的铅主要来源于食物。
10.2.8
宝气
在农村空气中铅的平均浓度为 0.1 微克/立方米白，叽 UL 通常在城市中的平均浓度为 0.5-2 微克/立方米 (8 ， 10) 。任何
一个特定区域的铅浓度取决于污染源〈例如交通、工业〉的 类型和程度以及这个区域的自然扩散状况〈主要的气候条 件〉。目前，在非工业城市空气中的大部分铅来源于机动车，
一般来说居住在高速公路附近的居民接触最多。据记载，在 一 154 一
一些工业区空气中的平均铅浓度高达 6 微克/立方米邸， 8) 。
空气中铅的沉降可污染土壤，据报道土壤中铅的含量为 2 克/ 公斤 (8) ，严重污染的土壤中铅含量为 10 克/公斤以上 (8) 。
大部分空气中的铅是以细颗粒的形式存在的。当这些颗 粒被吸入时只有 20""""60% 沉积在呼吸系统中 (8) 。根据每天
呼吸空气量为 15 ，....... 22.8 立方米比 83，一个城市居民(接触 1
微克铅/立方米空气，存留 40%) 通常每天铅的吸入量为 6"""" 9 微克，而存留的大部分铅最终会被吸收。
10.2.4
职业性接触
工厂工作区空气中铅含量可能比一般环境中高得多。但 是其含量高达 100 微克/立方米空气是罕见的 C830
10.2.5
吸烟
在烟草中发现少量的铅，但是一般来说经过这个途径吸 入的铅是很少的。
10.2.6
土壤、灰尘以及咀嚼油潦而食入的铅
土壤、灰尘、特别是家中的油漆(尤其是旧油漆〉都含
有高浓度的铅。由于小孩"抓食"(异食癖) ，因此他们可能从 这个途径食入大量铅 0 ， 3-8) 。虽然在许多的研究中分析比较
了土壤、灰尘和油漆对于儿童接触铅的相对意义，但是尚不 能肯定从上述来源的铅所占的精确比例 (6) 。
10.2.7
不同接触途径的相对意义
对于不同的个体和居民来说，从水、食物和空气等途径 摄入的铅可能很不相同。因为从每一种途径摄入的铅所占的 一 ]55 →
比例可能变化很大，所以不可能对于广泛的环境提出详细的
资件。但是近来已经考虑了某些环境的状况，并且做了数量 上的估价(1 0) 。为了对于-些可能的情况给予某些概念，在
此列举几个简单例子。在下表的最后两列提出了如下估计值 2 即从水中铅的摄入量与机体总摄入量的相对比例以及从水中 的吸收量与总吸收量的相对比例。然而下表中不包括一个重 要的婴儿组(直到 1 岁)。关于食物中铅含量以及吸收的现有
资料不足以对此做出估计，但是该组婴儿通过饮水每天摄入 的铅相当于或可能高于表中列出的 1~5 岁儿童组。此外，对 于吸烟、职业接触或各种其他途径摄入铅量的相对比例未予
计算。
10.2.7.1
成人每周摄入和吸入铅量的估计值
下表的例子假设 z 每人每天平均摄入 2 升水，呼吸空气 20 立方米/日(1 0) 。从各种途径对铅的吸收量不受从其他途径
摄入量的影响，从食物和水中的吸收量是摄入量的 10% 以
及从吸入途径而存留在呼吸道的 40% 的铅全部被吸收。
(a
)每天从食物中铅的摄入量为 100 微克 F
空气中铅的浓度为 4 微克/立方米 每周错的摄 λ 量(毫克) 水中铅浓度 水
摄 λ 量比吸收量.1"1:. 水/总量 水 j 总量 总量
空气
食物
(%)
(%)
20 微克/升 50 徽克/升
0.28 0.70 1. 40
0.14
。 .10
1, 12 1. 54
25 47 63
18 35 52
0.14 0.14
0.70 0.70
100 微克/升
2.24
( h)
每天从食物中铅的摄入量为 300 微克; 空气中铅的浓度为1. 0 微克/立方米 每周知的摄 λ 量(毫克)
水中知浓度 水
空气
食物
总量
摄 λ 量比吸收量地 水/总量 水/俨
(%)
(%)
20 徽克/升 50徽克/升 100 徽克/升
0.28 0.70 1. 40
0.14 0.14 0.14
2.10 2.10 2.10
2.52 2.94 3.64
11 24 40
10 20 34
10.2.7.2
1""'5 岁儿童每周攫入和吸收铅量的估计值
下表假设 z 每天水的摄入量为 1 升，呼吸量为 4.7 立方 米/日 (10) 。除从食物和水中对铅的吸收量为摄入量的 50% 以 夕俨4 ， 82，关于铅吸收的基本假设与成人相同。
每天从食物中铅的摄入量为 93 微克 p 空
气中铅的浓度为1. 0 微克/立方米 每周铅的摄 λ 量(毫克} 水中铅浓度 水 摄 λ 量比吸收量比 水/总量 水 IJ总量
空气
食物 。 .65 。 .65
总量
(%)
(%)
20 徽克/升 50 徽克/升 100 徽克/升
0.14 0.35 0.70
0.03 0.03 0.03
0.82 1. 03 1. 38
18 35 51
11 35 51
0.65
10.3
代
谢
一项很重要的考虑是从水中摄入的铅真正被吸收的比例
是多少。虽然关于存在于自来水中铅的微细粒子的吸收情况 知道得很少，但是已有一些水溶液中可溶性铅经肠道吸收的 -157 一
资料。通常认为成人约吸收 10%(4 ， 5 ，的 a ，这取决于饮水时
究竟是在胃充盈时还是空腹时。例如口服铅离子前后备禁食 6 小时的人对铅的吸收明显增加(例如 50% 或以上) 03 ， 14) 。 这个发现在小鼠实验中获得证实 (15) 。其他一些因素，诸如
食物中存在的一些元素〈钙、磷、铁、铜和辞等〉以及人的 年龄和身体状态都影响从胃肠道对铅的吸收(1， 6 ， 8 ， 11) 。
肺的吸收取决于铅粒子的大小及呼吸的深度和频 率。， 8) 。一些大的粒子沉积于呼吸道粘膜，其中一些最后被吞 咽下去 (8 ， 10) ，肺通常可存留 40% 白， 8 ， 10) 。
被吸收的铅进入血液，分布于软组织和骨路中。当长期
接触后，在血液和软组织中铅浓度达到平衡。相比之下，骨 悟能随着时间的延长而蓄积铅。尸体检查表明，铅在骨路中 的负荷随年龄而增加 z 为 2~4 周邸， 8 ， 14) 、
事实上大约机体负荷的 90% 是在骨
髓中 (8 ， 1 1)。据估计，铅在血、软组织和骨路中的半减期分别 4 周 (1 4)和 27.5 年 (10) 。
铅易通过胎盘，胎儿血中铅浓度几乎与母血中相同。铅 同样能通过血脑屏障，虱然脑不能蓄积铅(1 6) 。
对铅的摄入和血铅浓度之间的相互关系巳进行了广泛研 究和评论 '6) 。空气中铅和血铅之间存在着一种曲线的关系 z 当错的接触增加时，则相应的血铅增高的程度小 (17) 。从水
中摄入铅和血铅之间同样呈曲线关系，这在饮用含高铅浓度 自来水〈超过 50微克/升〉的人体中已得到了证实(1 8 ， 19) 。
应该强调，目前对接触低浓度铅与血铅之间的关系还有
明显的不能肯定的因素，因此，当使用从该关系中获得的资 料时要特别小心。 铅从尿、粪、汗、头发、手指甲和脚趾甲中排泄。在世 a 儿童对错的吸收率较高，通常认为 5 岁以下儿童吸收率达 50 ，%川， 81. -158 一
界卫生组织的两篇文章中已对铅的代谢进行了综述帆 16) 。
10.4
对健康的影响
几个世纪以来，已经认识到高剂量的铅是一种蓄积性的
毒物。急性中毒的一些症状包括疲劳、乏力、腹部轻度不 适、易激动、贫血 z 儿童还有行为改变 (8) 。这些症状难以定
量。目前对于各种可能的轻微影响，包括接触低剂量铅引起 的神经生理方面的变化，都给予很大的关注白， 1130 低剂量铅 能使胆色素原合成酶活性降低 (8) 。该酶在。一氨基』γ-嗣戊酸
转变为胆色素原的阶段参与正常血红素的合成，所以这种酶 活性降低可作为接触铅的指标。铅对含硫氨基酸也有亲和力。
此外，铅易于与线粒体结合，从而干扰对氧的转运和能量产 生的惆节作用 (10) 。在精神发育迟滞的儿童中巳发现血铅含 量明显增高(大于 400 微克/升〉 E2030
已有很多动物实验研究了铅对血液生成、神经、肾、心 血管和生殖系统的影响 (8 ， 10 ， 11) 。虽然这些结果不能直接外
推到人，但是确实提供了有价值的剂量-反应资料，这是在人 体流行病学研究中不具备的。
已经进行了广泛的人体流行病学和临床的研究，包括回 顾性研究，以确定接触铅人群的死亡和发病的原因，以及研 究铅对特殊器官和系统的作用 '8) 。特别重要的是研究水中铅
浓度与成人和儿童血铅浓度的关系以及儿童发生的轻度行为 影响 (6)。很多研究表明，水中铅浓度很高的地区血铅增加很
少，但一般而言，与个体或人群组可接受的血铅浓度相比，由 于饮水而产生的血铅浓度不高 E6 ， 830 例如，据估计饮水平均含
铅浓度为 100 微克/升时，成人血铅平均增加 25 微克/升。对于 幼儿和娃振妇女血铅约分别增加 40 和 50 微克/升队 10) 。但 一 159-
是对这些数值的解释必须很小心，因为铅的摄入量与血铅之
间具有曲线相关。 人群组从整个环境中接触铅的平均限值为 200 微克/升 血液，而对于个体来说，该值范围为 300-350 微克/升。美 国疾病控制中心和美国儿科科学院建议个体儿童对铅的接触 限匾为 300 微克/升血液队 10) 。对于一般人群铅浓度的典型
水平了解得还不太精确，但是对于非职业接触的成人通常引 用的数值低于 200 微克/升。估计在美国 (6) ， 99.5% 的儿童
血铅低于 300 微克/升，而几何均数维持在 150 微克/升。
很多关于铅对儿童行为影响的研究表明，这种影响与环 境中高浓度的铅有关，而与水中铅浓度升高无关。在一些情 况下有迹象表明，铅的有害作用与水中铅浓度之间有些关 系 (8 ， 10) 。
在几项研究中发现，接触铅的人群(血铅浓度为 100-
1000 微克/升〉外周血淋巴细胞染色体发生畸变。而在其他 研究中〈血铅范围为 40-500 微克/升)却得到阴性结果 (2 1)。
已知铅不是维持生物系统功能的必需元素，所以一般认 为应尽可能减少对铅的接触。虽然 1972 年联合国粮农组织/ 世界卫生组织专家委员会建议成人对铅的最大摄入量为 3 毫 克/周 (0.05 毫克/公斤体重) (12) ，但是对于儿童还没有建议
相应的数值。对于儿童〈包括婴儿〉的情况有所不同，因为 儿童对铅的吸收比成人高(1 2) ，儿童还具有较高的敏感性，其 部分原因是由于儿童生长速度快 E430 孕妇和发育中的胎儿对
铅更敏感，这是由于母亲对食物的摄入量增加以及激素状况 的改变 (8) 。如呆从膳食中铅的摄入量大于 220 微克/日，就会
超过成人每周限量 3 毫克的规定。
已有三个世界卫生组织的专家组关于铅对健康的影响进 一 160 一
行了综述 (8 ， 16 ， 22) 。
参考文献
1. Guidelines /or Canadian drinking water qua 1i ty , 1978. Quebec , Ministry of Supply and (supporting documentation). 2. Harrison , R. M. & Laxen , D. P. H. Natural source of tetra-alky 1 lead in air. N ature , 275: 738 (1 978). Service 目，
1979
3. The hazards to health and ecological effects 01 persistent substances in the environment: report of a working group. Copenhagen , WHO Regional Office for Europe , 1973. 4. National Research Counci l. Drinking water and
health. Washington , DC , National Academy of Sciences , 1977. 5. Toxicology
0/
w f!. ter , vo 1. 11. Washington , DC , US
Environmental Protection Agency , 1977 (Environmental Health Effects Research Series). 6. US Environmental Protection Agency. Ambient water
quality criteria /or lead. Washington , DC , Criteria and Standards Division , Office of Water Planning and Standards , 1980 (E P A 440/5-80-057).
7. Lead in
drinking ωater ， a
survey in Great Britain.
London , Department of the Environment , 1977 (Pollution paper No. 12). 8. Lead. Geneva , World Health Organization , 1977
(Environmental Health Criteria 3). 9. Zoeteman , B. C. 1. & Brinkmann , F. 1. 1. In: Amavis , R. et al., ed., Hardness
0/
drinking waler 一 161 一
and public health , Oxford , Pergamon Press. 1975. 10. Drill , S. et al. The environmental lead problem.
An assessment 0/ lead in drinking water /rom a multi-media perspective. Washington , DC , US Environmental Proteetion Ageney , 1979.
11. Underwood , E. J. Trace elements in human and
animal nutrition. New York , Academic Press , 1977. 12. WHO Technieal Report Series , No. 505 , 1972
(Evaluation
0/
certain /ood additives and the
contaminants: mercury , lead and cadmium). 13. Chambe r1 ain , A. C. et a 1. Investigations into lead
/rom motor vehicles. Harwell , Oxfordshire , Atomic Energy Research Estab 1ishment , Environmental and Medical Sciences Division , 1978 (AERE-R9198). 14. Rabinowitz , M. et a 1. Studies of human lead metabo 1i sm by use of stable isotope tracers.
Environmental health perspectives , 7 1 145 (1 974). 15. Garber , B. T. & Wei , E. In f1 uence of dietary factors on the gastrointestinal absorption of lead. Toxicology
and applied pharmacology , 27:685 (1 974). 16. WHO Technical Report Series , No. 647 , 1980
<Recommended health-based limits in occupational exposure to heavy metals). 17. Hammond , P. B. & Beliles , R. P. Met.lsl lead. Inl Doull , J. et a 1., ed. , Casarett and Doul l' s
toxicologYI the basic science New York , Macmillan , 1980.
0/
poisons , 2nd ed.
18. Moore , M. R. et a l. Contribution of lead in water to blood-lead. Lancef , 2:661 (1977). 19. Thomas , H. F. et a l. Relationship of blood lead in -162 一
Women and children to domestic water lead. 1'1 ature , 282: 712 (1 979). 20. Moore , M. R. et a l. A retrospective analpis of blood-lead in mentally retarded children. Lancet , '1717 (1 977). 2 1. Some metals and metallic compound!. Lyon ,
International Agency for Research on Cancer , 1980 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo l. 23). 22. Health hazards from drinking-waten report of a
Working Group. Copenhagen. WHO Regional Office for Europe , 1977 (ICP/PPE 005).
一 163 一
11. 采 11.1 11.1.1 来翻
概述
环境中东的主要来源是从地壳中天然释放的，每年为 25 ， 000'-""150 ， 000 吨。此外，虽然很多工业活动与亲的生产 或使用并无直接关系，但却向环境中排放大量的亲 p 这些工
业活动包括矿物燃料的燃烧，各种金属冶炼，生产水泥以及
废弃物的处理等。来可用于氯碱工厂(生产氯和氢氧化纳入 油漆(作为防腐剂或色素〉、电键和电池、测量和校正仪器 (例如温度计、医用仪器)、牙科和农业(特别是种子肥料)
中。在环境中来能以金属、一价和二价盐以及有机来〈最重
要的是甲基来〉的形式存在。存在于水的底泥和污泥中的微 生物可将无机亲转化为甲基来 p 而其他微生物却可使亲去甲 基化变为无机束。鱼和哺乳动物对甲基苯吸收和贮存的能力 比对无机宋强 p 因此甲基苯能在食物链中蓄积(1)。
1 1.1. 2
水中存在的亲
据报道，在瑞典的雨水中含来量约为 300 毫微克/升 (1)。 在大部分地面水中，主要的亲化合物为氢氧化柔和氯化
来 p 一般浓度低于 '0.001 毫克/升 (2-5) 。在污染的河流和湖 泊中，浓度达 0.03 毫克/升 (6) 。联邦德国内陆水中录浓度约 一 164 一
为 400 毫微克/升，而河流中的浓度范围为 100........1 ， 800 毫微 克/升 (7) 。
通常饮用水中的亲浓度很低 E830 例如在加拿大几个省的
饮用水中，来浓度的中值约为 0.0002 毫克/升 (9) 。当用铁和 明矶进行水的混凝处理时无机来可被去除(1 0) 。从联邦德国
饮用水贮水池收集的 700 份水样化验结果表明，最纯净的饮 用水中来浓度低于 0.00003 毫克/升 (1 1)。
11. 2 11. 2.1 饮用水
接触途径
在无证据表明受到来污染的地区，在淡水中家浓度低于 0.0002 毫克/升。常规水处理过程能明显降低水中的录浓 度，在制备饮抖时可能进一步降低。因此，按偏高的估计 (4) ，
在一般情况下从饮水中摄入的来不超过 0.1 微克/日 El30
11. 2.2
食物
食物是非职业接触人群中柔的主要来源，而鱼和鱼制品 又是食物中大部分甲基来的来源(1)。从食物中来的每日平均 摄入量为 10........12 微克口， 12 ， 13) ，但是在水体受到亲污染，而
且鱼又占膳食中很高比重的地区，从食物中摄入的录可能会 高得多。
11. 2.8
空气
由于金属亲和有机家化合物具有相当高的蒸气压，所以 能通过蒸发作用进入大气中。除污染区外，大气中的浓度为 0.02 微克/立方米。假设大气中的浓度为 0.05 微克/立方 一 165 一
米，则金属来蒸气的平均日摄入量约为 1 微克/日，其中 80% 存留在体内 E130
11.2.4
职业性接触
在很多职业和商业中接触景 (14) 。特别重要的是采矿、氯
碱工业和生产某些科学仪器的工作，空气中来的浓度可高达 5 毫克/立方米(1)。美国政府工业卫生医师会议建议可接受 的时间-加权平均阔限值为 0.05 毫克/立方米，假设工作时肺
的通气量为 10 立方米/日，则职业接触可使每日平均摄入量 为 500 微克或以下(1)。
11. 2.5
不同接触途径的相对意义
下表说明了成人从饮水中摄入来的相对意义， 每周柔的摄 λ 量(吸收量)(微克) 水 比值，水/总量(%) J总量
食物
空气
(a)70(5.6) (b)140 (11. 2) (c)70(5.6)
2.8(2.2) 2.8(2.2) 7.0(5.6) 7.0(5.6)
14(2.1) 14(2.1) 14(2.1) 14(2.1)
87(9.9) 157(15.5) 9 1( 13.3) 161 (1 8.9)
16(2 1. 0) 9(13.6) 15 (1 5.8) 9 (11.1)
(d)140 (11. 2)
假设 I (1)食物摄 λ量为 10 徽克/日
(a ， c) 或 20微克/日仙， d) ，其中摄 λ 量
的 8% 被吸收。
(到空气中的浓度为 0.02 微克/立方米价， b) 或 0.05 徽克/立方米衍， d) ，总通气量为 20.0 立方米/日，其中 80% 被存留。
(3)饮水中浓度为 0.001 毫克/升，每天摄 λ2 升水，其中摄入量的 15% 被吸收。
尚无有关儿童的可靠资料，。
11. 3
代谢
录对人体无有利的生理功能。各种理化状态的来(金
属、无机化合物、有机来化合物〉均具有各自的特性，决定 了独立的毒理学作用。4)。
从食物中摄入的无机来约吸收 7---8% ，而实际上胃肠 道对甲基来的吸收是完全的。从饮水中摄入的无机来可能吸 收 15% 或更少(1 5) ，但是对甲基录几乎能完全吸收。
无机来化合物能迅速蓄积在肾脏中，肾脏是这些化合物 作用的主要靶器官(1 5) 。被吸收的甲基宋能迅速出现在血液
中，在人的血液中， 80---90% 的亲与红细胞结合 F 而甲基亲
去甲基化为无机来的过程也以缓慢而明显的速度进行着。甲 基苯比无机录毒性大是由于其脂溶性，从而能更容易地穿过 生物膜，特别是进入脑、脊髓和外周神经，并且能透过胎盘。 来盐从肾、肝、肠粘膜、汗腺、唾液腺和奶中排泄，最 重要的途径是经尿和粪排出 (15) 。
1 1. 4
对健康的影响
来中毒的主要作用是使神经和肾脏功能紊乱，这些作用 分别与有机和无机录化合物有关 (16) 。根据 Krasovsk y Cl7)
的资料，来除了具有一般毒性作用外，还具有对性腺的毒性 和致突变作用以及干扰胆固醇代谢的作用。世界卫生组织环
境卫生标准丛书 ω 已经详细地综述了来化合物的毒理学。还 不可能鉴定无机、芳香基和皖氧基烧基来的近似最低作用剂
量。迄今无证据表明无机来为致癌物。而烧基示对实验动物 具有胚胎毒性和致畸作用 (9) 。
-167-
参考文献 1. Mercury. Geneva , World Health Organization , 1976 (Environmental Health Criteria 1). 2. Holden , A. V. Present levels of mercury in man and his environment. In. Mercury contamination in man
and his environment , Vienna , International Atomic Energy Agency , 1972 , p. 143 (Technical Report Series No. 137). 3. Wiklander , L. Mercury in ground and river water. σ rund /oerbaettring ,
21 '151 (1 968).
4. Wershaw , R. L. Sources and behaviour of mercury in surface water. In. , Mercury in the environment , Washington , DC , US Geological Survey , 1970 (Professional Paper No. 713). 5. Voege , F. A. Levels of mercury contamination in water and its boundaries. In , Proceedings
0/
the
symposium on mercury in man's environmen t. Ottawa , Royal Society of Canada , 1971 , p. 107.
6. Investigations 0/ mercury in the St. Clair RiverLake Erie systems. Washington , DC , US Department of the Interior , 1970 (Report of the Federal Water Quality Administration) , p. 108. 7. Schramel P. et al. Some determinations of Hg , As , Se , Sb , Sn and Br in water , plants , sediments and fishes in Bavarian rivers. International journal
0/
environmental studies , 5:37
(1 973).
8. National Research Council. Drinking water and
health. Washington , DC , National Academy of Sciences , 1977. 一 168-
9. Guidelines /or Canadian drinking water quality , 1978. Quebec , Ministry of Supply and Services , 1979 (supporting documentation).
10. Guidance /or the issuance 0/ variances and exemptions. Washington , DC , Office of Drinking Water , US Environmental Protection Agency , 1979. 11. Bouquiaux ,1. In , Proceedings
symposium on the problems
0/ an international 0/ contarnination 0/ 阴q
and his environment by mercury and cadmium. Luxembourg , Commission of the European Communities , 1974 , p. 23.
12. Meranger , J. C. & Smith , D. C. The heavy metal content of a typical Canadian diet. Canadian journal :1
0/
public health , 63:53 (1 972).
13. Neilsen-Kudsk , F. Absorption of mercury vapour from the respiratory tract in man. Acta pharmacologica et.
toxicologica , 23:250 (1 965). 14. Key , M. M. et a l., ed. Occupational diseases-a guide
to their recognition. Washington , DC , US Department of Health , Education and Welfare , 1977 (NIOSH publication 77-18 1), pp. 370-373.
15. WHO Technical Report Series , No. 647 , 1980
(Recornmended health-based Ii mits in occupational exposure to heavy metals). 16. Swedish Expert Group. Methyl mercury in fish. A toxicologic-epidemiologic evaluation of risks.
Nordisk hygienisk Tidskri/t , Supp l. 4 (1 971). 17. Krasovsky , G. N. et a l. (The need for revising the existing hygienic standard for mercury in water.) σ igiena
i sanit 宜 rija (2):20 (1 98 1) (in Russian). 一 169 一
12. 镰 12.1 12.1.1 来翻
概述
镇是普遍存在的 z 典型的土壤中含镶量为 10"""'100 毫克/ 公斤(1)。主要矿石是肿化物和硫化物。在矿石加工以及生产 和使用媒的过程中可造成环境污染。镇被用作某些合金和金
属板催化剂、电池、杀菌剂的成分，使用含媒的食品加工设 备也能引起食物的某些污染。
12.1.2
水中存在的锦
s 很多镇盐溶于水，故可造成水的污染，主要问题是由于 含镰化合物的工业废水向河中排放所致。巳报道在地面水中 镇的浓度可高达 1 毫克/升(1)，但是一般而言，浓度低得多， 仅为 5"""'20 微克/升。3. 巳发现，苏联地下水水源含镇浓度 达 0.13 毫克/升 E33.
12.2 12.2.1 饮用水
接触途径
常规水处理能去除一定量的镇，所以处理后的水中镇含 量通常低于未处理7]< (2) 。儿乎尚未对龙头水中镇的含量进行
综合性调查。有限的资料表明，浓度为 2.......5 微克/升是相当 一 170 一
普遍的Cl -5) ，有时也发现较高的浓度 (6) ，特别是当使用含壤
的设备时。此外，也有个别报道饮水中的浓度达 0.5 毫克/ 升 (4 ， 7) 。假设每人每日饮水量为 2 升，那么从饮水中摄入的
镇一般不会超过 10-20 微克/日。
12.2.2
盒物
在大部分食物中含有镇，但是含量低于〈常常远低于) 1 毫克/公斤 (8) 。虽然已知部分操可能与肌醇六磷酸结合，
但是关于食物中镇的化学形式了解甚少阳。从膳食中的摄入 量波动在小于 200 至 900 微克/日 (1 ， 2 ， 8-10) 。典型的膳食可
能提供约 400 微克/日。据报道，在葡萄酒和啤酒中镰含量 分别为 100 和 50 微克/升 ω。
12.2.1
空气
几乎没有关于空气中镇浓度的资料。但是，似乎空气中 的浓度通常低于 0.5 微克/立方米。过去曾报道空气中含有 较高浓度的镇，主要与工业区有关 ClH23. 在典型的城市空气
中镶的浓度为 0.2 微克/立方米。
12.2.4 12.2.4.1
其他接触途径
职业接触
巳报道在工业环境空气中镰浓度可达 400 微克/立方 米 E13，但是在工业中一般接触水平要低得多。在某些情况 下，工人主要的接触途径可能是其职业环境。 吸烟
12.2.4.2
据报道，香烟〈一般含镇量约为 3 微克/支〉中含镇量的
大约 10"""'20% 能被吸入 E132. 可能主要是一种挥发性镇化合 94
物，即棋基镇。 3 。每天吸 20 支香烟的人每周一般的摄入量 可能是 40~80 微克镇。
12.2.5
不同接触途径的相对意义
成人每周摄入镇的估计值如下a (a) 每周从食物、空气和水中的摄入量 水中锦 的浓度
每周镰的摄 λ 量(微克) 水 空气 食物 总量
比值 水/总量(%)
50微克/升 75 微克/升 ‘ 100 微克/升 150 微克/升
700 1050 .1 垂 00
28 211 28 28
3150 3150 3150 3150
3878 4228 4578 5278
18.0 25.0 30.6 40.0
2100
假设 z 每天从食物中镜的摄 λ 量为 450 微克; 量为 0.2 微克/立方米。
每天呼吸量为 20 立方米，镰含
( b) 每周从食物、空气和:水中的吸收量 水中镰
每周镇的吸收量(微克) 水
此值 总量
的浓度
空气
食物
水/总量(%)
50徽克/升
7.0 10.5 14.0 2 1. 0
14 14 14 14
3 1. 5 3 1. 5 3 1. 5 3 1. 5
52.5 56.0 59.5 66.5
13.3 18.8 23.5 3 1. 6
75 徽克/升 100微克/升 150 徽克/升
假设 I 50% 吸 λ 的镣被吸收;
1% 食 λ 的镰被吸收。
12.3
代
谢
几乎可以肯定地说，镇是动物的必需营养素，因此可能 也是人所必需的口气镇经胃肠道吸收很少，即 1% 或更低 E13 ， 虽然也有报道吸收率较高， I'!
p 10% (8) 。几乎没有证据表明
镇在各种组织中蓄积 (12) 。给大鼠饮用含操浓度为 5 毫克/升
的饮水未见明显蓄积。显然，至少在动物体内具有能限制过 量摄入键的机制。 人的某种疾病状态确实能引起组织中镇的增高 F 但是尚 不清楚原因何在∞。
镇易于排出，主要通过粪，小量通过尿排泄，汗液也能 排出相当一部分 (8) 。
12.4
对健康的影响
镇是相对无毒的元素。在食物和水中通常发现的浓度不 会造成严重的健康危害 0 ， 8) ，但是，在早期的动物实验中高
剂量(在词料中的含量为 1 ， 600 毫克/公斤〉能引起最低的 毒性效应〈断乳的小鼠数减少) (8) 。在以后的三代生殖试验
中这种作用并不明显。给大鼠和小鼠终生饮用含镇浓度为 5 毫克/升的饮水，未见有害作用 (13) 。 动物实验表明某些镇化合物是致癌的口， 6) ，而目前，认为
可溶性镰化合物对动物和人都不致癌 (2)。正如其他的二价阳
离子一样，体外致突变试验表明，镇能与 DNA 反应，高 浓度时能引起 DNA 损伤口 ARC ，个人通信) ,
皮炎是最常见的与职业接触键有关的问题。但是皮肤接 触货币或珠宝手饰也观察到同样的作用。高浓度的职业接触 与肾脏问题有关，而且也见到眩晕和呼吸困难等作用 ω 。
对大量病人调查了接触性皮炎对发生在手上湿密的作 用 F 发现 4~9% 的病人对镇的斑贴试验呈阳性反应(14)。妇 女对镶较男人敏感十倍 (5) 。这些研究对检查湿彦病人仅具
有有限的价值，但是不能完全反应在一般人群中过敏或接触 性皮炎的发生率 E162， -P3~
参考文献 1. Commission of the European Communities. Trace
meta!s: exposure and hea!th ef/ects. Oxford , Pergamon Press , 1979. 2. National Research Counci l. Drinking water and
hea!th. Washington , DC , National Academy of Sciences , 1977.
3. Sidorenko , G. I. & It skova , A. I. Nickel. Moscow , Medicina , 1980. 4. Durfor , C. N.
&:
ßecker , E. Pub!ic water supplies
0/
the 100 !argest cities in the United States , 1962. Washington , DC , Government Printing Office , 1964 (Geological Survey Water Supply Paper 1812).
5. Kopp , J. F. & Kroner , R. C. Trace metals in waters
0/
the United States. Cincinnati , US Department of
the Interior , 1967.
6. Cadmium , nicke/ , some epoxides , miscellaneous industrial chemica/s and general considerations on vo/atile anaesthetics. Lyon , International Agency for Research on Cancer , 1976 (I ARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo I. 11).
7. Kopp , J. F. The occurrence of ttace elemente in water.ln. Hemphill , D. P. , ed. Proceedings
0/ the
Third Annual Con/erence on Trade Substances in Envi ronmental H ea!th , 1969 , Columbia , University of Missouri , 1970 , pp. 59-73. 8. Underwood , E. J. Trace e!ements in human and anilllal
nutrition. New York , Academic Press , 1977. ~174-
9. Hamilton , E. 1. & Minski , M. J. Abundance of the chemical elements in man's diet and possible reIations with environmental factors. Science 01 the
total environment , "375 (1 973). 10. Masironi , R. How trace elements In water contribute to health. WHO chronicle , 32:382 (1 978). 11. WHO Technical Report Series , No. 532 , 1973 (Trace
elements in human nutrition. report of a WHO Expert Committee) .
12. Environmental Protection Agency. Water qua 1i ty criteriaJ ava i1 ab i1 ity. Fed eral register , 44: 43684 (1 979) .
13. Schroeder , H. A. et al. Long-term effects of nickel in ratsa survival , tumors , interactions with trace elements and tissue levels. J ournal 01 nutrition , 104'239 (1 974).
14. W i1 kinson , D. S. et al. The role of
CO l). tact
allergy in
hand eczema. Transactions 01 the St. J ohn's
Hospital Dermatological Society , 56:19-25 (1 970). 15. Fisher , A. A. & Shapiro , A. Allergic eczematous contact dermatitis due to metallic nicke 1. J ournal 01
the American Medical Association , 161 :717-721 (1 956) •
16. Kaalter , K. et a 1. Low nickel diet in the treatment of patients with chronic nickel dermatitis. British
journal 01 dermatology' 98:197-201 (1 978).
-175 一
13. 硝酸盐和亚硝酸盐 这两种化合物往往被一并考 J!P.，因为在环境中易于从一 种形式转变为另一种形式。硝酸盐对健康的影响通常是由于
在体内转变为亚硝酸盐的结果。 在水中的浓度表示为毫克/升(硝酸盐氮和亚硝酸盐 氮〉。
13.1 13. 1. 1 来源
概述
硝酸盐以一定的数量广泛存在于土壤、大多数水中和植 物内，包括蔬菜(1)。亚硝酸盐也广泛存在，但一般而言，其 浓度远低于硝酸盐 (1) 。硝酸盐是通过细菌的作用使有机氮氧
化的产物，这些细菌存在于土壤和氧气充足的水中。亚硝酸 盐是从有机氮通过细菌的不完全氧化而形成的 (2) 。硝酸盐的
主要用途之一是作为肥料，而大部分其他的含氮肥料是在土
壤中转化为硝酸盐 ω 。硝酸盐也用于爆破材料、化学工业的 氧化剂以及食品防腐剂 (2) 。亚硝酸盐(→般是铀盐和锦盐〉
的主要用途是用作为食品防腐剂(1)。环境中的某些硝酸盐是
由于大气中氮气被固定在土壤中产生的(细菌的合成作用〉。 而某些硝酸盐和亚硝酸盐的形成是由于闪电时或人为产生的 氮的氧化物经雨水冲洗的结果 (2) 。此外硝酸盐和某些亚硝酸
盐是由于动物和植物的有机物被细菌分解而在土壤中形成 一 176~
的。
由于硝酸盐和亚硝酸盐广泛地分布于环境中，因此在大 多数食物、大气和很多水源中都有发现。
13. 1. 2
水中存在的硝酸盐和亚硝酸盐
施肥、腐烂的动植物材料、生活污水、向田地施用污
泥、工业废物、垃坡堆〈从应坡堆放厂的渗漏〉和大气降落 物都是水源中这些离子的来源 0 ， 3 ， 4) 。改变土地的利用同样
可增加硝酸盐的水平。根据具体的情况，这些来源可污染小 溪、河流、湖泊和地下水，特别是井 (3) 。污染可能是由于直
接或间接的排放，或者在一段时间内有时是多年的渗透引起 的。在污染的水中硝酸盐含量几乎总是比亚硝酸盐含量高得 多(1)。
水中硝酸盐含量通常低于 5 毫克/升〈硝酸盐氮)，但是 在一些小水源巾硝酸盐的含量超过 10 毫克/升。在经加氯消
毒的给水中，亚硝酸盐含量通常低于检出限，即小于 0.005 毫克/升(亚硝酸盐氮) (~-4)，但在未经加氯消毒的水中，亚
硝酸盐含量可能很高。当水中亚硝酸盐的浓度高时，通常微
生物的质量是不令人满意的。 许多研究揭示，水中硝酸盐含量的范围可能达到 2(} 到 200 毫克/升(硝酸盐氨) (3) ，但是这种情况很罕见。大部分 较高的硝酸盐含量见于地下水中 (3) ，然而在地面水中的硝酸
盐几乎完全被水生植物所消耗 ω 。水中硝酸盐含量的增加与
施用氮肥有关。河流中硝酸盐含量可呈明显的季节性变化， 可能出现较高浓度的硝酸盐，特别是在严重干旱季节的大雨 后 (5)。但地下水中硝酸盐含量在一年期间恒定得多。
一 177 一
13.2 18.2.1 饮用水
接触途径
鉴于常规水处理和消毒过程都不能使水中的硝酸盐含量 发生很大变化，而且在配水系统中硝酸盐含量也没有明显改
变，所以自来水中的含量通常很近似于水源中的含量。但自
来水中亚硝酸盐的含量可能明显地低于水源水，因为在水处 理的过程中〈尤其在水氯化时〉亚硝酸盐被氧化 (2)。很难确
定水中硝酸盐或亚硝酸盐的范围和平均含量，因为该含量随 水源不同而变化很大。对于全世界的大多数人来说，接触的 浓度可能大大地低于 5 毫克/升(硝酸盐氮) (3). 而对于少数
人，通常居住在偏僻地区，接触的浓度可高达 100 毫克/升 或以上(硝酸盐氮)。假设每天水的摄入量为 2 升，那么每天 接触的硝酸盐量通常在 20 毫克(硝酸盐氮〉以下 z 但是在少 数情况下，接触硝酸盐量可能是这个值的 5 倍以上 (3) 。
18.2.2
盒物
在某些食物中存在着相当量的硝酸盐和较少的亚硝最 盐，通常食物是摄入这两种盐的主要来源 (0 。在一些庄稼 中，硝酸盐含量可高到 100 毫克/公斤。在一些诸如洋白菜、
芹菜、离宦、土豆、许多根茎蔬菜和夜菜等蔬菜中含有相当 高水平的硝酸盐，但仅含有少量的亚硝酸盐 (4) 。硝酸盐和亚
硝酸盐作为防腐剂被加入某些食物中，特别是某些肉类和乳 酷中 (2) 。
摄入硝酸盐和亚硝酸盐另一个重要来源是唾液 ωs 人类
分泌硝酸盐的量约为 10 毫克/日〈硝酸盐氮)，其中约 2 毫 -178 一
克硝酸盐氨被还原为亚硝酸盐(4)。唾液中的硝酸盐主要来自 于食物 (6) ，特别是来自于蔬菜。
从食物中摄入硝酸盐和亚硝酸盐量的波动范围是非常大 的(4)，例如，几乎不吃蔬菜和熏肉的人对硝酸盐和亚硝酸盐 的摄入量极低 (4) 。据估计，各种典型膳食对硝酸盐的摄入量
范围从约 120 毫克到 230~300 毫克/日(没有估计亚硝酸 盐) (4汀， 830 鉴于居民中婴儿对硝酸盐最敏感，因此限制婴儿
接触量很重要。对于两个月的婴儿来说，单纯从食物中每天 摄入的硝酸盐量估计大约为 25 毫克〈硝酸盐氮) (2) 。
18.2.3
空气
除了空气中的氮氧化物和硝酸盐的天然来源外，还有一
些重要的人为来源，特别是矿物燃料(煤、石油、天然气〉 的燃烧产物和来自化学工业的产物。这些硝酸盐分为无机和 有机形式两类，这些盐类和氮的氧化物可经呼吸系统吸入，
而其中一部分被吸收，在体内产生硝酸盐和亚硝酸盐的一种 混合物。据估计对于空气中含氮化合物高的地区，如果所有
这些化合物都能被成人所吸收，那么摄入的总量约为 0.1 毫 克/日〈硝酸盐氮)。
13.2.4 13.2.4.1
其他接触途径 职业接触
在工业上氮的氧化物是很普遍存在的，虽然可能出现硝 酸盐和亚硝酸盐的气溶胶，但其含量一般来说是低的。美国 工厂车间空气中二氧化氮最高容许浓度 约 30 毫克/日〈硝酸盐氮〉。 一 179 一
(8 小时工作〉为
5ppm(9) ，这个值相当于在特殊情况下接触硝酸盐含量高到
13.2.4.2
吸烟
在香烟燃烧时生成氮氧化物和硝酸盐气溶胶，但是与食 物和水相比，由于吸烟接触的剂量还是比较小的 (2) 。
13.2.5
不同接触途径的相对意义
对于不同的个体来说，可能从水、食物和空气中接触硝
酸盐和亚硝酸盐的量差别很大。一般来说，空气污染是不太 重要的来源，而食品、唾液和水通常是硝酸盐和亚硝酸盐的 主要来源。飞考虑了食物中硝酸盐的两个含量水平，而不考
虑从唾液中的摄入比例，计算了从硝酸盐含量为 10 毫克/升 (硝酸盐氮〉的水中摄入硝酸盐的比例。 这个评价是基于 s 硝酸盐和亚硝酸盐从食物和水中能
100% 被吸收 z 摄水量对于成人为 2 升/日，对于儿童为 1 升/ 日。而且假设从吸入、吸烟、职业性接触及其他接触途径摄
入的硝酸盐和亚硝酸盐忽略不计。 每天从食氮 物中 λ 量(毫克) 每周睛酸盐氮的吸收量(毫克) 水·
硝酸盐的摄
食物
总量
比例 水/总量
(%)
成人一例 1: 成人一例 2:
20 70 25
140 140 70
140 490 175
280 630 245
50 22 28
儿童一例 3:
"1 0 毫克/升(硝酸盐氮计)。
13.3
代谢
关于硝酸盐在体内的代谢尚不完全了解，似乎硝酸盐的
吸收发生在小肠的上部，一般情况下主要经肾脏排泄，但不 -180 一
完全经肾脏排泄(4)。巳知硝酸盐在上胃肠道吸收，最后被唾 液腺蓄积于唾液中(4)。硝酸盐在人体内的代谢没有得到充分 的研究，当把动物实验结果外推到人时还不是很可靠的 C430
总之硝酸盐和亚硝酸盐很容易在体内吸收。 一个很值得考虑的是在体内硝酸盐易被细菌还原为亚硝
酸盐(4)。已表明，接触高剂量硝酸盐能引起唾液中亚硝酸盐 浓度的明显增加(4)。但是个体间差异很大 E 这可能与其口腔 微生物群及其膳食成分方面的差异呈函数关系 (4) 。硝酸盐被
还原为亚硝酸盐也发生于体内其他部分〈包括胃 h 如果 pH 不大于 4.6 几乎就不发生这种转化(4)。在婴儿通常胃酸很 低， pH 约为 4 或以上 (3 ， 4) ，所以能产生高浓度的亚硝酸盐。
相反，成人胃的酸性 pH 1.-......5 ，硝酸盐很少被转化 (4) 。
生成亚硝酸盐是特别重要的，其理由 z 第一，亚硝酸盐
能使血红蛋白氧化变成高铁血红蛋白，后者是一种色素，但 它不具有携带氧的功能，第二，在某些情况下，一些亚硝酸 盐能在人体内与二级和三级股和氨化物(通常来自于食物和 其他源泉〉反应生成亚硝膀，而具中某些亚硝股被认为是致 癌物(1)。上述反应发生在人胃内正常酸度范围所特有 pHl'-""" 5 的酸性洛液中(4)。当 pH 为 3.5 或低于 3.5 时，其反应率
最大。
13.4 13.4.1
对健康的影响
高铁血红蛋白血症
通常人体内 1.-......2% 的血红蛋白是以高铁血红蛋白的形 式存在的。但是当后者超过 10% 时就会出现临床表现(高铁 血红蛋白血症)，若超过 30.-......40% 出现缺氧症状。 ~.-
一 181-
文献中已在充分的报道，在一些国家由于所供应的水含 有高浓度的硝酸盐，引起婴儿高铁血红蛋白血症和死 亡 (2 ， 10) 。在一篇世界卫生组织的文献中已经综述了世界范围
的问题\并建议不应使用含有高硝酸盐含量(大于 100 毫 克 N0 3 / 升〉的给水制备婴儿的食品，而推荐使用另外的含 有低硝酸盐的水，甚至使用瓶装的饮水 (5) 。婴儿对硝酸盐敏
感是由于婴儿按体重计算具有较高的进水量 (3) 。在上胃肠道
内存在着硝酸盐还原菌以及胎儿血红蛋白(在出生后最初的 几个月内是以这种形式存在〉易于被氧化(11)。在成年人 rjl
不出现高铁血红蛋白血症的问题。当婴儿胃肠道紊乱时能使 硝酸盐转变为亚硝酸盐的细菌数增加，从而增加了敏感 性问 12) 。同时认为奶粉加水，与其他形式的牛奶相反，能增
加对水中硝酸盐含量的敏感性。婴儿胃中的 pH 大约是中
性的，能使细菌在胃和肠的上部生长。与成年人比较，婴儿 还缺乏能使高铁血红蛋白转变为血红蛋白的两种特异性 酶 (3) 。
发生婴儿高铁血红蛋白血症最普遍的原因是由于制备婴 儿食品的水中含有过量的硝酸盐 (3) 。长时间煮沸的水可加重
这一问题，因为过量蒸发而增加了硝酸盐的含量。绝大多数 婴儿高铁血红蛋白血症病例是由于使用被微生物污染的私人 井水所致∞。 对于引起高铁血红蛋白血症的水中硝酸盐含量已进行了 大量的研究，但是对于这个作用的阔值还存在着矛盾的结 果 (3 ， 13) 。在饮用水中硝酸盐含量一直低于 10 毫克/升(硝酸
盐氮)的地区，未见婴儿高铁血红蛋白血症的病例报告 E230 a 次材井水中南浓度的硝酸盐引起的婴儿育紫症。世界卫生组织妇幼卫生专
家委员会，
1949( 未发袤的文件 WHO/MCH/13:1 的。
一 182-
很多婴儿使用比该值高得多的水也未见发病 s 仅仅全部病例 的 2.3% 是发生在饮用水中硝酸盐含量在 10 和 20 毫克/升 (硝酸盐氮〉之间(1)。因此对于在这种浓度下硝酸盐的作用 尚有一些怀疑。但是，尽管在这种浓度下婴儿高铁血红蛋白
血症的临床表现不明显，然而确实出现了血液中高铁血红蛋 白含量的升高 ω 。
同样还指出，任振妇女比一般成年人有较大的危险 性(1) ，但是需要进一步的工作确证这点。
虽然高铁血红蛋白血症易于诊断，对于在具有充分医疗 设备的地区可能是没有问题的 F 但是对于缺乏医疗设备发展 中的地区可能问题就比较大了。
13.4.2
.iI!硝脏的鼓癌性
由于摄入的硝酸盐，无论在嘴里还是在体内酸度较低 (pH 高)的其他部位，都能容易地被转变为亚硝酸盐，因
此有可能生成亚硝膀，而某些亚硝胶可能是致癌物。已经表 明，具有膀胧感染和胃酸缺乏的人〈一种低胃酸的情况)亚 硝腔的生成可能增加。在膀胧感染时有可能生成的亚硝胶将 被吸收进入血液中 E230
尽管动物实验表明，很多亚硝股是致癌的，但还没有直 接的证据证明它对人具有致高性 (1 ， 4 ， 15) 。尽管有一些研究，
但是关于水中硝酸盐可能与癌症有关的证据还不是有说服力 的问 16) 。硝酸盐经生成亚硝股致癌的证据取决于流行病学
研究。还没有与人有关的恰当的动物实验研究。在一篇关于 中国胃癌的综述 (17) 中报道，在福建省蒲田地区胃癌的死亡率
最高(1 20--147/10 万男人〉。资料表明，在这个地区的饮用 水和蔬菜中的硝酸盐和亚硝酸盐含量均高于低病区。关于这 一 183~
个病的流行病学和病因学的研究正在高危险区和低危险区进 行着。 参考文献
1. Nitrates , nitrites and N-nitroso compounds. Geneva , World Health Organization , 1978 (Environmental Health Criteria 5).
2. Guidelines /or Canadian drinking water quality , 1978. Quebec , Ministry of Supply and Services , 1980 (supporting documentation). 3. Nitrates in water supplies. Report by the International Standing Committee on Water Quality and Treatment.
Aqua , 1 :5-24 (1 974). 4. National Research Counci l. Drinking water and
health. Washington , DC , National Academy of Sciences , 1977.
5. Royal Commission on Environmental Pollution. London , HM Stationery Office , 1979 , Chapter 4. 6. Tannenbaum , S. R. et a l. Nitrite in human saliva.
It s possible relationship to nitrosamine formation.
Journal (1 974) .
0/ the National Cancer Institufe , 53:79
7. White , J. W. Relative significance of dictary sources of nitrate and nitrite. J ournal
0/ agricultural and
/ood chemistr y , 23: 886 (1 975). 8. Phillips , W. E. J. Change in nitrate and nitrite content of fresh and processed spinach during storage. J ournal 16:88 (1 968). -184 一
0/
agricultural and /ood chemistr y ,
9. TLVs-Threshold limit values /or chemÎcal substances and physical agents in the workroom environment with intended changes for 1976. Cincinnati , American Conference of Governmental Industrial Hygienists , 1976.
10. Health effects of nitrates 77-030) .
in ω ater. Cincinnati , US
Environmental Protection Agency , 1977 (EPA-600/111. Betke , K. et a l. Vergleichende Untersuchungen über die Spontanoxydation von Naberschnur-und Erwachsenenhiimoglobin. Zeitschrift für
Kinderheilkunde , 77:549 (1 956). 12. Shuval , H. l. & Gruener , N. Epidemiological and toxicological aspects of nitrates and nitrites in the environment. American journal of public health. 62: 1045 (1 972). 13. Frazer , P. & Chilvers , C. Health aspects of nitrate
in drinking water. Netherlands , 1980 (paper presented at the International Symposium on Water Supply and Health) • 14. Hill , M. J. et al. Bacteria , nitrosamines and cancer of the stomach. British journal of cancer , 28: 562 (1 973).
15. Some N-nitroso compounds. Lyon , International Agency for Research on Cancer , 1978 (l ARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , Vo l. 17). 16. Frazer , P. et a l. Nitrate and human cancerl a review of the evidence. International journal
0/
epidemiology , 9'3 (1 980).
17. Xu Guang-We i. Gastric cancer in China , a review.
Journa/ (1 981) .
0/
the Roya/ Society
0/
Medicilie , 74:210
一 186-
14. 晒 14.1 概述
14. 1. 1
来源
由于地质化学的差异，土壤和植物中晒的含量变动很 大(1)。晒的化学式和溶解度对其进入食物链以及存在于水中
都是决定性的因素。环境的作用能降低晒的溶解度，使可溶 性晒酸盐化合物或某些亚晒酸盐转化为溶解度很低的化合 物，例如元素晒或陋化物〈或甚至某些金属的亚晒酸盐) (2)。
通常晒以亚晒酸盐或晒酸盐的形式存在于水中，而 pH 和某 些金属(如铁〉盐的存在均可影响其化学形式 (2) 。
14. 1. 2
水中存在的画
已有数篇综述(1-7)收集了世界上不同地区的资料，结果
表明，在大部分分析的地面水样品中晒的浓度远低于 10 微 克/升。在 4 年期间从美国主要水域收集的 535 个样品中仅
有 2 个样品超过该值，最高值为 14 微克/升。在苏联非高晒 区; -晒的浓度范围从每升数十分之几微克至数微克，最高值
为 5.1 微克/升。在阿根廷的 22 个地面水中晒的浓度从小于
2 至 19 微克/升，中值为 3 微克/升。 有一项研究分析了美国克罗拉多州的 42 个地面水样品，
晒的浓度范围从小于 1 至 400 微克/升，中值为 1 微克/升。 -187~
在苏联最靠近黄铁矿矿床的乌拉尔山区排出的水中也含有类 似高浓度的晒。流经产晒区土壤的灌 j既水能使地面水(河水〉 中的晒浓度增高。一些泉水和浅井水含晒浓度超过 100 微克/
升，在美国南达科他的一个产晒区某些井水中晒的浓度高达 330 微克/升。， 5) 。
14.2
接触途径
除职业接触时经空气接触和皮肤接触特别重要以外，一 般人群主要是通过食物接触。
14.2.1
饮用水
在加拿大和美国各种公共给水系统及澳大利亚和联邦德 国的某些农村给水的龙头水中 F 晒的浓度只有个别样品超过 0.01 毫克/升 (2-8) 。在产晒区，特别是井水中晒的浓度较高。
但是，对于世界高晒区饮水中晒的含量进行系统的研究还很 少。
因此，象美国或加拿大等国，从饮水中摄入的晒通常不 超过每日从膳食中摄入量的 5 .......， 10% ，一般而言，在大部分
地区远低于该值。对产晒区的居民来说，饮水往往不是摄入
晒的唯一和主要来源。 14.2.2 盒物
从膳食中摄入的晒量取决于饮食类型和食物中晒的含
量，后者主要由食物特性和地质化学状况决定。 从蔬菜和水果食入的晒一般很少，而谷类、谷类产品、
肉(特别是内脏)和海产品却含育丰富的晒，一般含量远高
于 0.2 毫克/公斤湿重。土壤的化学组分和含晒量对来自不
-188-
同国家的谷类含晒量具有明显的影响，其含量范围为 0.04~ 21 毫克/公斤 (4 ， 6 叶飞
近年来已对不同国家每日从膳食中摄入晒量的资料进行 i 了综述，晒的摄入量范围，从 56 微克/日(新西兰低晒区〉
至 320 微克/日以上〈委内瑞拉，是一个在土壤和植物申含晒 量很高的国家) (7 ， 9) 。但是另一方面，有记录表明，在新西兰
某些很健康的人每日晒的摄入量低至 20 微克/日(1 0) 。营养
调查表明 F 例如在美国或加拿大等国，在成人中普通的膳食 每天可提供约 100~200 微克晒 (7 ，的。
14.2.3
空气
从现有的关于大气和烟草中晒含量的资料表明，对一般 人群而言，吸入的晒在每日晒的摄入量中不占很大比重 E4 ，730
14.3
代谢
可溶性晒盐，例如亚晒酸铺， .易于在大鼠胃肠道内被吸 收。无论膳食中含晒量为 20 或 4000 微克/公斤，其吸收率均 超过 95% (11)。当从水溶液中给人以毫克剂量级的亚晒酸铺
时，约吸收 93% 。由此可见，象大鼠一样，人也没有体内平 衡控制机制以限制大量亚晒酸制经胃肠道的吸收 (12) 。
吸收的晒广泛分布在器官和组织中，在肝和肾中含量较
高。晒能穿过胎盘，也能进入乳汁中，穿透的能力取决于晒的 化学式 (4 ， 6 ， 13) 。在体内有两种主要的代谢途径 s 一为直接掺入
或与蛋白质结合，另一种是先还原，继而甲基化，产生二甲 晒离子和三甲晒离子。当二甲晒的形成速度超过进一步甲基
化为一种尿中的代谢物三甲晒离子时，挥发性的二甲晒被呼 出。在一般人群接触的情况下，从尿中排泄晒为主 E3-1330 03
晒的排泄速度取决于给予晒的化学形式以及体内晒的营 养状态。现有人的资料表明，以亚晒酸盐形式给予的晒从体 内排泄的速度，比有机晒化合物(例如晒代蛋氨酸〉更快 (10 。
在大鼠中晒的生物半减期，随着膳食中晒水平的增加而缩 短 (15) 。
14.4
对健康的影响和剂量
一反应关系 巳证明晒是数种动物种属的一种必需营养素 (8 ， 16) 。在低
晒区发现家畜出现某些地方病，而且补充晒对预防这些疾病
具有明显效果。当然，如果晒的剂量过高也会出现毒性，引 起家畜的其他疾病 (1 ， 2)。
14.4.1
对人畴的研究
越来越多的证据表明，晒是保持人体健康所必需的元素。 只有在极端的情况下，当长期食用含晒量极低的当地产食物 时才会出现明显的晒缺乏表现。目前对克山病的研究表明， 摄入晒量过低能引起儿童的这种心肌病 (17自 19) 。 Smith 等 (20) 以及 Smith 和 Westfall(2 l)研究了一
组农民，他们居住在美国的高晒区，基本上食用当地产食物，
摄入晒的水平高达 200 微克/公斤体重/日。出现的症状和体 征是非特异性的。然而在一组摄入高晒的 100 人中，有 31 、 28 和 27 人分别出现胃肠功能紊乱、皮肤黄瘟和牙齿不好。 Jaffé 在委内瑞拉(是一个土壤和植物中含晒量高的国 家〉进行了另一项研究 ω ，233 ，将接触高晒的儿童与加拉加
斯的一组儿童(血晒和尿晒水平均较低〉加以比较，结果发 现第一组儿童的血晒是迄今在世界上一般人群中含量最高
的 (8)0 Jaffê 认识到，这两组委内瑞拉人不仅晒的摄入量不
同，而且其他因素，包括营养状况和寄生虫感染的情况也不 同。在高晒区的儿童比加拉加斯的儿童中更常见到的恶心、 皮炎和指甲的病理改变。但是，所见改变，特别是生长延滞 和贫血可能是由于其它因素所致。
曾进行了数项研究试图找出在不同人群中踊齿患病率高 与晒的接触量高以及在某些区域癌症发病率低与晒的接触量 高之间的相关关系〈见综述 4 ， 24) 。但是，这些研究未排除 其它干扰因素，而且在其他方面，特别是关于晒的接触问题 也受到批评邸， 26) 。
14.4.2
动物研究
在大多数受试动物种属中，基本的膳食需要约为 0.04-0.10 毫克/公斤饲抖 (8) 。但是维生素 E 的缺乏明显地增加了
对晒的需要量，而且也应考虑数种其它营养素的相互作 用
膳食晒的含量为 5 毫克/公斤食物或以上能引起慢性中
毒，而且该数值被当作在高晒区有毒和无毒食物之间的分界 线(4)。这一结论是基于对高晒区生长的家畜的现场经验以及
很多已有的实验动物研究资料得出的。摄入过量晒对动物的
主要作用表现为体重增长减慢、存活率降低以及损伤肝和其 他器官，有时也损伤心肌、肾和膜。
在一些动物实验中报道了长期食用含晒量低于上述水平 的饲料所引起的作用"飞给大鼠喂词半纯化饲料，按 0.5--
2.0 毫克/公斤饲料补充晒，结果发现实验组大鼠肝实质增生 较对照大鼠更多些 (27) 。
血液中谷胧甘肤浓度增加、肝中玻咱酸脱氢酶活性降低
以及肝脏排泄功能的某些损伤可能与长期摄入低水平的亚晒
酸纳有关，在这种接触水平的情况下还观察到很多行为的改 变 E2830
在两篇关于高晒饮用水毒性作用的报告之间有明显差 异。其中一篇报告，雄性大鼠饮用含亚晒酸盐浓度为 2 毫克/
升(以晒计〉的饮水，在不到 3 个月的时间内引起 50% 动物 死亡，而对雌性大鼠的毒性作用不明显 (29) 。但是，在另一篇
报告中，饮用含亚晒酸纳浓度为 3 毫克/升(以晒计〉的饮水 对雄性大鼠的存活率却无影响 (30) 。已经表明，晒化合物对于 食用比较高晒饲料的动物具有较小毒性剧， 32) 。给猴喂饲致
踊齿饲料和饮用含亚晒酸纳浓度为 2 毫克/升(以晒计〉的 饮水 15 个月，随后改饮用浓度为 1 毫克/升的饮水 45 个月， 结果表明，当牙齿发育时晒有致踊齿作用 F 而出牙以后再接 触晒，则无此作用 (3) 。
尚无充分证据支持摄入高晒对实验动物具有致癌作 用 (34) 。最近的一项研究揭示了小鼠长期接触二硫化晒具有
致癌作用 E53 ，但是对其实验设计和统计学分析所存在的缺陷
尚未予以评击。另一方面，有数篇报告表明，晒的剂量高于 营养需要量时对实验动物具有防癌作用 ω 。 最近已经发表了一篇关于晒对实验动物抗癌作用的综 述 (35) 。
参考文献 1. Rosenfeld , I. & Beath , D. A. Selenium. geobotany ,
biochemistry , toxicity and nutrition. New York , Academic Press , 1964. 2. National Research Council. Drinking water and n3 nr"
health. Washington , DC , National Academy of Sciences , 1977. 3. Muth , O. H. , ed. Selenium in biomedicine. Westport , 'CN , Avi Publishing Company Inc. , 1967.
4. Selenium. Washington , DC , National Academy of Sciencee , 1976.
5. Ambient water qua/ity criteria for selenium. Washington , DC , US Environmental Protection Agency , 1980. 6. Ermakov , V. V. & Kovalskij , V. V. Biolo !J ical si !J ni /icance House , 1974.
0/
seleni um. Moscow , Nauka Publishing
7. National Research Counci I. Drinkin !J water and
health. vo I. 3. Washington , DC , National Academy of Sciences , 1980.
8. WHO Technical Report Series , No. 532 , 1973 (Trace
elements in human
nut 俨 ition)
.
9. Levander , O. A. In: Proceedings
10.
0/ the symposium on selenium-tel/ urium in the environment , Pittsburgh , Industrial Health Foundation , Inc. , 1976. Stewart , R. D. H. et a I. Quantitative selenium metabolism in normal New Zealand women. British
journal
0/
nutrition , 40:45 (1 978). 75SeO~
1 1. Brown , D. G. et a l. Effect of dietary selenium on the gastrointestinal absorption of in the rat.
International journal /or ritamin and nutrition research. 42:588 (1972). 12. Thomson , C. D. In. Trace elements in human and animal health and disease in New Zealand. Hamilton , Waikato University Press , 1977. 。。
13. Diplock , A. T. Metabolic aspects of selenium action and toxicity. CRC critical revieω s in toxicology , 4:271 (1 976). 14. Griffithe , N. M. et al. The metabolism of (75 Se) selenomethionine in four women. British journal
0/
nutrition , 35:373
(1 976).
15. Burk , R. F. , Jr. et al. Influence of dietary and injected selenium on whole-body retention , route of excretion , and tissue retention of 75S e 02- in the rat.
J ournal 0/ nutrition , 102: 1049
(1 972).
16. Schwarz , K. et a l. Introduction. Symposium on nutritional significance of selenium (Factor 3).
Federation proceedings , 20:665
(1 96 1).
17. Keshan Disease Research Group of the Chinese Academy of Medical Sciences. Observations on the effects of sodium selenite in the prevention of Keshan disease. Chine St medical journal. , 92:471 (1 979) •
18. Keshan Disease Research Group of the Chinese Academy of Medical Sciences. Epidemiological studies on the etiologic relationship of selenium and Keshan disease. Chinese medical jou 俨 nal ，
921477 (1 979).
19. Xiaoshu Chen et al. The relations of selenium and Keshan disease. Biological trace element research , 2'91 (1 980). 20. Smith , M. 1. et a l. The selenium problem in relation to public health. A preliminary study to determine the possibility of selenium intoxication in the rural population living on seleniferous 80il. Public health
reports , 51: 1496 -194 一
(1 936).
2 1. Smith , M.
1. & Westfa !l, B. B. Further field studtes
on the selenium problem in relation to public health.
Public health reports , 52:1375 (1 937). 22. Jaffé , W. G. et a l. Estudio clinico bioquimico en ninos escolares de una zona selinifera. Archivos
latinoamericanos de nutrición , 22:595 (1 972). 23. Jaffé , W. G. Ina Proceedings
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the symposium on
selenium-tellurium in the environment , Pittsburgh , Industrial Health Foundation , Inc. , 1976.
24. Glover ,1. et al. Inl Friberg , L. et al. , ed. Handbook
on the toxico!ogy 25. Allaway ，顶T
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metals , Amsterdam , EIsevierj
North-Holland Biomedical Press , 1979. H. An overview of distribution patterns of trace elements in soil and plants. Annals
0/ the
New York Academy
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Sciences , 199:17 (1 972).
26. Schwarz , K. Inl Muth , O. H. , ed. Selenium in
biomedicine. Westport , CN , Avi Publishing Company Inc. , 1967 , pp. 225-226.
27. Harr ,
1. R. et a l. In. Muth , O. H. , ed. Selenium in
biomedicine. Westport , CN , Avi Publishing Company Inc. , 1967. 28. Pletnikova ,1. P. Biological effects and safe concentrations of selenium in drinking water.
H ygiene and sanitation , 35:176 (1 970). 29. Schroeder , H. A. & l\litchener , M. Selenium and tellurium in rats. Effects on growth survival and tumors. J ourna! 0/ nutriti on , 101: 1531 (1 97 1). 30. Palmer , I. S. & Olson , O. E. Relative toxicities of selenite and selenate in the drinking water of rats.
Journal
0/
nutrition , 104:306 (1 974).
31. Ja ffé , W. G. & Mondragon , M. C. Adaptation oí rats to selenium intake. Journal (196ω-
0/ nufrition , 97'431
32. Jaffé , W. G. & Mondragon , M. C. Effects of ingestion of organic selenium in adapted and non-adapted rats. British journal
0/ nutrition , 33:387
(1 975).
33. Bowen , W. H. The effects of selenium and vanadium on caries activity in monkeys (M. i rus). J ournal
0/
the Irish Dental Association , 18:83 (1 972). 34. Some aziridines , N- , S- and O-mustards and selenium. Lyon , International Agency for Research on Cancer , 1975 (IARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo l.
9) , p. 245. 35. Jllcobs , M. M. Effects of selenium on chemical carcinogens. Preventive medicine , 9.362-367 (1 980).
一 196-
15. 15.1 15. 1. 1 来源
号良
概述
银在自然界是以元素状态或在各种矿石(例如辉银矿和
角银)中存在的，也与铅、金、铜和铮矿共生。在地壳中的 含量约为 0.1 毫克/公斤(1)。它用作各种合金和焊剂的成分，
还用于照像、电器设备、电镀以及杀菌剂、银器、珠宝手饰、 货市和牙科用品(1)。由于银具有抑菌作用，所以银盐可用于 水的消毒和预防药剂 E230
15. 1. 2
水中存在的银
在天然水中银的浓度很低 (1) 。虽然尚无足够的资料证实 水中的精确含量，但是已发表的资料。 -3)表明，在水源中银
的含量几乎均不超过 1 微克/升，而大于 10 微克/升的情况更
为罕见。
15.2 15.2.1 饮用水
接触途径
已表明，很多常规水处理措施能有效地从水中去除银， 因此许多处理的水含银浓度很低∞‘但是，由于在配水系统中 一 197-
使用的一些金属(例如铅和铮〉可能含有微量银以及在某些 国家使用氧化银消毒饮水，所以有时在龙头水中银的浓度增 高。据记录，在个别情况下银的浓度超过 50 微克/升，特别 是当使用含银净化器净化饮水时更是如此 ω 。所以龙头水中 银的平均浓度很低，显然低于 1 微克/升。假如每日饮水量为 2 升，从饮水中每日银的平均摄入量可能不超过 2 微克。
15.2.2
盒物
虽然大多数食物仅含有微量银〈小于 1 毫克/公斤)，但 是几乎还没有发表关于食物或膳食中银含量的资料。蘑菇例 外，银的含量为数百毫克/公斤 (3) 。据估计由各种膳官提供的 银为 1~80 微克/日 EL43 ，而使用银具时其摄入量可能高得 多p 用1 含银水烹调的蔬菜能很有效地吸收该金属 (3) 。迄今尚
无足够资料准确地证明从膳食中银的平均摄入量，但是合理 的估计值可能是 20"""'80 微克/日。
15.2.3
空气
已发表的关于空气中银浓度的资料极少，美国报道其浓 度可达 0.1 微克/立方米白， 6) 。因为出于经济理由，严格控制
r 银随工业废气的排放，据报道，在矿物燃抖中的含量也很
低，所以在大气中银的浓度很低。预计在城市空气中通常的 浓度不会超过 0.05 微克/立方米，所以从空气中吸入的量是 微不足道的。
15.2.4 15.2.4.1
其他 职业接触
几乎没有关于职业接触水平的资料。除用铜钟合金进行 一 198 一
焊接和生产银袖的工作外，对银的职业接触很低。
15.2.4.2
药物制剂的使用
局部使用含银制剂可能造成一定量的接触 (7 ， 8) 。
15.2.5
不同接触途径的相对意义
根据在第 15.2.1........15.2.4 节中的考虑，可以粗略地评价 银的平均每日摄入量。
假如水中银的浓度为 1 微克/升，而且每日摄水量为 2 升，则银从水中的摄入量为 2 微克/日。如果银从空气巾的 摄入可以忽略不计，而每日膳食中含银量为 20........80 微克，因 此从各种来源的每日平均摄入量大约是 22........82 微克。
使用含银的抑菌定点净水器，饮水中银的浓度可达 50 微克/升，加上每日从膳食中摄入 80 微克银，因此每日总摄 入量可高达 180 微克/人。 一般而言，饮水中银的浓度很低，通常不超过 1 微克/ 升，故从饮水中摄入的银占各种来源正常总摄入量的比值， 不可能超过 5% 。
目前还没有关于人类对银吸收的精确资料，也不可能提 供摄入情况的数据。
15.3
代谢
除了知道个体和个别器官能选择性地吸收银外 E32，关于 人对银的吸收和代谢了解甚少(1)。动物约能吸收银摄入量的
10%(2) 。在各种器官中都能检出银 F 特别是肝和脾看来能 蓄积银 (2) 。人摄入银后 16 天，肝脏中的银含量能占身体负 荷的 50% 以上 (9) 。吸入的银也能少量地被吸收(2)。银与某些
酶系统的蔬基成分和其他生物学上重要的化学基团结合，因 -199 一
此影响蛋白质的沉着和使某些酶系统灭活 (10) 。动物实验也 表明，在代谢上银与铜和晒相互作用(4)。大部分吸收的银几
乎全部从粪中排泄 F 仅仅小量银长期贮存在组织中，但是皮 肤例外，能蓄积大量银 (3) 。将要被排泄的银在体内的生物学 寿命为数天至数周 (2) 。
15.4
对健康的影响
没有证据表明银是人体所必需的元素。已有致死性中毒
的病例记载，但是只有在较高浓度下才会发生。银的主要作 用表现为皮肤、头发和指甲着色(银质沉着症)。当应用银肿 凡制明作为医药时曾发生过这种病例 E330 一次注射剂量为 1 克的银〈银肺凡纳明〉能引起这一效应 (1 1)。而职业接触银
的工人也曾出现这种作用，但是现在已很罕见 (3) 。有时银质 沉着症能掩盖某些轻微的全身作用 (2 ， 3) 。目前还无证据表明 摄入银具有致癌性 (12) 。
大鼠饮用含银浓度为 400 微克/升或以上的饮水，观察到 肾和肝出现病理学改变。但是难以把这些结果外推到人。如 果假设首先出现的银质沉着症对健康无明显影响，那么就可 以使用银质沉着这一表现来估计安全的接触水平。能引起人 银质沉着症的最低剂量是 1000 毫克银 (13) 。终生 (70 岁〉摄
入 1000 毫克银就相当于每天连续接触 40 微克银。但是，由 于银被不断地排泄 F 并且只有大约 10% 能被吸收，因而引起
银质沉着症终生所需要的每日摄入量可能高达 400 微克。 参考文献
1. Guidelines /or Canadian drinking water quality , 1978. Quebec. Ministry of Supply and Services. 1980 一 200 一
(supporting documentation).
2. Toxicology
0/
metals , vo l. 11. Washington , DC , US
Environmental Protection Agency , 1977 (Environmental Health Effects Research Series). 3. National Research Counci l. Sciences , 1977. 4. Underwood , E. 1. Trace elements in human and Drinking ω ater
and
health. Washington , DC , National Academy of
animal nutrition. New York , Academic Press , 1977. 5. Greenberg , R. R. et al. Composition and size distributions of particles released in refuse incineration. Environmental science and technology , 12' 566 (1 978). 6. Ragaini , R. C. et a l. Environmental trace metal contamination in Kellogg , ldaho , near a lead smelting complex. Environmental science and technology , 11: 733 (1 977). 7. Pariser , R. 1. Generalized argyria. Clinicopathologic features and histochemical studies. Archives
0/
d ermatology , 144: 373 (1 978). 8. Marshall , 1. P. & Schneider , R. P. Systemic argyria secondary to topical silver nitrate. Archives
0/
d ermatology , 11311077 (1 977). 9. Newton , D. & Holmes , A. A case of accidental inhalation of Zn-65 and Silver-110. Radiation
research , 29: 403 (1 966). 10. Goodman , L. S. & Gilman , A. The pharmacological
basis 1975.
0/
therapeutics , 5th ed. New York , Macmillan ,
1 1. H ilI, W. B.
& Pillsbury. D.
M. Argyria. The ~201 →
pharmacology 01 siluer. Baltimore , MD , Williams and Wilkins , 1939. 12. Environmental Protection Agency. Water quality criteria (1 979) .
,
availability. Federal register , 44:15 964
一 202-
, /
16. 纳 a 16.1 概述
16.1.1
来源
铀存在于很多矿物中，主要的一种是石盐(氯化饷〉。海
水含相当高浓度的制。总起来看，在地壳中铀的含量为 26 克/公斤(1)。
纳及其盐类有广泛的用途，包括用于公路除冰，造纸， 玻璃和肥皂工业，制药和一般化学工业，水处理，食品工业 以及烹饪。
纳广泛存在于土壤、植物、水和很多食物中 F 有时含量 相当高。多数国家均有大量的铀矿床。此外，人为地还能排
出大量俐，它是生活污水中的一种普遍成分。
16.1.2
水中存在的铀
铀离子普遍存在于水中，这是由于制盐的洛解度大以及
铀矿床很丰富的缘故。海水含纳量约为 10 克/升。在低地的 河流和地下水中的铀含量是淡水中最高的。而在高地的小溪 流及与其者关的水库中纳含量是相当低的。特别是在富含铀
矿的地区以及由于海水侵入(来自海洋和港湾〉或其他形式 的污染都可能使地下水中制含量增高 (2) 。 凰销对水质其他方面的影响，将在第五章第 11 节进行讨论。(见 414 页) -203 一
靠近沿海的区域，凤所携带的海水飞谏是水中铀的主要 来源之一，因为飞沫可能落到地面进入水源或雨水将其冲洗 进入地面水源 (2) 。水中制的另一个重要来源是废水(生活、
商业和工业废水〉排入河流。在这些河流中铀的浓度与很多 因素有关，包括河水流通，和废水中铀的浓度。有时在一些河
流中铀的浓度很高，特且是当流速低时。
16.2 16.2.1 饮用水
接触途径
在多数国家的大部分给水中含铀量小于 20 毫克/升，但 是在某些国家可能超过 250 毫克/升。除海水侵入和天然污 染外，用于公路除冰的盐、水处理的化学物质、生活用水的
软化、海水飞沫以及污水排放均造成大量的纳进入水中。用 于水处理的化学物质，例如氟化锅、氟硅酸纳、氢氧化锅、 碳酸纳、碳酸氢铀和次氯酸锅，作为单个化学物质，可能使
水中铀的浓度增加不多 F 但是合起来的数量是可观的 E 在某 些情况下，可增加 30 毫克/升 (2) 。生活用水的软化可使水中
浓度增加 300 毫克/升以上，但是一般而言，远低于此 ω 。 大多数人从龙头水中摄入的纳少于 50 毫克/日〈基于每
天水的摄入量为 2 升〉。但是由于纳盐极易溶解，事实上在水
中存在的全部纳，无论是直接摄入，还是从饮料或掺入食物 中，均可被吸收。
16.2.2
盒物
铀天然地存在于全部食物中。食物加工的方式主要决定 了铀的最终含量。例如，冻豌豆比新鲜豌豆含制量高得多 ω 。 一 204 一
新鲜水果和蔬菜含的量从小予-1 0 毫克/公斤至 1 克/公斤，而 谷类食物和奶酷可能含 10~20 克/公斤 (2) 。牛奶含纳量相当 高，即1. 5 克/升 (2) 。瓶装饮水有时也可能含相近浓度的 纳 (2) 。但是难以估计从食物中每日铀的摄入量，因为在食物
中制的含量变动很大，而且很多人向食物中加盐。在西欧和 北美，估计从膳食中氧化铀的总摄入量为 5~20 克/日 (2~ 8 克制/日)，平均约为 10 克/日( 4 克制/日严"。由于健康 的原因，有些人需要特殊的低纳膳食，要求铀的摄入量小于 2 克/日 ω 。对于人造婴儿食物，已经制订了法规要求降低
含销量。
16.2.3
空气
从大气和车间空气中吸入的纳量远低于从膳食中吸收到 体内的量。
16.2.4
不同接触途径的相对意义
每个人摄入的销量相差很大。一般来说，主要来源是食 物。由于铀摄入后很容易被吸收，所以可以认为吸收量等于
摄入量。已经计算了从各类膳食和水源中铀的吸收量，并且 列于下表。成人的资料是根据已发表的从膳食中摄入纳的资 料 E235 而对于婴儿 (0~2 月)是根据估计值 250 毫克纳/日，
对于儿童(l ~5 岁〉是根据估计值为 2000 毫克/日来进行计 算的口气
16.2.-4.1
成人每周从水和膳食中纳的吸收量
qG nu
u 气 F
(a ) 水中纳的浓度
特殊的限纳膳食一-500 毫克纳/臼 每周纳的吸收量(毫克) 水
食物
总量
此值 水 /J总量
(%) 7 17 28 44
20毫克/升 50 毫克/升 100 毫克/升
280 700 1 400 2 800
3 500 3 500 3 500 3 500
3 780 4 200 4 900 6 300
200 毫克/升
(b) 水中纳的浓度
低纳膳食一一 2000 毫克纳/日 每周铀的吸收量(毫克) 水 比值 水/总量
食物
总量
(%) 2 5 9 17
20 毫克/升 50 毫克/升 100 毫克/升
280 700 1 400 2 800
14 000 14 000 14 000 14 000
14 280 14 700 15 400 16 800
200毫克/升
(c ) 水中纳的浓度
普通膳食一-5000 毫克纳/日 每周纳的吸收量(毫克) 水
此 f应 水/总量
食物
总量
(%)
20 毫克/刘 50 毫克/71.
280 700 1 400 2 800
35 000 35 000 35 000 35 000
35 280 35 700 36 400 37 800 2 4 7
100 毫克/升 200 毫克/升
-206~
(d) 水中铀的浓度
限制纳的强入量为 500 毫克/目 每周铀的吸收量(毫克) 水
食物
-一一一一←一
总量
比 f血 ←-水/总量
(%) 8 20 40 80
20 毫克/升
280 700 1 400 2 800
3 220 2 800 2 100 700
3 500 3 500 3 500 3 500
50毫克/升 100 毫克/升
200毫克/升
上述全部计算是根据每日水的摄入量为 2 升进行的。
16.2.4.2
儿童每周从水和膳食中纳的吸收量 ( a) 0~2 月龄婴儿一一从膳食中摄入锅
量为 250 毫克/日 每周铀的吸收量(毫克} 水中铀的浓度 水·
比值 水/总量
食物
总量
(%)
20毫克/升
140 350 700 1 400
1 750 I 750 1 750 1 750
1 890 2 100 2 450 3 150
7 17
50毫克/升 100 毫克/升
29 44
200 毫克/升
且很据水摄 λ 量为1. 0 升/日计算的。
( b)
1.-5 岁儿童-一从膳食中摄入纳量为
2000 毫克/目 每周纳的吸收量(毫克) 水中铀的浓度
水·
食物
总量
水/总量
地值
<%) ，唱，，。。
20 毫克/升 50 毫克/升 100 毫克/升 200 毫克/升
210 525 1 050 2 100
14 14 14 14
000 000 000 000
14 14 15 16
210 525 050 100
1
·棋锯水的慑 λ 量为1. 5 升/日计算的。
16.3
代谢
由于铀的生理学作用及其对机体的重要性，因此对的的 代谢已经进行了广泛的研究 (3) 。在此仅就纳代谢的一般问题
加以论述以供参考。纳是人血浆和细胞外液中最丰富的阳离
子，存在于骨、细胞和多数组织中。在内分泌、心血管和自 主调节机制的作用下 F 肾脏能精细地维持细胞外液中铀的浓 度。因此细胞外液中铀的总量决定了细胞外液的容量 (3) 。
铀的平衡是通过复杂而相互关联的系统，包括神经系统 和内分泌系统来控制的(1)。血浆铀浓度升高能剌激下丘脑
中心的渗透压感受器，无论体液容量如何，均会导致渴的感 觉 Elh 在炎热的气候和重体力劳动时，因出汗而丢失大量的 制，则需增加铀的摄入以补充丢失(1)。
铀的摄入并不是由生理学机制控制的。在食物中含量的 90% 以上都可被吸收 ω 。正常情况下，摄入量主要取决于膳
食。氯化铀的最低需要量约为 120 毫克/日〈约为 50 毫克 制) (2) 。
16.4 18.4.1 急性作用
对健康的影响
一般而言，铀盐不是急性毒性物质，因为发育成熟的肾
脏能有效地排泄锅。飞过量摄入氯化铀能引起呕吐和铀的大 量排出。急性作用可导致瘟孪、肌肉抽搞和僵化以及脑和肺 71< ]1中(1)。
铀对婴儿的作用与成人不同 F 因为婴儿肾脏发育尚未成
熟 ω 。据报道，当意外过量摄入氯化铀的情况下会出现急性 一 208 一
作用和死亡 (2) 。
过量摄入锅能引起慢性充血性心力;在竭病人的严重恶 化。巳有报道由于饮水中含高浓度制而引起有害作用 ω。
16.4.2
多铀血症
患严重胃肠道感染的婴儿，可能因体液丢失而引起脱水 以及血浆中铀含量升高(多纳血症)，在这种情况下常造成永 久性神经系统损伤 (2) 。
仅有有限而间接的证据表明，增加铀的摄入量是引起"婴 儿突然死亡"的原因之一，然而尚不能定论 (2) 。为了婴儿和
儿童的健康，应尽可能降低铀的总摄入量 (2) 。
已经表明，用牛奶加到固体食物中进行现代的婴儿喂养 是造成多铀血症的原因之 _(2) 。当把含高浓度铀的饮水掺入 该食物中时情况会更坏 E23 ，因为在牛奶中铀的浓度约为人奶
的三倍 (2) 。而婴儿肾脏发育尚未成熟，不能象成人那样有效
地维持渗透压 F 因此需要采取预防措施，例如制定法规，要 求降低婴儿食物中铀的浓度。
16.4.3
铀的摄入量增高与西血 E
对于摄入高制能影响高血压发病率的假说已有很多争 论，但是这一假说在科学上是具有说服力的 ω 。
16.4.3.1
从动物实验获得的证据
已经明确证实，接受高氯化铀饲料的不同种属动物发生 高血压 (2) 。尽管从动物实验外推到人还存在一些问题，但是
动物实验资料的一致性表明了这些实验的有效性。
16.4.3.2
从志愿者试验获得的证据
尚无结论性的证据表明，志愿者血压升高与高盐膳食有 一 209-
关。但是，在这些短期试验中有一些可疑现象说明可能与高
血压有关。西方社会的多数人从婴幼儿时起就摄入高盐膳食， 但是直至四十年代持续性高血压才成为常见病 ω 。
16.4.3.3 从流行病学调查获得的证据 ( a )对高铀膳食人群的研究 某些非西方化的人群组和西方人群之间有特别明显的不
同，前者为低纳膳食，高血压发病率很低p 而且血压不随年 龄的增长而升高 (2) 。虽然人们试图将其结论为因果关系，但
是两组人群之间的很多其他不同也可能解释这种差异 (2) 。尽
管如此，这些结果和其他研究之间的明显一致性 ω 进一步支
持摄入高纳和高血压之间具有直接的关系。
( b )对通过饮水摄入铀的研究 美国和荷兰现已完成的流行病学研究证实，居住在饮水 中铀浓度为中等水平(1 28........161 毫克/升〉区域学龄儿童(特
别是女孩〉的血压比居住在饮水中含纳量低 区域学龄儿童的血压稍高
(28 毫克/升〉
(3........5 毫米京柱) (4-6) 。苏联对
16~60 岁人群的类似调查也证实了饮水中铀的含量和血压 之间者类似的关系(7)。
在美国进行了一项研究 E町，把饮水中含高纳(1 08 毫克/
开〉的同一个居民区儿童(1 0........11 岁〉分为三组，以收缩压 为指标进行比较。其中两组给以含锅量为 108 毫克/升的瓶装
饮水，而另一组给以含低纳 (8 毫克/升〉的瓶装饮水。女孩 的结果与美国的其他两个研究报告一致，饮用低纳瓶装饮水 女孩的血压低于其他两组。而男孩的这种血压差异，在最初 6 周获得了与女孩一致的结果，但是在整个研究工作的其余
时间是不一致的。
一 210 一
-、.
.
16.4.4
饮水中铀的含量与其他接病的关系
虽然高血压和某些疾病(例如冠心病〉有关，但是由于 个体敏感性的遗传学差异，一些矿物质(锦和钙)可能具有
保护作用以及实验方法学方面的问题，因此难以对这种相关 进行定量。一般而言，饮水中制含量仅占膳食中含纳量的很
小一部分，故目前还不能对饮水中纳含量的重要性以及与疾 病的可能联系作出肯定的结论。 1978 年世界卫生组织的一个工作组综述了从各种来源摄 入铀的相对比例 ω 。该工作组的建议之一是在饮水中铀含量 超过 20 毫克/升的区域，公共卫生当局应该、通知公众，因为
某些人(高血压和充血性心力衰竭的病人)需要限制从膳食 中摄入铀的总量。鉴于是否需采取任何措施取决于当地情况
和政策，在本"准则"中未根据健康的考虑来推荐铀的特殊 建议值(见第五章 11.4 节，根据昧阔的建议值)。 (王有森译〉
参考文献 1. Guidelines for Canadian drinking water quality ,
1978. Quebec , Ministry of Supply and Services , 1980 (supporting documentation). 2. Sodium. chlorides and conductivity in drinkin!]
water. Copenhagen , WHO Regional Office for Europe. 1979 (EURO reports and studies. No. 2). 3. National Research Counci l. Drinking water and
health. Washington , DC , National Academy Sciences , 1977.
(,[
4. Calabrese , E. J. & T lI thill , R. W. Elevated blood
'-211-
pressure and high sodium levels in the public drinking water. Archives o{ environmenta/ health , 35:200 (1 977). 5. TuthiII , R. W. & Calabrese , E. 1. Elevated sodium levels in the public drinking water as a contributor to elevated blood pressure levels in the community.
Archives o{ environmenta/ hea/th , 37: 197
(1 979).
6. Tuth iIl, R. W. & Calabrese , E. 1. Drinking water sodium and blood pressure in children: a second look. American journal o{ public health , 71 :722729 (1 98 1). 7. Fatula , M. 1. The frequency of arterial hypertension among persons using water with an elevated sodium chloride content. Sovetskaja medicina , 30:123 (1 967).
一 212-
第四章与健康有关 的有机成分
1. 氯代皖怪 氯代院怪在化学工业中的主要用途之一是用作生产其他
有机氧化合物的中间体，因此，它可能会大量生产并在原水 和出厂水中被发现。根据现有资料，在已知的所有商品氯乙
皖中，仅 1 ， 2-二氯乙烧被明确证明具有致癌的危害。
1. 1 1.1. 1 概述
四氯化碳
四氯化碳 (CC1 4 ) 是一种卤代皖怪，在工业和化学方面 具有广泛的用途。在室温下，它是一种很重、无色的液体，
密度为 1594 克/升 F 相对无极性，可与乙醇、丙嗣和大多数 有机榕剂混合，在水中溶解度为 800 毫克/升 (25'(;) 。在美 国 11 个工厂生产量约为 423000 公吨 (932.7 X 10 6 磅严。主
要用于生产作为气洛胶推进剂的碳氟化合物
(1 973 年占
95%)0 1977 年 2 月由于渗漏事故造成 63.6 公吨的 CC1 4 排 入俄亥俄河，致使地面水浓度高达 340 微克/升 ω。也常常 在污染的地下水中检出 C Cl.，一般存在于原水中。据报道， 在出厂的饮用水中被度可达 2---3 微克/升。 CC1 4 的水解是从水中去除的一种方式，但与蒸发相比
并不明显。 • Jobn. , R ，四氯化碳对空气污染的评价。为美国环保局准备的文件. McLean , VA. , Mitre Corp. , 1976. 一 214 一
1. 1. 2 1.1. 2.1
接触途径 水
美国环保局进行了全国性的有机物调查 (2) ，发现至少在
10% 的饮用水中 CCl. 的浓度小于 2--3 微克/升。在新奥
尔良，人的血浆和饮用水中均发现 CCl. o C Cl.有时是饮 用水消毒剂氯的污染物，而不是在饮用水氯化过程中产生的。
1.1. 2.2
食物
在很多种食物中已检出 CCl.，其含量范围为 0.1'-"20 微克/公斤。 McConnell 等 (3) 总结了特别易受 CCl. 污染
的不同食物品种，他们注意到没有证据表明 C Cl.在食物 链的较高营养水平有明显的生物蓄积。食物受 C Cl.污染的 主要途径是由于用作熏烟剂所致。 空气
1.1. 2.3
已在大气中广泛地测定了 CCl. J 因而对其分布较为了 解。大气中 CCl. 的存在主要是由于它是一种挥发性化合 物，它的主要来源在于人类的活动 (4-7)。在大气中的分布接
近于均匀状态。虽然已有报告指出在城市空气中曾发现高浓 度的 C Cl• J 但是一般而言，近似于在大陆空气团中发现的
本底水平 0.00078--0.00091 毫克/立方米。
1.1. 3 1.1. 3.1
代谢 吸收
CCl. 易于经肺吸收 F 通过胃肠道吸收较慢，但是可完
全被吸收 E83 。 CCL 也可穿透皮肤进入机体。当 CCl. 与脂 肪 (8) 和乙醇 (8帽 10) 一起摄入时，其吸收速度和数量增加。
RobbinsC ll)在研究 CCl. 经狗的胃肠道吸收时发现，大量 一 215 一
的 CCl~ 是经小肠吸收的，经结肠吸收较少，而几乎不经胃 吸收。 分布
1.1. 3.2
Nielsen 和 Larsen (8) 发现在动物宰丸的脂肪组织、 肝、脑、骨髓和肾脏中 CCl~ 浓度高。 Robbins (ll)研究了
经口给狗 CCl~ 后的分布，发现在骨髓中 CC1 4 浓度最高，
在肝、膜手口牌中的含量是骨髓含量的五分之一。 Recknagel 和 Litteria (1 2) 的研究表明， CCl ，在器官中的分布随染毒
途径、浓度及接触时间的不同而异。 McLean 等 (3) 在含有 较高浓度核糖体的所有细胞成分中均发现 CC1 4 。
1.1. 3.3
生物转化
当给哺乳动物 CCI ，时，被代谢的很少，大部分通过肺
脏排出。代谢产物包括氯仿、六氯乙院和二氧化碳。研究工 作已证实，这些代谢产物在 CC1 4 整个毒性中起着重要的作 用 (14) 。
1.1. 4
对健康的影响
CCI ，对人的毒性常常不被认识 (15) 。经口、皮肤和吸入
接触均可引起急性和亚急性毒性，从而对皮肤、循环、呼吸、 血液以及肾、肝、眼和膜脏功能产生有害作用。
在很多急性中毒的情况下，病人在数天内出现肝损伤的 症状，表现为黄瘟，肝脏变大和变脆。当肝脏出现损伤或有 时未见损伤时，可能观察到肾损伤，而且可能是主要的临床 表现，这常常是早期死亡的原因(1 6) ，但是一般而言，在 CC1 4 引起的毒性反应中肝损伤比肾损伤更常见。还注意到有血液
指标、视力和膜脏的改变。慢性 CCl，中毒的临床表现远不
如急性中毒那样典型。 CCI‘中毒死亡病例的病理改变通常 一 216-
局限于肝和肾的病变。 关于 CCI. 的致突变性资料极少。 Kraemer 等(1 7)发
现 CCI. 在鼠伤寒沙门氏菌和埃希氏大肠杆菌回变试验中无 致突变性，但是在 Ames 试验中卤代怪通常为阴性。 现有资料足以得出结论: CCI. 是实验动物的致癌物(1 8'" 24)。对人接触的允许浓度标准是基于美国国立癌症研究 所 (2 4)对三氯乙烯的研究工作，在这一工作中 CCI. 用作阳
性对照，它对 B6C3-FI 小鼠是致癌的。虽然也进行了其它 的研究工作，但是未得到足够的剂量-反应关系的资料或实
验周期太短不能用于估计危害性。由于已知肝脏毒物(如 CCI.) 对该品系小鼠引起肝肿瘤的机制不清楚，所以用无阔 值模式外推是否合适尚不能肯定。而对非遗传毒物诱发的癌
症又没有其他的外推模式。所以就必须采用保守的方法确定 这一饮用水中常见污染物的限值。因此，使用线性多阶外推 模式来推导建议的试验性限值为 3 微克/升。这一限值是基于
假定每日饮水量为 2 升，在终生接触的情况下每 10 万人口 中增加的癌症少于 1 例而确定的。
参考文献
1. Amnican Chemical Society. CCI, spill causes along Ohio River. Chemical and engineering news , 55:7 (1 977) .
2. Ambient water quality criteria for carbon tetrachloride. Washington , DC , US Environmental Protection Agency , 1980 (EPA 440/5-80-026). 3. McConnell , G. et a 1. Chlorinated hydrocarbons in the environment. Endeavour , 34:13 (1975). 4. Altschuller , A. P. Average tropospheric concentration -217 一
of carbon tetrachloride based on industri r.l production , usage and emission 串 • Envi 俨 onmelltal
science
technology , 10 ' 596 (1 976). 5. Love lock , J. E. et a 1. Haloger ated hydrocar bOllS in and over the Atlantic. Nature. 247:194 (1 974). 6. Wilkniss , P. E. et a I. Atmospheric trace gases in the southern hemisphere. Nature , 245:45 (l 973). 7. Singh , H. B. et a I. Atmospheric carbon tetrachloridea Another man-made pollutant. Science , 192:1231 (1 976) .
8. Nielsen , V. K. & Larsen ,1. Acute renal failure due to car bon tetrachloride poisoning. Acta med i ca
Scandinavica , 178:363 (1965). 9. F olland , D. S. et al. Car bon tetr achloride toxici ty potentiated by isopropyl alcoho l. Investigation of an industrial outbreak. J ournal
0/
the Amer i can tetr 且 chloride
M edical Association , 23:1853
(1 976).
10. Moon , H. D. Pathology of fatal carbon
poisoning with special reference to histogenesis of the hepatic and renal lesions. American journal
0/
pathology , 26:1041
(1 950).
11. Robbins , B. H. The absorption , distribution , and excretion of carbon tetrachloride in dogs under various conditions. Journal
0/
pharmacology , 37 203 '
0929) . 12. Recknagel , R. O. & Litteria , M. Biochemical changes in carbon tetrachloride fatty livera Concentration of carbon tetrachloride in liver and blood. American
journal 0/ pathology , 36:521 (1 960). 13. McLean , A. S. M. et a I. Cellular necrosis in the liver -218 一
‘
induced and modified by drugs and other agents.
International review 127 (1 965).
0/
experimental pathology , 4:
14. Gordis , E. Li pid metabolites of carbon tetrachloride.
J ournal 0/ c /i ni cal investigation , 48: 203
(1 969).
15. Von Oettingen , W. F. The halogenated hydrocarhons of industrial and toxicological importance. In: Browning , E. , ed. Elsevier monographs on toxic
agents. New York , EIsevier Puhlishing Co. , 1964. 16. Von Oettingen , W. F. T he halogenated ali phatic ,
ole/inic , cyclic , aromatic , and aliphatic-arornatic hydrocarbons including the halogenated insecticides , their toxicity and potential dangers. Washington , DC , Department of Health , Education & Welfare , 1955. 17. Kraemer , M. et a l. S. typhimurium and E. coli to detect chemical mutagens. Naunyn-Schmiedebergs
archives
0/ pharmacology , 284: 46R hydroca 俨 bons.
(Ahstract).
18. Some halogenated
Lyons , International
Agency for Research on Cancer , 1979 (1 ARC Monographs on the evaluation of carcinogenic risk of chemicals to humans , vo l. 20). 19. Edwards ,1. Hepatomas in mice induced with carhon tetrachloride. Journal
0/
the National Cancer
1 nstitute , 2: 197 (1 94 1). 20. F. dwards , J. & Dalton , A. Induction of cirrhosis of the liver and hepatomas in mice with carhon tetrachloride. J ournal
0/
the N ati onal Cancer
Institute , 3:19 (1 942). 21. Edwards ,1. et a l. Induction of the carbon tetrachloride -219 •
hepatoma in strain L mice. Journal
0/
the National
Cancer 1 nstitute , 2 1 297 (1 942). 22. Eschenbrenner , A. B.
& Mi Il er , E. Studies on
hepatomas-size and spacing of multiple doses in the induction of carhon tetrachloride hepatomas. J ourna/
the N ati ona/ Cancer 1 nstitute ， 中 385 (1 943). 23. Eschenbrenner , A. B. & Mi Il er , E. Liver necrosis and the induction of carbon tetrachloride hepatomas in strain A mice. Journa/
0/
0/
the National Cancer
Institute , 6:325 (1 946). 24. National Cancer Institute. Carcinogenesis bioassay Health , Education & Welfare , 1976 (CAS No. 7901-6 , NCI-CG-TR-2).
0/
trichloroethylene. Washington , DC , US Department of
1. 2 1. 2.1 概述
1, 2-二氯乙烧
1 ， 2一二氯乙皖 (CH z CI-CH 2 CD
是一种液体，
比重
1. 25 ，嗅阔为 2 毫克/升(1)。广泛用作很多有机化学物质的
溶剂、化学合成的中间体和杀虫剂。在美国， 1976 年的产量 是 3.63X10 6 吨 (
8x
10 u 磅 ) (2) 。
1.2.2
接触途径 饮用水
1. 2.2.1
由于 1 ， 2一二氧乙昆在工业上使用的结果，它成为工业废 水的一种成分，并在美国的原水和自来水中被检出。在美国
的 28 个城市的水中已发现其浓度达 6 微克/升。 3。 -220-
1. 2.2.2
食物
1 ， 2-二氯乙;院摄入量的 0.4% 来自于食用水生物，其平
均的生物富集能力为1. 2 倍。摄入量的其余 99.6% 来自于饮 用水 (4) 。
1. 2.2.3
职业接触
据美国国立职业安全和卫生研究所 ω估计，有 450 万工
人经吸入和皮肤接触 1 ， 2-二氯乙烧。
1. 2.2.4
空气
在城市空气中可检出 1 ， 2- 二氯乙皖，其浓度为 0.04--
38 微克/立方米 (5) 。因生产、贮存和分配的结果，据 1974 年
美国的计算， 1 ， 2- 二氯乙炕向大气中的排放量约为 74 公斤，即大概是总产量的 1. 8% 。
X
10 6
1.2.3
代谢
几乎没在关于 1 ， 2-二氯乙院代谢的资料。巳知它易溶于 脑的脂类中，这一特性促进了它对神经系统的影响(1)。
1.2.4
对健康的影响
关于 1 ， 2-二氯乙炕毒性的资料主要与职业性的吸入接
触者关F 它是一种麻醉剂，引起肝、肾和心血管系统的损 害 (7) 。当空气中 1 ， 2-二氯乙烧的浓度低于 4 毫克/立方米 时，中毒症状不明显(1)。
基于现有资料，美国国立职业安全和卫生研究所建议职 业接触 1 ， 2-二氯乙脆不应超过 20 毫克/立方米，这是根据
10 小时一个工作日和 40 小时一个工作周的时间加权平均值 而确定的。高峰浓度不应超过 60 毫克/立方米，这是按 15 分钟样品确定的。美国职业安全和卫生科学院建议的接触标
准是 200 毫克/立方米，而苏联的职业接触标准为 10 毫克/ 立方米。 1 ， 2-二氯乙烧对大自鼠的径口半数致死量为 1120 士 142 毫克/公斤体重 (7)。
流行病学研究尚未揭示接触 1 ， 2-二氯乙烧和癌症之间
的关系，但是这种化合物在动物实验中是致癌的，引起雄性
大鼠的前胃鳞状细胞癌和循环系统血管肉瘤、雌性大鼠和小 鼠的乳腺腺癌和雌性小鼠子宫内膜瘤均具有统计学意义的明 显增加 (6 ， 7)。使用适当的美国国立癌症研究所生物试验资料，
通过线性多阶模式计算得到 1 ， 2一二氯乙院在水中的标准。 1 ， 2-二氯乙烧是一种已知的致突变物。在 Ames 沙门
氏菌 TA1530 和 TA1535 试验中以及埃希氏大肠菌 DNA 多聚酶缺乏的试验系统中是致突变的 (8)。还能引起果蝇体细
胞突变频率的极明显增加 (9)。用 1 ， 2-二氯乙烧处理种子可 引起 8 种豌豆的形态学和叶绿素的突变 (10)。
氯乙睦是假设的 1 ， 2-二氯乙妮的代谢物，对鼠伤寒沙门 氏菌 TA100 是致突变的 (11)。
对 1 ， 2-二氯乙饶的建议值是基于在 78 周内经口结雄性 Osborne-Mendel 大鼠 1 ， 2-二氯乙烧引起循环系统血管 肉瘤的资料 (10) 。为使终生患癌症的危险性低于十万分之一，
经计算得到在水中 1 ， 2-二氯乙饶的浓度为 9.4 微克/升， 约为 10 微克/升。 参考文献
即
1. Zoeteman , B. C. 1. Sensory assessment and chemícal composition 01 drinking-water. Leidschendam , Netherlands , Institute of Water Supply , 1978.
2. National Institute for Occupational Safety and Health. Ethylene dichloride 口 ， 2-dichloroethane). Washington , DC , Department of Health (NIOSH) publication No. 78149). 3. Symons. 1. M. et a l. National organics reconnaissance 8urvey for halogenated organics. J ournal Welfare , 1978 (Current Intelligence Bulletin 25
, Education and
01
the (1 975>.
American Water Works Association , 67 1 634
4. Ambient water quality criteria lor chlorinated ethanes. Washington , DC. Environmental Protection Agency , 1980 (440/5-80-029). 5. Okuno. T. et a l. (Gas chromatography of chlorinated hydrocarbons in urban air.) H yogo-ken kogai
kenkyusho kenkyu hokoku , 6 1} -6 (Chemical abstracts , 87.72564 f) (1 974).
6. National Cancer Institute. Bioassay
01 1, 2-
dichloroethane /or possible carcinogenicity. Washington , DC , US Department of Health , Education
& Welfare , 1978 ((NIH) 78-1305).
7. Some halogenated hydrocarbons. Lyon , International Agency for Research on Cancer. 1979 (l ARC Monographs on the evaluation of the carcinogenic risk of chemicals to humllns. vo l. 20). 8. Brem , H. et a l. The mutagenicity and DNA-modifying effect of haloalkanes. Cancer research , 34 1 2576 (1 974) .
9. Nylander P. O. et a l. Mutagenic effects of petrol in
Drosophila melanogaster. 1. Effect~ of benzene and of 1. 2-dichloroethane. M ufati on research , 57:163 (1978) • 一 223-
10. Kirichek , Y. F. Effect of 1 , 2-dichloroethane on mutations in peas. Usþehi himii mutageneza se. , 232 (1 974) .
11. Some monomers , plastics and synthetic elastomers ,
and acrolein. Lyon , International Agency for Research on Cancer , 1979 (I ARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo l. 19).
-224 一
2.
氯代乙烯 a
该组化合物广泛应用于多种工业加工过程，用作溶剂、
软化剂、油漆稀释剂、干洗液和中间体等，因此常在原水和
处理的饮用水中发现。已知在地下水中的浓度为数毫克/升， 由于氯代乙烯挥发性强，往往从地面水中蒸发至大气中，故 在地面水中的浓度一般较低。
在该组化合物中最令人担心的是那些对实验动物具有致 癌活性的化合物，其中包括众所周知的人类致癌物氯乙烯。 但是，由于氯乙烯在水中的出现似乎主要与使用聚合不好的
氯乙烯水管有关，所以通过控制产品规格来解决该问题比确 定建议值更合适飞
2.1 2. 1. 1 概述
氯乙烯
氯乙烯主要用于生产聚氯乙烯树脂，而后者则是世界上
最广泛使用的塑料。其次的用途〈小于总产量的 5%) 是生 产甲基氯仿的一种中间体以及在生产氯亚乙烯-氯乙烯聚合 体(广泛用于食品包装和涂抖)时与氯亚乙烯共聚用的单体。
以前，氯乙烯曾用作气溶胶推进剂和制冷剂，但是似乎现在 a 以前采用的名称为 Chlorin&ted ethylen". 。
b 美国国立卫生基金会标准 No ， 14 ，塑料管成分和有关物质，
1980 年 12 月
1圭订，允许每公斤饮水管含有 10毫克氯乙烯单体。 一 225 一
已停止了在这方面的应用 ω 。
聚氯乙烯的最大用途是生产水管，其他的重要用途是在 铺地板、消费用品、电器和运输方面的应用(1)。 氯乙烯具有挥发性，在大多数的实验室和生态条件下易 于从液相进入气相，因此，在化学工厂和生产橡浆的工厂排 出的废水中(1)以及在饮用水(由于氯乙烯从配水管网的聚氯
乙烯管中逸出) (2) 中检出的氯乙烯浓度很低。已经有许多的
产品标准，限定游离氯乙烯单体 (VCM) 的含量，从而控制 聚氯乙烯水管的质量。如果使用这种质量的水管，则在饮水 中氯乙烯单体的浓度就可能低于用计算其他致癌有机物时采
用的线性多阶外推模式而获得的限值〈该值为 20 微克/升， 它是基于可接受的危害性为一生中每 10 万人中增加的癌症 少于 1 例〉。
2. 1. 2 2. 1. 2.1
接触途径 水
在美国七个区域废水样品中氯乙烯的浓度(与聚氯乙烯 氯乙烯生产工厂有关〉范围为 0.05-20 毫克/升 (3)。出厂自 来水中检出的氯乙烯最高浓度为 10 微克/升(4)。而在五个城
市的调查中，从聚氯乙烯水管构成的配水系统中采集的饮水 中检出的氯乙烯浓度达1. 4 微克/升 (2) 食物
2. 1. 2.2
由于氯乙烯能从聚氯乙烯包装材料中进入食品，所以可 能食入小量的氯乙烯。美国食品和药物管理局的研究表明， 在聚氯乙烯容器内的含醇饮料中氯乙烯的浓度达 20 毫克/公 斤 (5) 。已发现，在聚氯乙烯容器内包装和贮存的食用油、黄
油和人造黄油中氧乙烯的浓度达 14.8 毫克/公斤(1)。现在， 一 226 一
很多国家在聚氯乙烯包装材料中限制氯乙烯单体的含量，并 且禁止其用于含醇制品或食用油。
2. 1. 2.3
空气
氯乙烯在常温常压下是一种气体，存在于生产氯乙烯或
聚氯乙烯工厂的附近。在生产氯乙烯工广附近的空气中已检 出其浓度达 8.8 毫克/立方米(1)。以前，氯乙烯用作很多气 洛胶产品的推进剂(如农药、头发喷雾剂和除臭剂);无疑， 反复使用这些产品的人会接触中等浓度的氯乙烯。在新汽车 内的空气中己检出氯乙烯，其浓度为 1""'3 毫克/立方米 ω 。
2. 1. 3 2. 1. 3.1 2. 1. 3.2
代谢 吸收
经口给予 (6 ， 7) 或吸入氯乙烯 (7)后易于被吸收。
分布
氯乙烯在大鼠体内分布的研究表明，在肝、肾和膊中检 出的浓度最高 (6 ，的。
2. 1. 3.3
生物转化
微粒体混合功能氧化酶(主要经 P-450 系统〉将氯乙 烯代谢为氧化氯乙烯，后者可自然重排为氯乙睦。而氯乙醒 代谢的主要途径是被氧化为氯乙酸 z 它可能被直接排出或与 谷眈甘肤结合，经进一步的酶降解作用后排出。氯乙睦还经 许多其他途径代谢 (8) 。
2. 1. 3.4
排泄
已经报道了吸入和静脉注射氯乙烯后，其排泄的动力学 和生物半减期 (7) 。经胃肠道给大鼠剂量为 250 微克/公斤体
重的氯乙烯后，在 24 小时内排出 96% 以上 (3.7% 以氯乙烯 的形式呼出， 12.6% 以二氧化碳的形式呼出， 7 1. 5% 从尿中 一 227 一
排出，
2.8% 从粪中排出) (1 0) 。
2. 1. 4
对健康的影响
2. 1. 4.1
急性和亚急性毒性
急性接触氯乙烯的主要反应是中枢神经系统的抑制。尸 检时的病理学所见包括肺充血、水肿及肝、肾充血 (1 1)。
2. 1. 4.2
致癌性
已对动物的致癌实验和人的流行病学观察进行了研究和
评述。已经证实，摄入或吸入氯乙烯对大鼠、小鼠、地鼠和 家兔具有致癌作用 E 引起几个部位的肿瘤，包括肝血管肉瘤。
氯乙烯引起人肝脏的血管肉瘤以及脑、肺和血液淋巴生戒系 统的肿瘤(1)。国际肿瘤研究机构认为这些证据足以证实接触 氯乙烯和引起肿瘤之间的因果关系 (12) 。
2. 1. 4.3
致突变性
已经评述了氯乙烯及其数种代谢物的致突变性(1)。在很
多生物系统中它是致突变的，这些生物系统包括鼠伤寒沙门 氏菌、埃希氏大肠菌 K12 生物营养缺陷株、数种酵母、果蝇
的生磕细胞和中国地鼠 V79 细胞。而致突变作用取决于代 谢活化。
2. 1. 4.4
致畸性
娃振期接触氯乙烯的小鼠和大鼠出现骨悟异常(1 3)。
参考文献
1. Scme monorners , plastics and synthetic elastomers ,
and acrolein. Lyon , International Agency for Research on Cancer , 1979 flARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo l. -228 一
19) . 2. Dressrnan , R. C. & McFarren , E. F. Deterrnination of vinyl chloride rnigration frorn polyvinyl chloride pipe into water. Journal
0/ the American Water Works (1 978).
Association , 70:29
3. Preliminary assessment
0/
the environmental
problems associated with vinyl ch/ oride and polyvinyl chloride. Springfield , VA , US Environrnental Protection Agency , 1974 (E P A 560/4-
74-00 1) . 4. Safe Drinking Water Cornrnittee. Drinking water and
health. Washington , DC , National Acadern y of Sciences , 1977 , p. 794. 5. Anon. FDA to propose ban on use of PVC for liquor use. Food chemical news , 14 MaYI 3-4 (1 973). 6. Watanabe , P. G. et a l. Fate of (UC) vinyl chloride after single dose adrninistration in rats. Toxicology
and applied pharmacology , 36:339
(1 976).
7. Withey , 1. R. Pharrnacodynarnics and uptake of vinyl chloride rnonorner adrninistered by various routes to rats. Journal
0/
toxicology and environmental 口， 2_ 14 CJ
health , 1: 381 (1 976). 8. Bolt , H. M. et a l. Disposition of chloride in the rat. Archives (1 976) .
vinyl
0/
toxicology , 35:153
9. Plugge , H. & Safe , S. Vinyl chloride rnetabolism. A review , Chemosphe 俨 e ， 6: 309 (1 977). 10. Green , T. & Hathaway , D. E. The biological fate in rats of vinyl chloride in relation to its carcinogenieity.
Chemico-biological interactions , 11 :545
(1 975).
-229-
1 1. Patty , F. A. , ed. /ndustrial hygiene and toxicDlogy. Vo l. II , New York , Interscience , 1963.
12. Chemicals and industrial processes associated with cancer in humans. Lyon , International Agency for Researcb on Cancer , 1979 (l ARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , supp l. 1). 13. John , J. A. et a l. The effects of maternally-inhaled vinyl chloride on embryonal and foetal development in mice , rats and rabbits. Toxicology and applied
pharmacolo9Y' 39: 497 (197 7).
--230-
2.2 2.2.1 概述
1， 1-二氯乙烯
在二氯乙烯的三种异构体中，
1 ， 1-二氧乙烯(1， 1-
DCE) 是化学工业中最广泛使用的。它在合成甲基氯仿和
生产聚氯亚乙烯聚合物中是一种中间体。聚氯亚乙烯聚合物 在包装工业中用作载体的涂料，而 Saran (含聚合物的 1 ，卜
二氯乙烯〉广泛用于食品包装工业。 1 ， 1-二氯乙烯在水中的 溶解度为 2500 毫克/升，已报道其在辛醇/水中的分配系数 为 5.37 ，表明不会在动物体内明显蓄积。
2.2.2 2.2.2.1
接触途径 水
美国环保局的国家有机物监测调查(1)报道了在饮水中检 出了 1 ， 1-二氯乙烯，但未予定量。它的来源之一可能是 1 ， 1 ， 1-
三氯乙烧的降解，后者有时在饮水中被检出的浓度约为 1 微 克/升。， 3)。在某些欧洲的地下水中已发现二氯乙烯。
2.2.2.2
食物
1 ， 1-二氯乙烯聚合物的食品包装材料巳广泛使用，但遗
憾的是，没有资料表明未反应的单体逸入包装食品中的程 度。看来，人通过膳食在其他方面的接触很少。
2.2.2.3
空气
通过吸入途径的接触主要是职业性接触。阔限值为 40
毫克/立方米空气，相当于在使用和生产二氯乙烯工厂中的 工人接触 280 毫克/天 E43.
-231-
2.2.3 2.2.3.1
代谢 吸收
根据对有关化合物(如三氯乙烯)的研究，估计 1 ， 1-二 氯乙烯摄入量的 100% 最终会被全身吸收吼的。
2.2.3.2
分布
对 1 ， 1-二氯乙烯在大鼠体内分布的研究 (7)发现，在肾中
浓度最高，其次是肝、牌、心和脑。在血液中的浓度与组织 中的浓度密切相关。对亚细胞水平分布的研究资料表明， 1 ， 1-二氯乙烯代谢产物与大分子物质紧密结合，并且也与脂
类结合。
2.2.3.3
生物转化
Liebman 和 ûrtiz(S)检出了由 1 ，1-二氯乙烯生成的
氯乙酸。看来，氯乙烯通过环氧化物中间体而被代谢。这种 中间体是有活性的，并且可能与组织中的大分子物质形成共 价结合 (9) 。在完整的实验动物体内，从全身吸收的大部分 1 ， 1-二氯乙烯被代谢转化。但是关于 1 ， 1-二氯乙烯代谢产 物与其毒性之间的关系还不很清楚。
2.2.4 2.2.4.1
对健康的影响
急性、亚慢性和慢性毒性
1 ， 1-二氯乙烯和其他氯代乙烯类一样，具有麻醉作用。 Prendergast C1 O )报道了接触含 1 ， 1一二氯乙烯空气的大鼠
和豚鼠发生肾和肝损伤。在接触浓度和总接触时间相似的情 况下，断续接触和连续接触的结果不同，在较低浓度下，连 续接触比断续接触引起的死亡率高。经口给予一次剂量为
200~400 毫克/公斤的 1 ， 1-二氯乙烯对肝脏的酶活性具在明
-232-
显的影响。仅有一篇已发表的流行病学研究调查了接触 1 ， 1二氯乙烯的工人(11)。在 138 名工人中，未见任何与接触 1 ， 1-
二氯乙烯有关的异常发现;
测定的工作场所的浓度范围为 9
~280 毫克/立方米(时间加权平均值〉。
Z.2.4.2
致突变性 K12(13)试验系统中是致突变
已表明， 1 ， 1-二氯乙烯在鼠伤寒沙门氏菌 TA1530 和
T A100 (12 ) 以及埃希氏大肠菌 的。 Henschler (9
) 及其助手指出，氯乙烯系列化合物的致
突变活性和推测的致癌活性与氯非对称性地取代各自的环氧 化物中间体有关。这种取代物与对称取代的环氧化物相比， 稳定性差，但活性强。在哺乳动物试验系统中尚未确证象在
细菌试验系统中所见到的使突变率增加。
2.2.4.3 2.2.4.4
致畸性
还未对二氯乙烯类的致畸作用进行评价。 致癌性 Maltoni 及具同事 C14 ， 15)报道了吸入 1 ，1一二氯乙烯的
实验结果。当浓度为 100 毫克/立方米时，在 300 只瑞士小
鼠中有 25 只发生肾腺癌，而对照组未发生。吸入浓度为 100 毫克/立方米的瑞士小鼠和吸入浓度为 600 毫克/立方米的 Sprague-Dawley 大鼠的乳腺腺癌发生率明显增加。 Lee 等 (16) 观察到，动物吸入浓度为 220 毫克/立方米的
1 ， 1-二氯乙烯， 4 小时/天， 5 天/周，共 7"'_'12 个月，结果
表明，肝血管肉瘤的发生率稍有增加。 在 Rampy 等 (7) 的实验中，使 Sprague-Dawley 大
鼠饮用含 1 ， 1-二氯乙烯浓度为 200 毫克/升的饮水 2 年，
并
使其吸入 1 ， 1一二氯乙烯的浓度为 100 和 300 毫克/立方米，
结果未见处理组动物肿瘤发生率增加的证据。但是，鉴于在 -233 一
Maltoni(15) 的研究中证实 Sprague-Dawley 大鼠不敏
感，所以不能认为根据这一资料可以改变对瑞士小鼠出现阳 性结果的解释。
有一些证据表明，通常大鼠对低分子量氯代怪的致癌作 用敏感(1 8) 。目前对接触氯亚乙烯工人的流行病学调查资料 还不足以作出评价 E182.
2.2.5 2.2.5.1
基准值 现行标准
美国的现行标准是针对吸入途径的职业接触而确定的， 美国政府工业卫生医师会议 ω 确定的车间空气中 1 ， 1- 工氯
乙烯的阔限值为 40 毫克/立方米。该值的规定意味着容许每 天接触 286 毫克 1 ，1-二氯乙烯。这一标准是基于上述 Pre ndergast 等 (10) 的工作而确定的。
2.2.5.2
致癌的危害限值
已表明， 1 ， 1-二.氯乙烯引起小鼠和大鼠的乳腺肿瘤以及 小鼠的肾腺癌。此外，它在 Ames 试验〈为致癌活性的定
性指标〉中具有致突变性。基于该资料，使用线性多阶外推
模式确定限值，通过计算该限值所造成的危害是 z 假设 70
公斤体重的人每天水的摄入量为 2 升，贝IJ 每 10 万人群中额外 增加的癌症病例少于一例。这个限值为 0.3 微克/升，该值 低于根据非致癌危害所确定的数值。 参考文献
1. Statement
0/ basis and purpose /or an amendment prima俨 yd 门inking
to the national interim regul ation 冒 一 234 一 on
water
a treatment technique /or synthetic
organics. Washington , DC , US Environmental Protection Agency , 1978.
2. P reliminar y assessment 0/ suspected carcinogens in drinking water. Washington , DC , US Envirollmental Protection Agency , 1975.
3. List 0/ organic compounds identi/ied in US drinking water. Cincinnati , OH , US Environmental Protection Agency , 1978.
4. T LVs-Threshold limit values /or chemical substances and physical agents in the workroom environment with intended changes /or 1976. Cincinnati , American Conferenc e. of Governmental Industrial Hygienists. 1976. 5. McKenna , M. J. et a J. The fate of (14C) vinylidene chloride following inhalation exposure and oral administration in the rat. Proceedings
0/ the
Society 01 Toxicology , 206 (1 977). 6. McKenna , M. J. et a l. Pharmacokinetics of vinylidene chloride in the rat. Environmental health
perspectives , 21 :99-106 hepatotoxicity
(1 977).
7. Jaeger , R. L. .et a l. 1 , I-dichloroethylene
,
proposed mechanism of action of
distribution alld binding of 14C radio-activity following inhalation exposure in rats. Environmental
health perspectives , 21: 113-120
(1 977).
8. Leibman , K. C. & Ortiz , E. Metabolism of halogenated ethylenes. Environmental health perspectives , 21: 91-98 (1 977). 9. Henschler , D. Metabolism and mutagenicity of halogenated olefins一 A
comparison of structure and
activity. Envi 俨 onmental health perspectives , 21: 61~64
(1 977).
10. Prendergast ,
J. A. et a 1. Effects on experimental
animals of long-term inhalation of trichloroethylene , carbon tetrachloride , 1 , 1, 1-trichloroethane , dichlorodifluoromethane , and 1 , 1-dichloroethylene.
Toxicology and applied pharmacology , (1 967) •
10:270~289
1 1. Ott , M. G. et a 1. A health study of employees exposed to v iny lidene chloride. J ourna/ 01 occupati onal
medicine , 18'735 (1 976). 12. Bartsch , H. et a 1. Tissue-mediated mutagenicity of vinylidene chloride and 2-chlorobutadiene in
Salmonella typhimurium. Nature , 255:641 (1 975). 13. Greim , H. et a 1. Mutagenicity in vUro and potential carcinogenicity of chlorinated ethylenes as a function of metabolic oxirane formation. Biochemica/
pharmac%gy , 24: 2013 (1 975). 14. Maltoni , C. et a 1. Carcinogenicity bioassays of vinylidene chloride. Research plan and early results.
Medicina de lavoro , 68:241 (1 977). 15. Maltoni , C. Recent findings on the carcinogenicity of chlorinated olefins. Environmental health
perspectives , 21: 1 (1 977). 16. Lee , C. C. et a 1. Inhalation toxicity of vinyl chloride and viny lidene chloride. Envi ronmental health
perspectives , 21: 25 (1 977). 17. Ram py , L. W. et a 1. Interim results of a two-year toxico logica 1 stud y in ra ts of v in y lidene chloride incorporated in the drinking water or administered by 。，" 。。
repeated inhalation. Environmental health
perspectives , 21'33
(1 977).
18. National Cancer Institute. Bioassay
0/
tetrachloroethylene lor possible carcinogenicity. Washington , DC , US Department of Health , Education and Welfare. 1977 (Technical Report Series No. 13. (NIH) 77-813).
19. Some monomers , plastics and synthetic e/astomers. and acrolein. Lyon , International Agency for Research on Cancer , 1979 (l ARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo I. 19).
-237 一
2.3 2.3.1 概述
三氯乙烯
三氯乙烯(1， 1 ， 2-三氯乙烯 J
TCE) 是一种清晰无色的
液体，其经验式为 C 2 HCI g 。在金属工业中它主要用作脱脂 溶剂，也用作家庭和工业的干洗剂、在食物中的提取谣剂以
及在某些短时间的外科手术时作为吸入麻醉剂 (1)。 在生产和使用时，三氯乙烯的挥发是其在环境中含有一
定浓度的主要来源。三氯乙烯在空气、食物和人体组织中已 被检出口 3。在河流、城市给水、海洋和水生物中的检测表明， 它已广泛分布在水环境中 α'"'.)。预计它不会在地面水中持续
存在，因为其具有挥发性，但是已经发现它是地下水的常见 污染物。
2.3.2 2.3.2.1
接触途径 水
美国国家有机物监测调查表明.
1976 年 3"""4 月在 112
个城市的饮水中有 4 个城市饮水中的三氯乙烯平均浓度为
11 微克/升， 1976 年 5"""7 月在 113 个城市中 28 个城市的 平均值为 21 微克/升， 1976 年 11 月 """1977 年 1 月在 113 个
城市中 19 个城市的平均值为1. 3 微克/升。水中的三氯乙烯 可能来源于直接的污染或因下雨来自于大气污染。)，也可能 在水加氯消毒时形成邸， 6) 。
2.3.2.2
食物
几乎没有关于食品中三氯乙烯的资料。在英国友现肉类
中的含量达 10 微克/公斤，水果、藏菜和饮料中为 5 微克/
公斤 (3) 。包装的茶叶含 60 微克/公斤。预计除在咖啡粉和速
洛咖啡以及在香料提取物中当三氯乙烯作为溶剂时含有三氯 乙烯以外，在其他食品中几乎不含有三氯乙烯。 空气
2.3.2.3
迄今，人对三氯乙烯的接触仅限于相当小量的职业人 群 (7) 。其他经吸入途径的接触是与使用含三氯乙烯的清洗剂
者关，但是这种接触的危害表现为急性作用。
2.3.3 2.3.3.1
代甜
吸收
三氯乙烯易于经所有的接触途径吸收，基于其理化性质 这是可以预见的阳。由于对三氯乙烯作为工业毒物和麻醉
剂的兴趣，所以有关人对三氯乙烯吸收的大部分资料是通过 吸入接触途径获得的。但是关于人类食入三氯乙烯后的吸收
问题尚未进行研究。在大鼠的实验中，经口给予三氯乙烯总 剂量的 72-'85% 和 10-20% 分别经呼出气和尿中排出，小于 0.5% 经粪便排出 (9) 。这表明至少食入量的 80% (可能更
多〉是经全身吸收的。
2.3.3.2
分布
三氯乙烯在体内的分布，正如基于理化性质所预见的那 样 (8)。在豚鼠的实验中观察到，在脂肪中浓度的大约 50% 在 卵巢中，约 25% 在其他组织中。 Laham(lO)证实在人体中
三氯乙烯可经胎盘扩散，胎儿和母体血中浓度之比波动于 0.52- 1. 90。
2.3.3.3
生物转化
三氯乙烯的代谢似乎是引起其长期有害作用的关键。从 定性的角度而言，其在不同种属中的代谢类似 (11-13) 。在原
-239-
中测定的主要代谢产物为三氯乙睦、三氯乙醇、三氯乙酸和 三氯乙醇的结合产物〈葡糖醒酸试) (14) 。已表明，代谢产物
三氯乙醇是在吸入时引起长期对中枢神经系统作用的原 因 (15) 。但是，就已报道的三氯乙烯的致癌和致突变作用来 说，代谢途径是极为重要的，而最终代谢物却是次要的。代
谢途径的基本特点是生成具有活性的环氧化物，即三氯乙烯 氧化物，它可使核酸和蛋白质皖基化 (9 ， 16 I9) ， 随着环氧化物 N
水解酶的抑制，这种共价结合可能增加(1 6) 。
2.3.3.4
排泄
三氯乙烯及其代谢物从呼出气、尿、汗、粪和唾液中排 泄 C12 ， 13) ，其从体内排出的半减期为1. 5 小时 (20) 。而三氯乙
酸、三氯乙醇和三氯乙醇的葡糖睦酸苦排泄较慢，在人尿中 测得的三氯乙醇的生物半减期范围为 12.........50 小时，三氯乙酸 为 36.........73 小时(1 5 ， 2 1)。
2.3.4 2.3.4.1
对健康的影响 急性、亚慢性和慢性毒性
从分类上来讲，已知三氯乙烯是中枢神经系统的抑制剂。 事实上，它在医学上巳用作一般的麻醉剂 E2230 蛊然当接触
高浓度时主要的临床表现是直接引起中枢神经系统的抑制， 但是有证据表明，接触三氯乙烯可引起长期的对中枢神经系 统的作用 (23) 。
当三氯乙烯用作麻醉剂时曾见到致命的肝功能衰竭，通
常这涉及到患复杂疾病的病人，如营养不良、毒血症和烧伤以 及接受输血的病人 (22)。实验动物的肝功衰竭是以三氯乙烯代 谢物与蛋白质和核酸的普遍结合为特征的(1 7) 。
三氯乙烯作麻醉剂时，肾功衰竭是不常见，的 E223 ，虽然已 一 240 一
有在实验动物中出现肾功减退的报道，但是却需很高的剂 量 E242 ，相对而言，其作用远远弱于氯仿和四氯化碳。据报道，
由于误用三氯乙烯引起的死亡病例中出现肾损伤 Clk
2.3.4.2
致突变性
据报道，三氯乙烯对许多细菌菌株具有致突变活性。 Greim 等 (25)证实，当与苯巴比妥诱导的小鼠肝微粒体结合
使用，在培养基中浓度为 3.3 毫克分子时引起埃希民大肠菌 K12 的回复突变。在 Aroclor1254 诱导的大鼠肝微粒体或 B6C3-F1 小鼠肝微粒体存在的情况下， 也观察到类似的结果 (27) 。
三氯乙烯使鼠伤寒
沙门氏菌的回变率增加 (26) 。在酿酒酵母 (XV 185"":'14C) 中 但是，关于三氯乙烯的致突变性
还有些疑问。在化学分析时发现工业品三氯乙烯含有氯甲代 氧丙环和环氧丁烧， Henschler(28) 观察到这两种化合物对
鼠伤寒沙门氏菌 (TA 100) 的致突变性比三氯乙烯强，而纯 三氯乙烯的致突变性弱。这些研究者的结论是 z 以前把致突 变活性归于三氯乙烯，但是起码应部分地归于在某些三氯乙 烯的样品中所发现的致突变污染物。在无代谢活化时，三氯 乙烯在致突变试验中均为阴性 (25 27) ， 这表明被鉴别出来的 N
两个直接作用的化合物不能完全说明这种活性。
2.3.4.3
致崎性
小鼠和大鼠在娃振的第 6.........15 天吸入浓度为 1600 毫克/
立方米的三氯乙烯，每天吸入 7 小时，结果未出现致畸作 用 (29) 。尽管没有统计学上的显著性，但是有证据表明引起脑
室出血 (2/12 窝)。还观察到在三氯乙烯处理组小鼠中有几
例辜丸未下降 (2/12 窝) ，但发生率很低。这可能是仅有的 关于三氯乙烯致畸的研究。
2.3.4.4
致癌性
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美国国立癌症研究所 (30) 观察到，经三氯乙烯处理的小鼠
(B6C3-Fl 品系〉肝细胞癌发生率增高。该实验每周给动物
染毒 5 天，共染毒 78 周，给雄性小鼠的时间加权剂量为 1169 和 2339 毫克/公斤体重 F 对于雌性小鼠为 869 和 1739 毫克/ 公斤体重。对于 Osborne-Mendel 大鼠的类似实验未能
引起肿瘤发生率的增高。但是，大鼠对阳性对照四氯化碳的 反应也不敏感，表明 B6C3-F1 小鼠与上述大鼠比较，对氯
代化合物诱发癌症要敏感得多。从小鼠所获得的资料总结于 表 1 。在用低剂量和高剂量染毒的雄性小鼠中均观察到肝细
胞癌转移到肺的某些证据(分别为 4/50 和 3/48 )。 表 1 用三氯乙烯处理的 B6C3-F1 小鼠肝细胞 癌的发生率 (30)
雄性 对照 低剂量
雌性 。 /20
1/20 26/50 31/48
4/50
高剂量
11/47
已表明，在高度敏感的体外 Fischer 大鼠胚胎细胞系 统 (F1706 ，用于鉴定致癌物的测试系统〉中，三氯乙烯能引
起细胞转化。在浓度为 1M/ 升时，三氯乙烯引起大鼠胚胎细 胞转化，其特点是细胞缺乏接触抑制，形成渐进性生长灶， 如果接种在半固体琼脂培养基内，则形成肉眼可见灶。将转 化的细胞接种于新生的 Fischer 大鼠，于接种后 27""""68 天，在 100% 的动物接种部位形成未分化的纤维肉瘤 (3 1)。 已经指出，在国立癌症研究所的生物试验 (30) 中使用的三
氯乙烯含有单官能烧化剂氯甲代氧丙环和环氧丁烧作为稳定 ~242-
剂 (32) 。但是已表明，使用纯度为 99.9% 的三氯乙烯在 Fi
scher 大鼠胚胎细胞系统中也引起细胞转化，而且三氯乙 烯也与细胞大分子形成共价结合 (16叫 9) 。在大多数情况下，
使用放射性标记的三氯乙烯，使共价结合的三氯乙烯与杂质 分离开 El830 这种共价结合的性质通常与化学致癌物有关。此
外，似乎不可能以氯甲代氧丙环和环氧丁;院来解释三氯乙烯
引起的肝肿瘤 p 致癌性 (33) 。
由于它们在水溶液中不稳定，所以不可能在
远离接触部位引起肿瘤。 Henschler 研究了纯三氯乙烯的 一般而言，尽管对于仅在一种动物实验中证实具有致癌
活性的化合物来说，估计其致癌性的强度既是不理想的，也 是不可行的，但是国立癌症研究所对三氯乙烯的生物试验是
研究其经口毒性的唯一长期实验。特别是接触肝脏毒物引起 肝肿瘤会带来诱发肿瘤的机制问题，即是否是启动机制或者 是促进机制。用于估计致癌强度的模式是根据化学致癌物使
体细胞突变的理论，该理论假设了肿瘤启动的遗传基础。迄 今还没有合适的外推模式适用于肿瘤的启动剂，但是由于迫 切需要推荐在污染的地下水中三氯乙烯的可接受水平，因此 采取保守的态度，依据对 B6C3-F1 小鼠所证实的致癌性资 料，使用多阶模式，通过线性外推的方法来估计与接触三氯 乙烯有关的致癌危害 (30) 。基于一个 70 公斤体重的人每天摄
入 2 升水，在一生中可接受的危害水平是在 10 万人口中额外
增加的癌症病例少于一例，计算所得出的三氯乙烯试行建议 值为 30 微克/升。而对接触三氯乙烯工人的流行病学调查尚 不足以进行评价 EMH373.
一 243 一
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trichloroethylene.lndustrial medicine , 40:25 (1 97 1). 2. Pearson , C. R. & McConnell , G. Chlorinated C1 and C~
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to the national primar y drinking water regulations on a treatment technique /or synthetic organics. Washington , DC , US Environmental Protection Agency , 1978. 5. National Research Counci I. Drinking water and
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American Water Works Association , 66:703 (1974). 7. Fishbein , L. Industrial mutagens and potential mutagens. 1. Halogenated aliphatic derivatives.
Mutation research , 32:267 (1 976). 8. Goldstein , A. et al. The absorption , distribution , and elimination of drugs. In: Principles
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drug
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York , John Wiley and Sons , 1974 , pp. 129-154. 9. Danie 1, 1. W. The metabolism of CI-Iabe Il ed trichloroethylene and tetrachloroethylene in the rat , 一 244-
Biochemical pharmacology. 12:795 (1 963). 10. Laham , S. Studies of placental tJ: ansfer
of
trichloroethylene. /ndustrial medicine , 39146 (1 970). 1 1. Ikeda , M. & Ohtsuji , H. A comparative study of the excretion of Fujiwara reactionpositive su bs tances in urine of humans and rodents given trichloro- ortetrachloroderivatives of ethane and the ylene. British
journal
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industrial medicine , 29:99 (1972). 1' trichloroethylene after inhaJation. 1.
12. Kimmerle , G ‘ and Eben , A. Metabolism , excretion and toxicology Experimental exposure on rats. Archives
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toxicology , 30:115 (1 973). 13. Kimmerle , G. & Eben , A. Metabolism , excretion and toxicology of trichloroethylene after inhalation. 2. Experimental human exposure. Archives
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toxicology , 301127 (1 973). 14. Ikeda , M. et a l. Urinary excretion of total trìchlorocompounds , trichloroethanol , trichloroacetic acid , as a measure of exposure to trichloroethylene and tetrachloroethylene. British journal o{ industrial
medicine , 29:46 (1 970). 15. Ertle , T. et a I. Metabolism of trichloroethylene in man.
I. The significance of trichloroethanol in Iong-
term exposure conditions. Archives
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29:171 (1 972). 16. Van Duuren , B. L. & Banerjee. S. Covalent interaction of metabolites of the carcinogen trichloroethylene in rat hepatic microsomes. Cancer
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research , 36:2419 (1976). Bolt ，日. M. & Filser , 1. G. Irreversible binding of 一 245-
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macromolecule~.
Enuironmental health perspectiues , 21 :107 (1 977). 18. Uehleke , H. & Poplawski-Tabarelli , S. Irreversible binding of "C-labelled trichloroethylene in mice liver constituents in uiuo and in uitro. Archiues
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toxicology , 37:289 (1 977). 19. Allemand , H. et a l. Metabolic activation of trichloroethylene illto a chemica lI y reactive metabolite toxic to the liver. Journal
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pharmacology and experimental therapeutics , 204: 714 (1 978). 20. Stewart , R. D. et a l. Observations on the concentrations of trichloroethylene ill hlood and expired air following exposure to humans. American
lndustrial HY9iene Association journal , 23:167 (1962) . 2 1. Ikeda , M. & Imamura , T. Biologic a.l half-life of trichloroethylene and tetrachloroethylene in human subjects. lnternationales Archiu /Ur Arbeitsmedizin , 31: 209 (1 973). 22. Defalque. F. 1. Pharmacology and toxicology of trichloroethy lene. A critical review of the world literature. Clinical pharmacology and therapeutics. 21665 (1 96 1). 23. Grandjean , E. et al. Investigations into the effects of exposure to trichloroethylene in mechanical engineering. Briti sh journal 12:131 (1 955). 24. Klaasen , C. D. & Plaa. G. L. Relative effects of various chlorinated hydrocarbons on liver and kidney ~246 →
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pharmacology , 24'2013 (1 975). 26. Simmon , V. F. et a l. Mutagenic activity of chemicals identified in drinking-water. In , Scott , D. et a 1., eb. ,
Progress in genetic toxicolo [J Y. Amsterdam , ElsevierjNorth Holland Biomedical Press , 1977 , pp. 249-258. 27. Shahin M. & von Barstel R. Mutagenic and lethal effects of benzene hexachloride , dibu, ty 1phthalate and trichloroethylene in Saccharomyces cervisiae.
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2.4 2.4.1 概述
四氯乙烯
四氯乙烯(1， 1 ， 2 ， 2- 四氯乙烯，过氯乙烯 PCE) ·是一种 无鱼不易燃的液体，主要在干洗工业中用作溶剂，其次在金 ‘过氯乙烯，缩写为 PCE ，是四氯乙烯的旧名称。虽然本文徒用现代命名， 但是仍使用缩写 PCE ，以越兔与三氯乙烯的缩写混淆. -248 一
属工业中用作去油污剂(1)。
四氯乙烯广泛分布在环境中，巳发现在水、水生物、空 气、食物和人体组织中都含有微量。 3 。在环境中的最高浓度
见于商业上的干洗业和金属去油污工业时。 尽管四氯乙烯随生产工厂、消费工业和生活污水所排放 的废水而释放到水中，但是据报道，因其具有高度的挥发 性，所以在水体中的浓度很低。在英国生产氯代短工厂周围 的水中 (4) 以及美国的地面水中 (5)检出的四氯乙烯浓度低于
1 微克/升。而在污染的地下水中常常发现其浓度很高。
2.4.2 2.4.2.1
接触途径 水
1976 年 11 月 ........1977 年 1 月美国全国有机物监测调
查 ω发现，在 105 个饮用水样品中的 9 个样品内检出四氯乙 烯，这 9 个阳性样品的浓度范围为 0.2........3.1 微克/升，平均
浓度为 0.81 微克/升。在瑞士污染的地下水中四氯乙烯浓度
高达 954 微克/升的。它是在新奥尔良饮水和居民血浆中均 能检出的两种卤代化合物中的一种 (8) 。在英国城市饮水中含 量为 0.38 微克/升 ω。
2.4.2.2
食物
从英国利物浦港区采集的海产品中四氯乙烯的浓度范围 为 0.5......30 微克/公斤 (4 ， 9)。在食品中的浓度范围从桔汁中的
未检出量〈低于 0.01 微克/公斤〉至英国黄油中的 13 微克/ 公斤 (2) 。
2.4.2.3
空气
，→般环境中四氯乙烯浓度较低。 Pearson 和 McCon nell(4)观察到在英国城市大气中的浓度范围从低于 0.68 -249-'
至 68 微克/立方米。 Loechner(10) 发现在慕尼黑郊区的浓度
为 4 微克/立方米，而在市中心为 6 微克/立方米。在美国 8
个地区的调查表明城区法度达 6.7 微克/立方米，而农村低于 0.013 微克/立方米 (1 1)。与许多有关的低分子量氯代怪溶剂
一样，迄今对四氯乙烯的大量接触是在工业环境中 (12)。它主
要用于纺织和干洗工业 (69%) ，金属去污(1 6%) 和用作化
学中间体(1 2%) 0
2.4.3 2.4.3.1
代谢 吸收
Stewart 等 (13) 发现志愿者连续吸入 2 小时四氯乙烯，
血中浓度接近于稳定状态，这表明四氯乙烯在体内能迅速达 到稳定。而对其经口吸收尚未进行专门的研究，但是根据具 有类似性质的化合物(如三氯乙烯，氯仿〉的资料，有足够 理由相信，它可经胃肠道完全被吸收。
2.4.3.2
生物转化
迄今巳对四氯乙烯在人和实验动物体内的代谢进行了广 泛的研究。从定性的意义而言，似乎在人 (14 ， 15)和实验动 物肘， 17) 中的代谢产物是类似的。四氯乙烯基本上是通过环氧
化物〈四氯乙烯氧化物〉和一种酸性的氯化物中间体〈三氯 乙曹先氯〉被代谢为终产物三氯乙酸的。 Ogata 等 (19)报道，
在 67 小时内人体内存留的四氯乙烯的1. 8% 转化为三氯乙 酸 J 1. 0% 转化为未知的代谢物。
2.4.3.3
排泄
四氯乙烯本身主要是通过肺脏从体内排出 (13 ， 20 ， 2 1)，估
计经呼吸道排出的生物半减期为 65 小时 (20 ， 2 1)。三氯乙酸作
为四氯乙烯的一种代谢产物，经尿排出的半减期为 144 小 -250 一
时 (2 1)。
2.4.4 2.4.4.1
对健康的影响 急性、亚慢性和慢性毒性
和四氯乙烯族的所有其它化合物一样，四氯乙姆的急性 作用主要是对中枢神经系统的抑制。对中枢神经系统长期作 用的唯一指标是在接触浓度低至 100 毫克/立方米，每天 4 小 时，共接触 15........30 天的情况下，引起脑电图的改变。这种改 变与大脑皮质的电阻抗增加有关阳，23) 。据报道，这些作用与 在一些细胞中出现的散在性水肿和细胞质空泡有关 (23)。尽管
从实验动物所获得的资料是有限的，但一般来说它支持所发 现的急性中枢神经系统抑制的表现E242. 正如对人体的临床
研究那样，几乎没有长期接触的资料。几篇零星的病例报 告邸， 26) 以及对一组职业接触者(其浓度为 1890........2600 毫克/ 立方米〉的流行病学调查和临床研究 (27)表明，更严重的中枢
神经系统问题可能与人体慢性接触四氯乙烯有关。但是，往 往由于接触其他溶剂而使这些研究更复杂化了 (28) 。
短期接触较高浓度和长期接触较低浓度的四氯乙烯能引 起狗的肾和肝损伤 (29)0 Kylin 等 (30) 注意到，一次吸入浓度
为 1340 毫克/立方米的四氯乙烯达 240 分钟，导致小鼠肝脏
中度脂肪变性。每天吸入上述浓度 4 小时，每周 6 天，共吸 入 8 周，使肝损伤的严重程度增加 (3 1)。对于接触四氯乙烯 工人的流行病学调查尚不足以评价 EUh342.
2.4.4.2
致突变性
Cerna 和 K ype nova (35)发现在对碱基置换和移码突
变都敏感的沙门氏菌菌株中四氯乙烯引起致突变活性的增
加。但是，在肝微粒体存在时，四氯乙烯对埃希氏大肠菌 k 一 :!51 ~
12 的自发突变率并无影响 E3630
2.4.4.3
致癌性
已证实四氯乙烯为 B6C3-F1 小鼠的肝脏致癌物 E3730 对
Osborne-Mendel 大鼠的实验结果为阴性，但是由于早期
的死亡率高，所以在评价四氯乙烯的致癌性时排除了使用大 鼠的资料。而近来使用四氯化碳作为阳性对照的实验资料表 明，_ Osborne-Mendel 大鼠对氯代怪致肝细胞癌的敏感性 低 E3830
可以根据诱发 B6C3-F1 小鼠肝细胞癌的资料 E373 ，使用
钱性多阶外推模式来估计接触四氯乙烯可能产生的致癌危 害。该模式是根据化学物致癌的体细胞突变理论而建立的，
主要适用于肿瘤启动剂。但是已知引起组织(如肝脏)坏死 性损伤的化学物是通过外遗传机制使肿瘤的发生率增加 C393 ，
然而，还没有对这种作用的外推模式。在没有其他致癌性证
据的情况下使用肝脏毒性剂量的实验资料来估计致癌危害性 有些不妥。但是，由于缺乏其他合适的资料以及对四氯乙烯 污染的地下水迫切需要确定限值，所以采取了保守的方案，
将四氯乙烯作为肿瘤启动剂看待而确定建议的试验性限值。
计算的结果表明，四氯乙烯在水中浓度为 10 微克/升时，假 定每人每天水的摄入量为 2 升，预计在每 10 万人口中增加 的癌症病例少于 1 例。
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Canadian Psychiatric .4 ssociation journa/ , 14'627 (1 969) .
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British med ical journal , 2: 1563 (1 976). 27. Coler , H. R. & Rossmi Il er , H. R. Tetrachloroethylene exposure in a sma Il industry. .4 rchiues 01 industria/
hygiene & occupational medicine , ß:227 (1 953). -255 一
28. Tuttle , T. C. et a 1. A behavioral and neurological
evaluation 0/ dry cleaners exposed to perchloroethylene. Washington , DC , Department of Health. Education and Welfare. 1977 ((NIOSH) No. 77-214). 29. Klaasen. C. D. & Plaa , G. L. Relative effects of chlorinated hydrocarbons on liver and 10:119 (1 96 7), 30. Kylin. B. et a 1. Hepatotoxicity of inhaled trichloroethy lene , tetrachloroethy l.ene i! nd ki~ney
funcqón
in dogs. Toxicology and applied pharmacology ,
chlorof orm.
Single exposure. Acta pharmacologica
et toxicologica.20:16 (1 963). 3 1. Kylin. B. et at llepatotoxicity of inhaled trichloroethylene. Long-term exposure. Acta
pharmacologica et toxicologica. 22 :379 (1 965). 32. Blair , A. et al. Causes of death among laundry and dry cleaning workers. American journal
0/
public
health , 691508 (1979). 33. Blair. A. Mortality among workers in the metal polishing and plating industry. J ournal
0/
occupational med icine. 221158 (1 980). 34. Blair , A. & Mason , T. J. Cancer and mortality in the United States counties with metal plating industries.
Archives
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environmental health. 35:92 (1 980).
35. Cerna , M. & Kypenova. H. Mutagenic activity of chloroethylenes analysed by screening system teets.
Mutation research , 46:214 (1 977). 36. Greim , H. et a 1. Mut电 genicity in vitro and potential carcinogenicity of chlorinated ethylenes as a functiO I1 QHW ED
。f
metabolic oxirane formation. Biochemical
pharmacology , 24:2013 (1 975). 37. National Cancer Institute. Bioassay
0/
tetrachloroethylene /or possible carcinogenicity. Washington , DC , Department of Health , Education and Welfare , 1977 (CAS No. 127-18-4 , NCI-CG-TR-13 (NIH) 77-813). 38. National Cancer Institute. Carcinogenesis bioassay
39.
0/ trichloroethylene. Washington , DC , Department of Health , Education and Welfare , 1976 (CAS No. 79-01-6 , NCI-C6-TR-2 (NIH) 76-802). Schumann , A. et a l. The pharmacokinetics and macromolecular interac (i ons Qf perchloroethylene in mice and rats as related to oncogenicity. Toxicology
and applied pharmacology , 55:207 (1 980).
-257 一
·‘
3. 多环芳怪 (PAH) 3.1 概述
多环芳怪〈以下简称 PAH) 是一大类有机化合物，含 有两个或两个以上的苯环，在某些情况下还带有非芳香环， 邻近环共用两个碳原子。
03\/ .........../ 茶
J 1 1 >\</\/\/ C!|1||11 ，/一一一，/ ~/.........../.........../ r 荧惠 苯并 (a) 在
/"-....
,
/...........~
PAH 是由有机化合物不完全燃烧形成的，也可以由某 些细菌、藻类和高等植物合成(1， 2) 。在水中溶解度相当低，
但可牢固地吸附在颗粒物和底泥上 E33 ，从而导致在水环境中 的浓度高于仅从溶解度考虑所可能含有的浓度(1)。如果海解
在水中和吸附在颗粒物上的 PAH 受到具有足够能量的日光 中紫外线的照射时就会发生光解作用民 5) 。土壤中的某些微
生物可以使 PAH 降解 E63 ，而且在底泥中的 PAH 也至少 可能发生某些降解的。
3.2
接触途径
在环境中存在的 PAH 来自于天然的和人类活动的两
种来源。 PAH 作为一类化合物广语分布在环境中。已在动
-258-
物和植物组织、底泥、土壤、空气和各种水中检出 (8 ， 9) 。
3.2.1
空气
PAH 在空气中的浓度因地点和季节的不同而异，在工
业城市环境中浓度较高，而当寒冷季节使用更多取暖设备时 浓度趋于最高(1 0)0
P AH
的浓度在农付和城市是不同的，
前者约为 0.001""""2 微克 /1000 立方米，后者约为 0.1........60 微
克 /1000 立方米 (1 口，已有报道在某些污染的大气中苯井( 蓝的浓度变动在 20----400 微克 /1000 立方米 (12) 。
a)
3.2.2
盒楠
PAH 在某些食品中含量很高，这取决于烹调、保存和 贮存的方法，在各种肉、鱼、蔬菜和水果中均能检出 El330 美
国的资料表明，从食物中 PAH 的总摄入量为1. 6-16 微 克/天 (14) 。
3.2.3
水
在地面水中 PAH 的浓度受工业废水排放的影响，巳 表明在德国的各河流中浓度范围为 0.12........3.1 微克/升 EI530
对很多的地下水和饮用水检验了六种 PAH ，即荧窟、苯并 CbJ 荧惠、苯并 CkJ 荧惠、苯并 C a) æ 、苯并 (ghD
:1E和苟并口， 2 ， 3-cdJ 眩，结果表明其总浓度在地下水中一 般不超过 0.05 微克/升 2 在饮用水中不超过 0.1 微克/升 (6) 。 后来 BorneffCl7)进行的大量分析结果表明，仅在 1% 的样
品中六种 PAH 的浓度高于 0.11 微克/升， 90% 的样品浓度
在 0.001""""0.01 微克/升。苯并 (a )æ 在地面水中的浓度范
围为 0.6----114 毫微克/升，在饮用水中为 0.1-23.4 毫微克/ 一 259 一
升 E 盹山。
根据 Borneff(19) 的工作，在地面水中三分之二的
PAH 与颗植物结合，可通过沉淀、絮凝作用手口过滤过程而 去除，其余三分之一溶解的 PAH 经氧化作用可被去除，其 效率因所用的系统而异。水的氯化能去除 50---60% 的苯并
(a) 距，而在现场试验中用活性炭过滤可去除 99% 的苯并 〈川在 E1939
已知在某些情况下，在配水过程中，与内衬沥青的水管 接触会导致水中 PAH 浓度的增高，尤以荧蘑浓度增高特别 明显 (20) 。
估计在发达国家每年每人 PAH 的摄入量为 1---10 毫 克，苯并 (a) 踵为 0.1--- 1. 5 毫克 (21 ， 22) 。但是估计每日经
日摄入量的大约 99% 来自食物，从饮用水中的典型摄入量仅 占总摄入量的 0.1%α32.
3.3
代谢
PAH 具有高度的脂溶性，易于经哺乳动物的内脏和肺 吸收 (24-26) ，首先能迅速地从血液和肝脏清除 E273 ，并广泛分
布于各种组织内，特别倾向于分布在体脂中 EmH2930
PAH 的代谢是通过细胞色素 P450 混合功能氧化酶系 统进行的，其第一步为氧化和经化作用，产生的环氧化物或 盼类可能再经解毒反应生成葡糖昔酸、硫酸盐或谷脱甘肤结 合物，但是某些环氧化物可能代谢成二氢二醇 z 它依次通过结 合而生成可溶性的解毒产物或氧化成二醇-环氧化物 (30 ， 3 1)， 这后一类化合物被认为是引起癌症的终致癌物 (32)0 的代谢物主要以水洛性化合物从尿和粪中排泄 (33 ， 34) 。
PAH
PAH 能诱导其本身代谢所需酶的合成 E353，这意味着
-260-
PAH 的代谢效率将随着 PAH 的继续接触而增高。但是由 于各种外源和内源性诱导剂的作用、遗传因素、年龄、性别 和营养状况均影响这些酶系统，因而难以预计是否 PAH 的 致癌性随着接触 PAH 时间的延长而增高或降低。
很多研究表明，虽然 PAH 具有高度的脂溶性，但是在 动物或人的脂肪中几乎无生物蓄积的倾向，主要因为 PAH 能迅速和广泛地被代谢 (34 ， 38 ， 3 7)。
3.4
对健康的影响
几乎没有任何关于摄入 PAH 后的急性、亚急性和慢性 毒性的资料。但是已经表明能引起皮肤的角化过度、增生和 皮脂腺消失 PS) 。某些 PAH ，如二甲基苯并惠，对大鼠骨髓 和淋巴样组织具有明显的作用 (39) 。但是通常只有当具有致
癌性的 PAH 的剂量足以引起肿瘤时才能引起明显的中毒
表现。
关于 PAH 致畸作用资料很有限，有关苯并 (a) 眩的 资料表明仅当剂量相当高时才呈现此作用邸， 4 1)。
很多的 PAH 在细菌试验系统、体外细胞系和体内姐妹 染色单体交换试验中呈现致突变性E42H453 ，但是这些改变的
意义和对人群的影响尚不清楚，虽然体外的致突变性试验和
致癌性之间具有良好的相关。 很多 PAH 能引起多种实验动物的皮肤和其他上皮组 织的肿瘤 03 ， 46) 。很小量的受试物常常可引起恶性肿瘤，而
且潜伏期很短。对经口染毒时 PAH 的致癌性研究很少，但 已获得阳性结果，特别是胃肠道和食道肿瘤 EI3 ， t730
在环境中很少遇到单一的 PAH ，在 PAH 混合物中可
能发生很多相互作用，因而巳知敢癌的 PAH 的强度可能增
强 ':48吨的。但是这些系统并未被很好地了解，而且关于在环 境中接触 PAH 的意义还不清楚。
对与人体皮肤癌有关职业的流行病学研究为 PAH 在
某些人类癌症中起肯定作用的假设提供了强有力的证 据 (51-53) 。但是与职业接触相比，尚未能提供有关环境接触
或有关接触 PAH 混合物的重要性方面的定量资料。
3.4.1
建议值
基于所有饮用水中 PAH 的含量应相当于未污染的地
下水的前提，世界卫生组织 1970 和 1971 年的标准确定在饮 用水中六种指示性 PAH 的总限量为 200 毫微克/升。这六
种 PAH 是荧惠、苯并 (a) 蓝、苯并 (gh i)挠、苯并 (b) 荧惠、苯并 (k) 荧窟和苟并(1， 2 ， 3-cd) 览。发现 指示性 PAH 在地下水中的浓度为 10~50 毫微克/升，在相 对未污染的河水中为 50~250 毫微克/升，在污染的河水和排 出的废水中含量较高。 后来的研究表明，这些 PAH 在饮用水中的浓度明显 低于标准，而且其浓度明显受到从内涂沥青的配水系统中逸 出的荧踵的影响。此外，所选择的代表性 PAH 和确定的限
量都不是基于任何的毒理学考虑。 1976 年苏联建议苯并 (a) 眩在地面水中的限量为 5 毫 微克/升，并经苏联卫生部批准 (54) 。
虽然 PAH 从饮用水中的摄入量仅占总摄入量的很小
部分，但是由于它具有潜在危害，故应尽量减少接触。目前 除苯并 (a) æ外，对于 PAH 混合物或单个的化合物均
无足够的资料确定其限值。 对苯并( 一 262 一
a )踵的建议值可能将影响饮用水中其他
PAH
的浓度水平， -因为把苯并 (a) 眩降到可接受水平的方法将 会降低所有的 PAH 。
对苯并 (a) 蓝的建议值是基于毒理学考虑和采用 Neal 和 Rigdon(48) 的实验资料。在实验中作者给 CFW
小鼠喂词含苯并 (a )EE为 1......250 毫克/公斤的饲料共约 110 天，结果胃肿瘤〈乳头状瘤和癌〉的发生率呈现与剂量相关
的、具有统计学意义的明显增高。在危险性评价中采用线性 多阶模式，真可接受的危险性为 1/100 ， 000(55) 。
现将建议简述如下 z
1.基于采用多阶模式来分析苯并 (a) æ 现有的毒理 学资料以及考虑到苯并 (a) EE 与水中已知致癌的其他 PAH
有关的事实，推荐苯并 (a) 眩在饮水中的建议值为 0.01
微克/升。 2. 由于 PAH 与悬浮固体紧密结合，所以必要时可采 用处理措施将 j虫度降至可接受的水平飞从而能保证 PAH 含量降至最低水平。 3. 在水处理和配水时，不应将 PAH 带入水中，所以
应该停止在管道内涂沥青。由于认识到去除现有管道中的沥 青内衬是不现实的，所以应该研究和探索使 PAH 的逸出 减至最低的方法。 4. 应该继续监测 PAH 的浓度水平，以便确定本底水
平，从而可以评价任何的改变以及必要时采取治理措施。为 监测 PAH 的水平，使用数种特定化合物作为 PAH 类化
合物总体的指示性化合物是可取的，而指示性化合物的选择 将因具体情况而异。一些专家认为，在某些情况下苯并 (a) æ可作为 PAH 类化合物总体污染的指标。 机第五章第 16.4 节(第 448 页)中限定可接受的水平.
一 263 一
5. 对饮用水中 PAH 的控制应继续遵循下列原则 z 即
未污染的地下水代表了本底污染，不应超过。该原则体现在 1970 和 1971 年世界卫生组织的标准中。这是一个有用的原 则，但是不能应用于所有情况，而且它也不是基于毒理学的 考虑。
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Proceedings 01 the National Acad my øl Sciences , 73 =1 88 (1976). ~269 一
45.
Bay 凹，
U. In , Freudenthal , R.
I. & Jones , P. W. , ed.
Pol ynucl ear aromatic hyd rocarbons: S econd lnternational Symposium on Analysis , Chemistry and Biology. New York , Raven Press , 1978. 46. Ihall , J. Relative potency of carcinogenic compounds.
American journal 0/ cancer , 35.188 (1 939). 47. Neal , J. & Rigdon , R. H. Gastric tumours in mice fed benzo(a)pyrene , a quantitative study. Texas repo 俨 ts
on biology and medicine , 25:553 (1 967). 48. Pfeiffer , E. H. Investigations on the carcinogenic burden by air pollution in man. VII. Studies on the oncogenic interaction of polycyclic aromatic hydrocarhone. Zentralblatt /Ur Bakteriologie ,
Parasitenkunde , ln/ektionskrankheiten und Hygiene (Q rig. B).158:69 (1 973). 49. Pfeiffer , E. H. Oncogenic interaction of carcinogenic
and non-carcinogenic polycyclic aromatic hydrocarbons in mice. In , Mohr , U. et a I., ed. Air pollution and cancer in man. Lyon , International Agency for Research on Cancer , 1977 (1 ARC Scientific PublicatioM No. 16). 50. Schmlhl , D. et a I. Syncarcinogenic action of polycyclic hydrocarbons in automobile exhaust gas condensatee. In , Jl.I ohr , U. et al. , ed. Air pollution and cancer in man. Lyon , International Agency for Research on
Cancer , 1977 (l ARC Scientific Publications No. 16) , 5 1. Pott , P. In: Chirurgical obseruations. London , Hawes , Clark and Collins , 1975. 52. Eckhardt , R. D. lndustrial carcinogens. New York , Grune and Stratton. 1959. 一 270 一
53. Selikoff ,1.
J. Air pollution and asbestos
carcinogenesis. investigation of possible synergism. In. Mohr , U. et a I., ed. Air pollution and cancer in
man. Lyon , International Agency for Research on Cancer , 1977 (l ARC Scientific Publications No. 16).
54. CRules 0/ sur/ace waters protection /rom pollution by sewage and indust 俨 ial wastes.) Moscow , Ministry of Health , 1976.
55. Ambient water qua/ity criteria /or polynuclear aromatic hydrocarbolls. Washington , DC , US Environmental Protection Agency , 1980 (E P A 440/570-069) •
-271 一
·、
6
4. 农
药
对水质可能具有重要意义的农药包括氯代怪及其衍生 物、持久性除草剂、土壤杀菌剂、易从土壤中逸出的农药和 h 控制疾病传播媒介或其他目的而经常加入给水中的农药。
但是，其中仅仅氯代怪杀虫剂经常被发现。 d氯代怪农药长期存在于环境中，并且分布极广。例如已
在几干米高空的飘尘中和南极融化的雪水中发现了微量的滴 滴涕。水中的氯代怪农药可能呈进行性地蓄积在食物链的不 同阶段 F 例如滴滴涕在鱼体中生物蓄积的浓度可能是栖息地
水中浓度的一万倍以上。已表明该组中的数种农药，包括过 去被广泛地用于农业上和继续被用于诸如控制疾病等目的的 农药，对动物具育致癌性。
已经推荐了数种氯代居农药的建议值，因为已知它们是 水中残留的外来物质。该值是根据联合国粮农组织/世界卫 生组织农药残留量联合专家委员会规定的日许量 (ADD 确
定的 F 以 1% 的日许量作为饮水中建议值的基础。如果使用
钱性多阶外推模式来估计建议值的潜在致癌危害，结果表 明，含有该建议浓度残留农药的水会使在 10 万人口中终生增 加的癌症不超过一例。短时间超过该值〈例如为控制昆虫媒
介〉是可以容许的，但应对健康影响进行监测和评价。 已经确定了建议值的农药并不能代表已在水中鉴别出来
的全部农药。各地的具体情况可能要求使用尚未确定建议值 的农药，因此，对饮水中存在的这些农药进行监测是比较理 一 272 一
想的。
建议值是钟对保护人体健康而确定的，所以这个值不一 定适用于保护水生物。
4.1 4.1. 1 概述
滴滴涕〈全部异构体〉
滴滴涕是 1 ， 1-(2 ， 2 ， 2-三氯亚乙基〉双-(4-氯苯)的字 首缩写，也称为 p ， p'-DDT 。其结构有数种不同的异构体 形式。 DDT 的名称也用于商业产品，该产品主要含有 p ， p'-DDT 和一些 0 ， p'-DDT 以及少量的 p ， p'-TDE 、 0 ， p'一
TDE 、 物a 。
p ， p'一 DDE 、
0 ， p'-DDE 和其它未鉴定的化合
DDT 几乎不溶于水，但溶于有机溶剂。蒸气压在 20 0C 时为 2.53
x 10- 5 Pa (1 .9 x
10- 7 毫米来柱〉。
在第二次世界大战时 ， DDT 首先用于军事区域和个人 的防护，主要是被用于预防疙疾、斑痊伤寒和某些其他的虫
媒疾病。而广泛的农业应用则是在 1946 年始于美国，稍后 在大多数其它国家也陆续使用。 DDT 还广泛用于森林。但
是，由于生态学原因以及昆虫对这种杀虫剂的抗药性增加，
所以在一些国家已经被限用，甚至禁用。特别是在 1970 年 以后巳观察到生产量降低，然而在某些热带国家， DDT 仍 广泛用于农业和控制昆虫媒介。 DDT 是一种持久性杀虫剂，在大多数环境条件下稳定，
并能抵抗土壤微生物和较高等生物体内酶的完全分解作用。 • p , p'-TDE (也称为 DDD) 是 1 ， 1 '- (2 , 2-二氯亚乙基)双 (4-氯苯) ;
p. p'-DDE 是 1 ， 1'-(2 , 2-二氯亚乙烯基)双 (4- 氯苯)。 "4 。。 。，"
某些代谢物，特别是 1 ， 1-(2 ， 2一二氯亚乙烯基)双 (4一氯苯) (DDE) 的稳定性相当于或大于 DDT 。世界卫生组织已经
发表了关于 "DDT 及其衍生物"的环境卫生基准文件 ω 。
4. 1. 2
接触途径 空气
4. 1. 2.1
在农田使用 DDT 后 6 个多月，仍可检出 DDT 从农田的 蒸发，而大部分蒸发的 DDT 又会返回使用区域的土壤中，
其返回的数量与使用 DDT 地点距离的对数几乎成直线反比 关系。但是有少量 DDT 会被转移到全世界的范围，已知从 一千多公里高空飘移的尘土中发现微量的 DDT 。正常情况
下，非农业区空气中的浓度范围为 1........2.36x10- 6 毫克/立方 米，而在进行灭蚊喷雾的区域 ， DDT 浓度可能高得多。
一般而言，有关 DDT 及其类似物在环境中循环和结局
的知识是缺乏的。目前已经证实，在类似于上层大气的实验 条件下， DDT 被分解为二氧化碳和盐酸。
4. 1. 2.2
水
DDT 在雨水中的浓度，无论是在农业区，还是在偏僻 的非农业区，通常均处于相同的数量级水平(1 .8 X 10- 5 ........
6.6
X 10- 6 毫克/升〉。
DDT 在地面水中的浓度取决于在雨水和土壤中的浓度
以及土壤的性质。据说在美国的浓度 1966 年达到高峰，随后 在 1967 和 1968 年明显降低 s 那时，与 DDT 有关的化合物
在地面水中的最高浓度为 0.84 微克/升。到 1971 年在联邦 德国的平均浓度为 0.01 微克/升，从未高达 1 微克/升。近年
来，在自来水中的浓度远远低于 1 微克/升，其平均浓度与雨 水浓度类似. 一 274-
4. 1. 2.3
食物
在若干国家已经测定了 DDT 从食物中的每日摄入量。 1953 和 1954 年在美国 DDT 和总 DDT
(DDT + DDE +
TDE) 的平均每日摄入量分别为 0.184 和 0.286 毫克/人， 其中大部分来自于动物性食物。继在家畜中限用 DDT 十年 以后，摄入量降低了 75% 以上。
美国市场篮装食品的调查表明， DDT 的每日摄入量逐 渐降低，到 1970 年为 Q.015 毫克/人。在同期间，加拿大和 英国的摄入量稍低。在很多膳食类同的欧洲国家和其他国 家， DDT 的摄入量大致相同。关于在人体脂肪中 DDT 及
其代谢物贮存情况的世界性测定表明，总接触量的极值技动
范围的系数约为 10 ，但是对于大多数人群的总接触量波动范 围的系数不大于 3 。
食物是一般人群 DDT 的主要来源，人体内贮存的 DDT 90% 以上来源于食物。
4.1.3
代甜
DDT 通过吸入和食入均可被吸收。事实上，对于小剂 量的吸收是完全的，并且因食物中存在脂肪而被加速吸收。 但是难以经皮吸收。 DDT 贮存在脂肪组织中的浓度水平起
初迅速升高，继而逐渐升高，直至达到稳定的状态。人体贮存 DDT 达到平衡所需要的时间至少为一年。如果停止接触， 则贮存在组织中的 DDT 含量会逐渐降低。
象大多数的动物种属那样，人体能把一定量的 DDT 转 化为 DDE 。在组织中也能发现少量的 TDE (DDD) ，它 是在生成主要排泄产物 2 ， 2-双 (4-氯苯〉乙酸 (DDA) 过
程中的一种中间体。 一 275 -
在不同国家一般人群血液中总 DDT 的浓度范围为 0.01""0.07 毫克/升。在胎儿或新生儿血液和其他组织中，
DDT 的浓度低于母体相应组织中的浓度。 通常报道，人奶中的 DDT 浓度范围为 0.01""0.10 毫 克/:1t;而人奶中 DDT 及其代谢物〈特别是 DDE) 的浓度
总和，约为上述浓度的二倍。 一般人群尿中 DDA 的平均浓度是 0.014 毫克/升。
动物实验表明，血清中浓度能最精确地反应在脑这一重 要组织中的浓度。
4.1.4
对健康的影响
DDT 的急性经口毒性受溶剂的影响，其对大鼠的典型 致死中量为 250 毫克/公斤(溶于油中给予) ，而难于经皮吸 收。
DDT 主要作用于神经系统，致使中枢和外周神经系统
均在某种程度上受累。似乎其毒理作用与对神经系统细胞膜 的作用有关。
肝脏是明显受到 DDT 作用的唯一的另一个器官。对于 数种动物种属，致死剂量的 DDT 可能引起肝细胞的局灶性 坏死。 DDT 还引起数种品系小鼠的肝脏肿瘤，在一项研究
中引起大鼠未转移的肝脏肿瘤 p 而对地鼠却未见致癌作 用 (2-4)0 DDT 能诱导所有受试的各种动物的微粒体酶，但
是仅仅对于某些瞄齿类动物引起平滑内质网明显增加，致使
整个肝脏增大，并且在细胞的边缘出现颗粒。这些改变伴随 脂滴的中度增加，其中某些脂滴则形成所谓的脂肪球。在小 鼠和大鼠的长期词喂试验中，肝脏的改变从边缘增生和脂肪
球到形成细胞结节。微粒体酶的其它诱导剂也能引起喝齿类 。，"
动物一系列相同的改变。这种从单个细胞的即刻反应到最终 形成肿瘤的整个过程是某些啃齿类动物所特有的，而对其他
种属动物的反应在形态学上却与此不同。当饲料中浓度高至 200 毫克/公斤时，对大鼠的繁殖功能未引起有害的影响。当
接受的剂量为 10 毫克/公斤/天时，狗的繁殖功能也正常。 在数种动物种属均未观察到致畸作用。在细菌试验系统 中未发现 DDT 是致突变的。而哺乳动物试验系统的体外和 体内试验结果尚不能定论。 对接触 DDT 的浓度高于一般人群长达 25 年的工人进 行了调查，结果未发现 DDT 引起人类癌症的任何证据。虽
然从 1960 至 1981 年对接触 DDT 的工人进行了大量的流行 病学研究，但其结果尚不足以评价对人的致癌性。
对毒性资料进行了评价，并在 1969 年估算了人的每日 暂行容许摄入量，为 0........0.005 毫克/公斤体重 (5) 。
参考文献
1. DDT and its derivatives. Geneva , World Health Organization , 1979 (Environmental Health Criteria 9).
2. Some or !J anochlorine pesticides. Lyon , International Agency for Research on Cancer , 1974 (l ARC Monographs ιon the evaluation of the carcinogenic risk of chemicals to humans , vol. 5). 3. Rossi , L. et a l. Long term administration of DDT or phenobarbital-Na in Wistar rats. International journal
0/ cancer , 19'179 (1977). 4. National Cancer Institute Bioassay , Bethesda , MD , Department of Health , Education , and Welfare , 1978 (Technical Report No. 131). 一去 77-
5. Evaluations
some pesficide residues in /ood. Geneva , World Health Organization , 1970 (F AO/PL: 1969/M/17/11 WHO/Food Add./70. 38).
0/
4.2 4.2.1 概述
艾氏剂和狄氏剂
这两种有关系的农药化学各称如下 z
艾氏剂 (HHDN): 1 ， 2 ， 3 ， 4 ， 10 ， 10-六氯-1 ， 4. 机， 5 ， 8 ， 8a-六氢-1 ， 4- 桥 -5 ， 8-挂二甲撑案。 狄氏剂 (HEOD): 1 ， 2 ， 3 ， 4 ， 10 ， 10一六氯-6 ， 7-环氧一 1 ，
4 ，饵， 5 ， 6 ， 7 ， 8 ， 8a-八氢-1 ， 4-桥-5 ， 8-挂二甲撑茶。 艾氏剂和狄氏剂是持久性杀虫剂，在食物链中富集。狄 氏剂是艾氏剂在动物体内经代谢氧化和在土壤中经化学氧化
而生成的。两种杀虫剂用于针对各种土壤昆虫的土壤处理、 种子处理以及各种农作物的叶子。近几年几乎停止在叶子处
理方面的应用，在其他方面的应用也逐渐被限制或禁 lb
目
前其最大的用途为控制自蚁。在一些国家，需要用其处理水 果树周围的土壤以便形成屏障来对付自蚁对树根和树干的危
害。 自从 1966 年以来，许多专家组已经评价和综述了因使用 文民剂和狄氏剂所带来的危害(1 -5) 。
4.2.2
接触途径
由于艾氏剂在植物和动物体内易于转化为放民剂，所以
在土壤、食物阳水中很少被发现。
4.2.2.1 一 278 一
空气
在大气中检出的狄氏剂浓度很低。据报道，在美国最高 的浓度为 29.7 毫微克/立方米，在伦敦及其郊区的浓度达
1. 9 X 10- 4 毫微克/立方米，它附着在空气中的灰尘上。在美 国已经确定对 8 小时时间加权平均职业接触的阔限值为 0.25 毫克/立方米。 4.2.2.2 水
在雨水中的浓度范围为 5........42 毫微克/升，在地面水(美 国的大部分河流和湖泊〉中为 0-....0.1 微克/升。在饮水中
的浓度为 1........2 毫微克/升。
4.2.2.3
食物
狄氏剂贮存于哺乳动物、鱼和鸟的脂肪组织、肝、脑和 肌肉中，以及食物链的其他部分。鱼可将狄氏剂从水申浓度
为数毫微克/升水平富集到数毫克/公斤体重的水平。美国在 1965........1970 年总膳食摄入量的范围为 0.05-....0.08 微克/公 斤体重/天。在 1966........1967 年英国对总膳食的研究表明，狄
氏剂的摄入量为 0.09 微克/公斤体重/天 F 但到 1970........1971 年，其摄入量降至 0.03 微克/公斤体重/天。据估计，日本的 摄入量为 0.07 微克/公斤体重/天。
1961--1968 年在美国和英国以及 1964........1966 年在六个 其他国家一般人群体脂中狄氏剂的浓度范围为 0.03........0.45
毫克/公斤组织。基于狄氏剂在体脂中的贮存和接触之间存在 着线性关系，估计一个 70 公斤体重的人，对狄氏剂的平均摄 入量范围为 0.01--0.35 微克/公斤体重/天 (5) 。
4.2.3
代制
在所有受试动物种属(小鼠、大鼠、家兔、恒河猴、黑 猩猩〉中， 12-经基狄氏剂和 4 ， 5-文氏剂-反〈式)-二氢二醇 一 279 一
是主要的代谢物。至于这两种代谢物的比率如何，似乎对于 大鼠和黑猩猩， 主要是通过直接氧化将狄氏剂代谢为 12-捏 基狄氏剂，而对于小鼠和家兔的 1-IÇ谢过程，似乎是将环氧化
物的环打开，进而生成二醇。与其他受试动物种属相比，由 于小鼠的这一代谢途径，加之具育较高的代谢速度，从而可
能导致在小鼠肝中产生相当高浓度的 4 ， 5- 艾氏剂-反(式) 二氢二醇。
在狄氏剂的职业接触者和喂词该杀虫剂的猴体内，既未
鉴定出原型狄氏剂，也未鉴定出曾在大鼠尿中发现的代谢物。 狄氏剂是微粒体酶的强诱导剂。但是小鼠对其反应有些
例外，因此这种动物可能不是人类的适宜模型。长期酶诱导 的结果能引起细胞平滑内质网增生以及肝脏的肥大和增生。
人长期摄入大约 0.01 毫克/公斤体重/天，未引起可检出 的微粒体酶的诱导。对于脱离接触的工人，计算所得的狄氏 剂在体内的生物半减期为 7 个月。在接触的最后半年，、血中 狄氏剂的平均浓度是 0.1 毫克/升，相当于平均每天经口摄入 约 0.17 毫克/公斤体重 (2 ， 4) 。
4.2.4
对健康的影响
狄氏剂的主要作用部位是中枢神经系统。急性中毒时中 枢神经系统的兴奋是死亡的原因(1)。在任何体外或体内的致
突变试验中，均未证实狄氏剂具有致突变性。不同动物种属 的致畸研究表明，它不是致畸物。给予小鼠一次剂量高达 15 毫克/公斤体重或给地鼠 30 毫克/公斤体重的狄氏剂引起轻 度畸形，这是由于对于母体的毒性所致。
在小鼠和其他哺乳动物种属进行的各种致癌实验的结果 表明，艾氏剂和狄氏剂对小鼠肝脏具育种属特异性的作用，
-280-
表现为肝肿瘤发生率增高，这→反应是这种动物所特有 的 E43 ，而且在雌、雄两性动物均被证实呈现剂量-反应关系， 在最低剂量(约 0.015 毫克/公斤体重/天〉实验组雌性动物
即出现肿瘤发生率增高。大鼠的现有资料表明，当剂量达 2.5 毫克/公斤体重/天时，未见致癌作用。在地鼠的致癌实 验结果亦为阴性 (6) 。
根据对职业接触者的一项流行病学调查，还不能对其引 起癌症的危害性得出任何结论。飞总之流行病学的研究资料 尚不足以进行评价口， 8) 。
截止 1977 年的毒理学资料支持狄氏剂和文氏剂不是致 癌物的观点，因此对原估计的文氏剂和狄氏剂无论是分别接 触，或是同时接触的日许量为 0.0001 毫克/公斤体重再一次 得到确证 (4)。
参考文献 1. Evaluations
01
some pesticide residues in lood.
Geneva , W orld Health Organization , 1967 (F AO/PL , CP /15; WHO/Food Add. /67.32).
2. Evaluations
01
some pesticide residues in lood.
Geneva , World Health Organization , 1971 (F AO/ AGP /1970/M/12/1; WHO/Food Add. /7 1. 42). 3. WHO Technical Report Series , No. 612 , 1976{Pesticide
resid ues in lood , report of the 1976 Joint F AO/WHO Meeting) . 4. Eva/uations
0/ some pesticide residues in lood. Report 1977. Rome , FAO , 1978 (Plant Production and Protection Paper , No. 10).
5. Some orOanochlorine pesticides. Lyon , International -281 一
Agency for Research on Cancer 1974. (IARC Monographs on the evaluation of carcinogenic risk of chemicals to humans vo l. 5) fi. Cabral ,
J. R. et a l. A carcinogenicity study of
pesticid 吃
dieldrin in hamsters. Cancer I etter , 6'241 (1 979). 7. Van Raalte , H. G. Human experience with dieldrin in perspective. Ecotoxicology and envi ronmental
safety , l'201 (1 977). 8. Deichmann , W. B. & MacDonald , W. E. Organochlorine pesticides and liver cancer deaths in the United States , 1932-1972. Ecotoxicology and environmental
safet y , 1'89 (1 977).
4.3 4.3.1 概述
氯丹
纯氯丹是一种谈黄色液体，分子式 t C10HðCl u 分子
量 : 409.8 (1， 2) 。化学各称 1 ， 2 ， 4 ， 5 ， 6 ， 7 ， 8 ， 8- 八氯一2 ， 3 ，切， 4 ， 7 ， 7a-六氢 -4 ， 7-亚甲桥苟(1)。它是由顺式和反式立体结
构的异构体混合物组成的，分别称为 α-租用-异构体 E330 氯 丹能洛于水中的浓度可以构成对水生物的毒性。 Brooks(3)
报道了氯丹在水中的溶解度在 25 'C时约为 9 微克/升。 氯丹是多环氯代居〈所谓环二烯杀虫剂〉中的广谱杀虫
剂。在过去的三十多年中被广泛用于控制白蚁，用作家庭和 花园的杀虫剂以及生产农作物(如玉米〉时控制土壤昆虫。
但是，近几年来它的使用和产量均大幅度地减少.
一 282-
4.3.2 4.3.2.1
接触途径
水
关于对工业品氯丹在河水中存在持久性的研究表明，
8
周后还存留 85%(4.)。工业品氯丹的主要成分顺式氯丹和反 式氯丹在 8 周中是完全稳定的，但在其余的成分中，除两种 外至少都发生了部分的改变。 Schaefer 等 (5) 检查了来源于密西西比和密苏里河的
500 多个瞬间给水样品，结果在 20% 以上的出厂水中检出 氯丹，最高浓度达 8 微克/升。在雨水中也检出了氯丹 (6 ， 7) 。 虽然有时收到个别家庭由于使用氯丹控制白蚁而造成井
水受到污染的报告，但是仅仅发表了一篇关于城市给水系统 受到污染的报告 (8)0 1976 年 3 月 24 日，在美国田纳西州查
塔努加供应 105 人的一段公共给水系统受到污染，原因是在 负压期间用管子稀释氯丹浓缩液时发生了明显的反虹吸作
用。结果在 71 名受影响的居民中， 13 名出现轻度的急性氯 丹中毒症状，但无一例发生长期的后遗症。
4.3.2.2
食物
美国食品和药物管理局自 1965 年以来已对美国食品中
氯丹的含量进行了定期的监测。在调查的最初 11 年中，不 经常检出氯丹 (9) 。已收集的唯一定量样品是 1972 年的一个
谷类样品，其含量为 0.059 毫克/公斤 (10) 。在大多数近来 (1 975 年〉发表的结果中均未检出氯丹。 美国国立科学院 (1 1)综述了 Moore' 的结果，指出 1971.-
1973 年在伊利诺斯采集的 200 份奶样中， 87% 的样品检 • Moore , 5. 1 Proc 27th Il liuois Custom 5pray oporatoro Training 5çhool. Urhaua , IL , 1975. -283 一
出氯丹为阳性，其平均浓度为 50 微克/升。环二烯类(如氧 丹〉显然是随饲料摄入的，并倾向于在脂类中富集。氧化氯
丹是氯丹和七氯在哺乳动物体内的主要代谢物，于 1973~ 1974 年在阿肯色和密西西比州采集的 57 份人奶样品中，有 46% 的样品中发现这种代谢物，均值为 5 微克/升，最高值 为 20 微克/升 (12)。
4.3.2.3
大气
美国在关于农药对大气污染程度的调查中，在城市和农 业区采集了有代表性的 9 个地区的空气样品，结果在所有地
区至少检出了一种有机氯农药，但是在任何样品中却都未发 现氯丹 (13) 。在由 Nisbet(9) 总结的一项较大规模的调查中，
采集了美国 16 个州 45 个城市的 2479 份样品，结果仅在 2 份样品中检出氯丹，其浓度分别为 84 和 204 毫微克/立方米。
4.3.2.4
其他接触途径
氯丹可经皮吸收而引起毒作用 E143. 经皮接触可能仅仅
在生产时的职业接触或使用农药时才会发生。吸收的量可能 从少至微量直到引起急性作用，这取决于接触的程度。对一
般人群而言，经皮接触可忽略不计。而对于使用氯丹的人来 说，可能造成氯丹长期持久地存在于皮肤上。在一项研究中 报告，一位过去控制害虫的工作人员在已知最后一次接触后 的 2 年，在其己烧洗手液中仍含有氯丹 E153.
4.3.2.5
接触途径的相对意义
对于一个非职业接触者，鱼和贝类可能是接触氯丹的最 主要来源。海产品的相对意义就在于其生物富集系数大，从 而在水中微量的氯丹就可在海产品中富集到高得多的浓度。
相对来说，饮水和大气不是氯丹的主要来源。无疑，职业接 触，特别是对于使用氯丹控制虫害的人来说，是对这种化合
-284-
物的一次最大的接触。
4 , 3.3
代谢
一次经口给予大鼠氯丹，大约被吸收 6%(16) 。每天给
予小剂量，则可吸收 10.-....15% 。给大鼠喂词含氧丹浓度为 1. 5 和 25 毫克/公斤的词料共 56 天，结果在大鼠脂肪中的
浓度约为饲料中浓度的 3 倍，在肝、肾、脑和肌肉中的浓度 分别为饲料中浓度的 12 、 10 、 4 和 2% 。
一旦停止向饲料中加入氯丹，在 4 周内所有存在于体内 的氯丹量稳定地减少，在此期间浓度约降低 60%. 但在以
后的 4 周内，仅稍有降低。 大鼠体内的大部分氯丹从大鼠的粪便排出。仅仅约总摄
入量的 6% 从尿中排泄。但是家兔则不同，从尿中的排泄量 多于粪。
由于一个小男孩偶然误服氯丹，从而获得了有关人体内 氯丹半减期的资料(1 7) 。计算得出对全身的半减期为 21 天， 这比用于治疗的药物长，然而却比其他有机氯杀虫剂短得多。 Barnett 和 Dorough (16) 给大鼠喂词氯丹 56 天，所得到
的半减期约为 23 天，与上述结果类似。当排出量达 60% 以 后，进-~步的排泄是缓慢的。 Aldrich 和 Holmes Cl8 )报
道，在一个小女孩血清中氯丹的半减期为 88 天。 Polen 等(1 9) 以及 Street 和 BlaU(20)发现，氧化氯
丹是氯丹在哺乳动物体内的代谢物，贮存于脂肪组织中。 Street 和 Blau(20) 观察到氧化氯丹的毒性大于氯丹。
Barnett 和 Doroug h (16) 在大鼠排泄物中除发现氧化氯
丹外，还试验性地鉴别出数种氯丹的是基代谢物，因此得出
结论，氯丹的代谢是通过一系列氧化酶的反应而进行的。 一 285-
4.3.4 4.3.4.1
对健康的影响 毒性
有关氯丹对人体毒性的资料通常是由于事故性接触而获 得的。 Curley 和 Garrettson (l 7) 报道，一个 20 个月的 小男孩偶然喝了未知量的氯丹，出现呕吐和阵发性抽搞，每
次持续 3---5 分钟。给予苯巴比妥 04 毫克/公斤体重〉后 停止了发作，体温升至 38.9 "C，然后逐渐恢复正常。未见肺 部疾患的指征。当发作时做神经科检查发现，所有末梢的深
臆反射亢进。但颅神经功能未受损伤，无眼球震颤。接触后
48 小时心电图正常。接触后 3 个月所有试验均正常。 Dadey 和 Krammer(2 1)以及 Aldrich 和 Holmes (l 8)
也报道了类似的情况。 美国国立职业安全和卫生研究所的文献综述 (22)表明，
氯丹的半数致死量范围从家兔经口毒性为 100 毫克/公斤至
大鼠的经皮毒性为 700 毫克/公斤。给大鼠喂饲含氯丹为 2.5 毫克/公斤的饲料引起轻度肝损伤 (11) 。
4.3.4.2
致畸性
在饪振期给大鼠喂饲含氯丹浓度为 150~300 毫克/公斤 的饲料，未发现氯丹是致畸的 (23) 。
4.3.4.3
致突变性
Arnold 等 (24) 给 Charles
River
CD-] 雄性小鼠
一次剂量为 50 或 100 毫克/公斤体重的氯丹，继而将这些动
物与未染毒的雌性小鼠交配，结果未引起显性致死性改变。 Ahmed 等 (25) 用 SV-40 转化的人体成纤维细胞株 VA-4
进行的研究表明，氯丹引起非程序 DNA 合成，说明氯丹
是潜在致突变物。而代谢活{七阻止了非程序 DNA 合成的 一 286 一
诱导。 Simmon 等 (26 )发现纯的顺式或反式-氯丹在 Ames
沙门氏菌微粒体测试系统中都不是致突变的。但是工业品氯 丹在鼠伤寒沙门氏菌菌株 TA 1535 、 TA98 和 TA 100 中
是致突变的，而 S-9 肝活化系统不能增强这种致突变性。
4.3.4.4
致癌性
美国 Gulf South 研究所根据与国立癌症研究所的合
同，对氯丹致癌的可能性进行了生物实验口7) 。每组包括雌、 雄两种性别的小鼠各 50 只，在 35 天龄时给予两种浓度的受
试物共 80 周，随后再观察 10 周，结果小鼠发生肝细胞癌， 并呈明显的剂量关系。
Gulf
South 研究所使用 Osborne-Mendel 品系大
鼠进行了类似的实验。与小鼠的结果相比，给予氯丹的大鼠 未出现肝细胞癌发生率明显增加。 已有数篇关于对职业接触者的流行病学研究。在伊利诺
斯的马绍尔，自 1946 年起生产氯丹，结果在生产氯丹的工 人中未发现可检出的危害，也无任何证据表明氯丹是致癌 的 (28) 。对美国三大虫害控制公司的工作人员〈包括控制自
蚁的操作工〉的研究未发现癌症死亡率增加的证据，也没有 出现因肝癌死亡的病例 (28) 。总之，对于接触氯丹农药使用 者的流行病学调查尚不足以进行评价 E2930
据一篇报告， 14 名患成神经细胞瘤的儿童中有 5 名在 出生前和/或出生后接触了氯丹，在一篇流行病调查中发现，
3 名患急性白血病的人也曾接触过氯丹〈含 3~7% 的七 氯) (30) 。
由于仅仅获得对于一种动物(即小鼠)具有致癌性的明
确证明，所以应根据毒性确定限值，估计对人的日许量为 0.001 毫克/公斤体重 (3 1)。因此对于一个 70 公斤体重的人这
-287-
样得出的日许量应为 0.07 毫克。若将该数值的 1% 归于来
源于水，并假设平均每日水的摄入量为 2 升，则计算所得的 氯丹在饮水中的建议值为 0.35 微克/升或调整为 0.3 微克/
升。 参考文献
1. Windholz , M. The Merck index. Rahway , NJ , Merck and Co. , 1976. 2. Whetstone , R. R. Kirk-Othmer encyclopedia Sons , 1972. 3. Brookes , G. T. Chlorinated insecticides. Cleveland , OH , Chemical Rubber Company Press , 1974. 4. Eichelberger , J. W. & Li chtenberg , J. J. Persistence of pesticides in river water. Envi ronmental science
0/
chemical technology. New York , John Wiley and
and technology , 5:541 (1 97 1). 5. Schaefer , M. L. et a l. Pesticides in drinking water. Environmental science and technology , 311261 (1 969) .
6. Bevenue , A. et a l. Organochlorine pesticides in rainwater , Oahu , Hawaii , (1 972) . 1971~72.
environmental contamination and
0/ toxicology , 8:238 Bulletin
7. Environmental Protection Agency. Consolidated heptachlor /chlordane hearing. F ed eral register. 41 : 7552 (1 976). 8. Harrington , J. M. et a l. Chlordane contamination of a municipal water system. Envi ronmental research , 15: 155 (1 978). -288 一
9. Nisbet ,1. C. T. Human exposure to chlordane ,
heptachlor , and their metabolites. Washington , DC , US Environmental Protection Agency , 1976. 10. Manske , D. D. & Johnson , R. D. Pesticide residues in total diet samples (VIII). Pesticides monitoring
journal , 9: 94 (1 975). 1 1. National Research Council Drinking water and
.health. Washington , DC , National Academy of Sciences , 1977._ 12. Strassman , S. C. & Kutz , F. W. Insecticide residues in human milk from Arkansas and Mississippi , 1973-74.
Pesticides monitoring journal , 10:130 (1 977). 13. Stanley , C. W. et al. Measurement at atmospheric levels of pesticides , Environmental science and
technology , 5: 430 (1 971). 14. Gosselin , R. E. et a l. Clinical toxicology
0/
commercial products , 4th ed. Baltimore , MD , Wi lIi ams and Wilkins Co. , 1976. 15. Kazen , C. et a l. Persistence of pesticides on the hands of some occupationally exposed people.
envi ronmental health , 29: 315 (1 974). 16. Barnett , J. R. & Dorough , H. W. Metabolism of Archives
0/
chlordane in rats. Journal
0/
agricultur-a l and /ood
chemistry , 22:612 (1 974). 17. Curley , A. & Garrettson , L. K. Acute chlordane poisoning. Archives (1 963) .
0/
environmental health , 18:211
18. Aldrich , F. D. & Holmes , J. H. Acute chlordane intoxication in a child. Archives
0/
environmental
health , 19:129 (1 969). 一 289-
19. Polell , P. N. et a l. Characterization of oxychlordane , animal metabolite of chlordane. Bu l1 etin (1 97 1) .
01
environmental contamination and toxicology , 5 1 521 20. Street , J. E. & Blau , S. E. Oxychlordane: accumulation in rat adipose tissue on feeding chlordane isomers or technical chlordane. Journal
01
agricultural and lood chemistry , 201395. (1972). 2 1. Dadey , J. L. & Krammer , A. G. Chlordane intoxication. J ournal Association , 153:723
0/ the American M edical (1 953).
22. National Institute for Occupational Safety and Health. Suspected carcinogens-sub/ile
01
the
NIOSH registry 01 toxic e//ects 0/ chemical substances. 2nd ed. Washington , DC , NIOSH , 1976 (NISH-77-149) . 23. lngle , L. Chronic oral toxicity of chlordane to rats.
Archives 0/ industrial hygiene and occupational medicine , 6'357 (1 952). 24. Arnold , D. W. et al. Dominant lethal studies with technical chlordane , HCS-3260 , and heptachlor. heptachlor epoxide. Journal
01
toxicology and
envi ronmental health , 2: 547
(1 977).
25. Ahmed , F. E. et a l. Pesticide induced DNA damage and its repair in cultured human cells. Mutatation
research , 421161
(1 977). ch 雪 micals
26. Simmon , V. F. et a l. Mutagenic activity of
identified in drinking water. Developments in 27. National Cancer Institute , Bioassay 290 一
toxicology and environmental science. 2:249 (1 977). 01 chlordane lor
possi bl e carcinogenicit y. Be thcsda , MD , Department of Health , Education and Welfare , 1977 8) . (NCI-CG-T r..一
28. Environmentul Protection Agency. Con!olidated heptachlor /chlordane hearing. Fed eral register , 41: 7552~7572 ，
7584--7585 (1976).
29. Wang , H. H. 741 (1 979).
& MacMahon , B. 0/
Mortality of pesticide
.pplicators. Journal
occupational medicine , 21:
30. Some halogenated hydrocarbons. Lyon , International Agency for Research on Cancer , 1979 (I ARC Monographs on the evaluation of the carcinogenic risk 。f
chemicals to humans , vo l. 20).
31. Evaluations
0/
some pesticide
re 宫 idues in
/ood.
(F AO/PL 1967/M/ll/IJ WHO Food Add./68. 30.)
4.4 4.4.1 帽罐
/、
.品.
氯苯
六氯苯是一种自色固体，熔点 230 "C，在水中溶解度低
(6 微克/公斤) C I)，但是溶于有机溶剂。纯品是商业产品， 主要用作杀真菌剂。然而造成六氯苯最大量地进入环境中的 原因是由于生产氯和其他氯代化合物〈特别是溶剂〉时它作
为一种副产品而产生的。 4.4.2 4.4.2.1 接触途径 空气
人通过空气、水和食物接触六氯苯。六氯苯随灰尘颗拉 -291 一
以及由于从浓度高的地点挥发而在空气中扩散。看来空气中 含六氯苯的灰尘颗粒可能是居住在美国路易斯安那工业区附
近的一般人群血液中六氯苯浓度增高的主要原因 ω 。 4.4.2.2 水
在欧洲和美国不同地区的 4 个河水样品、
8 个出厂水样 在美国城
品、 l 个污水处理厂样品以及 7 个化学工厂排出的废水中均 发现了六氯苯 (3) 。在下列样品中已检出六氯苯 z 市的雨水中浓度为 0.........33!:' 毫微克/升 (4) ，在莱茵河 (5); 在 意大利的 108 个地面水样品中平均浓度为 2.5 毫微克/升 (6);
在美国工业区的大多数河水中的残留量一般低于 2 微克/升， 但是在一个样品中却高达 90 微克/升的。 食物
4.4.2.3
六氯苯存在于多种食物中，特别是陆栖动物产品(包括 奶制品和蛋〉中 (8) 。估计 1973 年在美国对于六氯苯从食物 中的平均摄入量为 0.4 微克/天， 1974 年为 0.07 微克/天 (9) 。 在日本的摄入量估计为 0.5 微克/天 ElQ3 ，在澳大利亚为 35
微克/天 (11) 。在澳大利亚和挪威由母乳哺养的婴儿可能摄入 40 微克/天 (11 ， 12) 。在人体脂肪组织中六氯苯平均含量的范
围为 0.02.........8.2 毫克/公斤 E133，在人血样中为 0.004.........0.06 毫克/公斤。 3) 。
在猪和羊体脂中六氯苯的含量分别比词料中的含量高 6' 和 8 倍 Z1430 如果这个比值适用于人体，则某些成人摄入的
六氯苯应为数毫克/公斤体重/天，将这一数值外推到人的血 液，可得出类似的结论。大鼠血液中六氯苯的浓度大约是饲 料中含量的十分之 _(5) 。
目前的证据表明，尽管对所选择的人群〈如工厂工人〉
来说，吸入和经皮接触可能更重要 F 但是对一般人群而言， 一 292 一
从食物摄入可能是六氯苯在身体中负荷的主要来源。
4.4.3
代蹦
经口给大鼠六氯苯 (16) 以后，主要经淋巴系统从肠道缓 慢吸收， 48 小时后被广泛地贮存于脂肪中 (7) 。经腹腔注射
和经口给大鼠 e.c)-六氯苯的定量回收与剂量有关，而且从
粪中回收的 14c 量多于从尿中的回收。在尿中的主要代谢物 为五氯盼、四氯氢醒和五氯硫代酷，其他的代谢物是四氯苯、 五氯苯、 2 ， 4 ， 5-和 2 ， 4 ， 6-三氯盼以及 2 ， 3 ， 4 ， 6- 和 2 ， 3 ， 5 ， 6四氯酌，而 2 ， 3 ， 4- 三氯盼和其他四氯盼仅含有微量。这些 代谢物以结合或游离形式从尿中排出。在粪和脂肪中还发现 原型的六氯苯 (18 吨。。
按剂量 110 微克/天经口给予
Macaca
mula tta
猴
(I 4C)-六氯苯共 11.......15 个月，结果发现在尿中放射性的 50%
为五氯酌， 25% 为五氯苯，其余部分为未鉴定的代谢物和原 型的六氯苯。在粪中， 99% 的放射性为原型六氯苯。在实验 的最后 10 天，雄性动物从尿中排出给予剂量的 7.2% ，从粪 中排出 52% ，雌性动物相应地分别排出 4.6 和 42.2%(22)。
4.4.4 4.4.4.1
对健康的影响
毒性
虽然六氯苯对大多数动物种属的急性毒性低〈大于 1000 毫克/公斤) ，但是长期中等程度的接触却具有广泛的生物学 作用。 给大鼠喂饲含六氯苯的饲料 15 周，观察到亚急性毒性
作用，仅肝和脾出现组织病理学的改变。当饲料中含六氯苯 的剂量低至 2 毫克/公斤词料/天时，肝小叶中心损伤的严重 一 293-
程度就会加重。对大鼠的无作用剂量似乎是 0.5 毫克/公斤 体重/天 (15) 。在类似的研究中发现对猪的无作用剂量是在饲 料中六氯苯的含量为 0.05 毫克/公斤饲科/天 (23) 。
给大鼠剂量为 50 毫克/公斤的六氯苯，隔天一次共 53 周，结果于 9 周以后摄入和排出之间达到平衡。一般而言，
在长期实验中所观察到的改变与短期实验中描述的改变类 似。当停止给予六氯苯后，其排出将缓慢地持续至数月之 久 (20 。
1955 ，.._. 1959 年期间，在土耳其曾发生六氯苯引起的迟 发性皮肤叶琳症的流行 (25) 。在 5 年中出现 600 多病人，估
计受影响的总人口达 3000 人之多。该病的爆发流行是由于 食用了经六氯苯处理的小麦种子。出现的症状包括在身体接 触部位，特别是脸和手，出现水店和表皮松解。估计病人在 皮肤出现明显症状前一个相当长的时间内每天摄入 50 ，.._. 2 00
毫克的六氯苯。这些症状多见于夏季，因为强烈的阳光可使 症状加剧。当停止食用被六氯苯污染的面包以后 20 ，-... 30 天
病情好转，症状消失。但常常复发，这是由于病人再次食用 含六氯苯的小麦或在体脂中贮存的六氯苯重新分配造成的。 在土耳其，由于母亲患六氯苯所致的叶琳症或食用被六 氯苯污染的面包致使其婴儿出现一种紊乱，称为 pembe yara(26)。在母乳中含有六氯苯，至少 95% 的患病婴儿在
一年内死亡。基于毒理学的考虑，联合国粮农组织/世界卫 生组织 (27)建议人的日许量为 0.6 微克/公斤，但是现已撤销。
4.4.4.2
致畸性
已报道，六氯苯可经小鼠和大鼠的胎盘转移 (28 ， 29) 。在
Wistar 大鼠的实验中，观察到六氯苯引起很轻微的致畸 作用，但在同一实验室，于器官形成期给予剂量高达 120 毫 。，" 言 AU
克/公斤体重的六氯苯却未能再现上述作用 (30) 。在其他的研
究中，使用六氯苯和被六氯苯污染的五氯硝基苯(剂量为 100 毫克/公斤体重) ，观察到小鼠出现膊裂和某些肾脏畸 形 (3 1)。
在 4 代试验中，每组包括 10 只雄性和 20 只雌性 Spr
ague-Dawley 大鼠，从断奶起给动物喂词含六氯苯剂量
分别为 0 、 10 、 20 、 40 、 80 、 160 、 320 或 640 毫克/公斤的 词料。结果 F1 代吃奶的幼鼠特别敏感，食用含六氯苯为 320 或 640 毫克/公斤饲料的母鼠所生的很多幼鼠于断奶前 死亡，但未见肉眼可见的畸形 E322.
4.4.4.3
致癌性
已进行的两项研究表明，六氯苯是致癌物。对终生每天 喂饲含六氯苯为 4 、 8 或 16 毫克/公斤体重的地鼠的致癌活 性进行了评价 E333，六氯苯具有多种潜在的致癌活性 F 使肝
细胞瘤、血管内皮细胞瘤和甲状腺瘤的发生率明显增加。未 接触六氯苯的地鼠中，仅 ]0% 发生肿瘤，而每天接受剂量 为 16 毫克/公斤体重六氯苯的地鼠中，有 92% 发生肿瘤。
受试动物肿瘤的发生率与剂量有关s 每天接受剂量为 4 毫克/ 公斤体重的地鼠中， 56% 发生肿瘤，而 8 毫克/公斤体重组 中， 75% 发生肿瘤。在非接触组中，未发现甲状腺肿瘤、肝 细胞瘤或肝血管内皮细胞瘤。当地鼠摄入六氯苯的剂量为 4~16 毫克/公斤体重/天时，其对六氯苯的摄入量接近于土 耳其人因事故食用被六氯苯污染的谷物所估计的摄入量范 围 E333.
对于六氯苯对小鼠终生每天接受其剂量为 6.5 、 13 或 26 毫克/公斤体重的致癌活性进行了评价。每天给小鼠六氯
苯的剂量为 13 或 26 毫克/公斤体重，结果引起肝细胞瘤发 一 295 一
生率明显增加，而在未接触六氯苯的对照组未出现肝细胞瘤 的发生或转移。 Cabral 等的结果阳， 35) 进一步证明了他们
以前的结论 z 六氯苯具有致癌性。但是，给品系 A 小鼠每次
注射六氯苯的剂量为 40 毫克/公斤体重，每周 3 次，共注射 24 次，结果肺部肿瘤的发生率不明显高于对照小鼠 (36) 。六 氯苯对大鼠也是致癌的 (37) 。
基于可接受的危害水平为十万分之一，所以推荐饮水中
六氯苯的建议值为 0.01 微克/升。
参考文献
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& Metcalf , R.
L. Environmental fate and
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14. Hansen , L. G. et a l. Effects and residues of dietary hexachlorobenzene in growing 9wine. J ournal
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applied pharmacology , 40:529 (1 977). 16. Some halogenated hydrocarbons. Lyon , International Agency for Research on Cancer , 1979 (l ARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo l. 20.). 17. Latropoulos , M.
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organotropism of dich lorobipheny 1 (DCB) , die ldri :l, and hexachlorobenzene (HCB) in rats. Environmental
research , 10'384 (1 975). 18. Engst , R. et a l. The metabolism of (HCB) in rats. Bulletin hexachlorobenzen 号
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the American Medical Association , 183:88 (1 963). 26. Cam , C. Une nouvelle dermatose épidémique des enfants. Annales de dermatologie et de
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4.5 4.5.1 概述
七氯和环氧七氯
纯品七氯 (C 1o H 6 CI 7 ，分子量 373.35) 是一种白色结 晶固体，具有棒脑样气味，化学名称为 1 ， 4 ， 5 ， 6 ， 7 ， 8 ， 8-七 氯-3a ， 4 ， 7 ， 7a-四氢 -4 ， 7- 甲撑昂。在 25 "C时的蒸气压为
4 x 10-2 pa (3 x 10-.mmHg) , 25""'29 "C时在水中的溶解度 二-
300-
为 0.056 毫克/升，易溶于相对非极性的恪剂中(1)。
七氯属于多环氯代怪类的一种广谱杀虫剂，被称为环戊 二烯类杀虫剂。从 1971.......1975 年，七氯的最重要用途是控 制农田土壤害虫(1)。
自 1975 年以来，生产和使用量大幅度地减少，这是由于 唯一的生产商自愿加以限制，以及 1976 年 8 月 2 日美国环保 局对七氯在食物性庄稼和家庭中的使用，暂停注册的缘故。但
是，在商业上，仍然继续大量地用于控制自蚁和非大田庄稼。 七氯能在环境中长期存留。并且能在土壤 (2 5) 、植物 (6) N
和哺乳动物 (7) 体内转化成毒性更大的代谢产物，环氧七氯。 在溶液中或薄膜内经光分解作用转变为光七氯 (8) ，后者对昆 虫 (9)和水生无脊椎动物 (10 ， 1 1)的毒性比原化合物更高。 4
七氯及其环氧化物经生物富集在很多种属中，并蓄积在 食物链中 E130
4.5.2 4.5.2.1
接触途径 水
很多研究者已在美国的主要河流流域检出七氯和/或环 氧七氯，两者的平均浓度均为 0.0063 微克J 升 E123，'而七氯 的浓度范围为 0.001"""'0.035 微克/升 (13) 。
4.5.2.2
食物
食品和药物管理局于 1974.......1975 年对 20 个不同城市篮
装食品的研究表明，在 12 类食品中有 3 类环氧七氯的残留 量范围为 0.0006"""'0.003 毫克/公斤 (14) 。已发现，在 1964--
1974 年采集的牛肉、家禽和奶制品中，有 14"""'19% 的样 品含环氧七氧的残留量大于 0.03 毫克/公斤·。从全国得到 • Ni.bett , 1.
C.
T. 人对氯丹、七氯及其代 i射物的接触(未发表的综遮) ~
nv nq
的 590 个鱼样品中有 32% 检出七氯和/或环氧七氯，其整个 全鱼的残留量为 0.01~8.33 毫克/公斤(1 5) 。
人奶可能受到环氧七氯的污染。全国调查表明，在 1936 份母奶样品中有 63.1% 的样品含有环氧七氯 (16) 。 空气
4.5.2.3
七氯能从被处理的表面、植物和土壤中挥发。七氯和较 少量的环氧七氯广泛存在于周围空气中，一般平均浓度约为
0.5 毫微克/立方米。基于该资料，计算所得人的典型接触
为 0.01 微克/人/天气因此，似乎呼吸不是人接触七氯的主 要途径二
4.5.3
代谢
七氯易于被哺乳动物代谢为环氧七氯。该代谢物主要贮 存于脂肪组织中，也贮存于肝、肾和肌肉中(1 7) 。大鼠和狗
能将食入的七氯迅速代谢为环氧七氯，后者主要蓄积在脂肪 组织中。已发现在饲料中七氯的含量与在脂肪组织中七氯的
贮存量之间呈正相关，雌性大鼠在脂肪中蓄积的环氧七氯约 为雄鼠的六倍 (7 ， 18) 。
尽管没有直接的证据表明在人体内七氯被转化为环氧七 氯，但是几乎无疑在人体组织中发现的这种环氧化物是来源 于七氯。已在人的血液、脂肪和乳汁中发现不同含量的环氧 七氯。由于乳汁中脂肪含量高，所以乳汁是有机卤化物(包
括环氧七氯〉的主要排泄途径之一。在美国所进行的一项广
泛调查表明，曾生产过几个孩子而且自己哺乳的妇女乳汁中
农药的含量低于初产妇。环氧七氯、狄民剂和氧化氯丹是人奶 中最常见的农药，仅 2% 的人奶样品有七氯残留，而 63.1%
的样品有环氧七氯残留，其浓度范围为 15~2050 微亮/升 -302 一
〈以脂肪进行校正) ，平均浓度为 91 微克/升。在残留量高 的妇女中，有 11% 的妇女或者是职业接触者，或者是家庭 成员中有职业接触者 E162.
4.5.4 4.5.4.1
对健康的影响 毒性
巳报道，七氯及其代谢物对各种哺乳动物的半数致死量 范围为 6~531 毫克/公斤体重 ω。' ，几乎没有关于慢性作用的资料。当长时间每天给予大鼠 小量的七氯，引起体内葡萄糖平-衡的改变，并认为这种改变 与肝和肾皮质的环腺昔酸-腺苦酸环化酶系统受到剌激有 关 (19-2 1)。
联合国粮农组织/世界卫生组织联席会议 (22) 已经确定七
氯加环氧七氯两者的最大日许量为 0.5 微克/公斤/天。
4.5.4.2
致畸性
长期喂词七氯的研究表明，、在双亲大鼠中以及在子鼠睁 开眼睛后很短时间内即出现白内障 (23) 。
4.5.4.3
对其他生殖功能的影响
在长期喂饲大鼠的实验中表明，七氯能引起子鼠身长的 明显变短以及乳鼠寿命缩短 E232.
4.5.4.4
致突变性
巳报道，七氯在哺乳动物试验中是致突变的，但是在细 菌试验中却呈阴性。对娃振大鼠体内吸收胎鼠数的观察证明， 七氯能引起雄性大鼠出现显性致死性的改变 (24)。在处理组
动物的骨髓细胞中观察到异常的有丝分裂、染色单体异常、破
碎和易位的发生率增加。七氯和环氧七氯在 SV-40 转化的 人细胞 (VA-4) 试验系统中，与代谢活化系统一起培养， -303 一
引起非程序 DNA 合成 (25) 。两者在鼠伤寒险门氏菌的 Ames 试验中均未呈现致突变性 (26) 。
4.5.4.5
致癌性
在小鼠的三项慢性喂饲实验研究中，七氯和/或环氧七 氯能诱发肝细胞癌，环氧七氯在大鼠的一项研究中也引起了 相同的反应 (27)'。
4.5.4.6
建议值
根据粮农组织/世界卫生组织联席会议 (22)关于最大日许
量为 0.5 微克/公斤体重的建议，按一个 70 公斤体重的人每 日水的摄入量为 2 升来计算，在饮水中的建议值为 0.1 微克/
升。由于仅在一种动物种属中充分确证了致癌反应，所以使 用该建议值是合理的。
参考文献
1. Heptachlorl ambient water quality criteria. Washington , DC , US Environmenta~
ProtectioD
Agency , 1980 (Document No. 440/5-80-052). 2. Lichtenstein , E. P. Insecticidal residues in various crops grown in soils treated with abnormal rates of aldrin and heptachlor. Journal 01 agricultural and
lood chemistr y , 8= 448
(1 960).
3. Li chtenstein , E. P. et a I. Degradation of aldrin and heptachlor in field soils. J ournal 01 agricultural
ànd lood chemistr y , 18: 100 (1 970儿 4. Lichtenstein , E. P. et a l. Effects of a cover crop ·美国环保局。对氯丹和七氯危害性的评价。致癌;物评价小组，华盛顿地
区，
1977
(未发表的报告九
-304 一
versus soil cultivation on the f ate of vertical distribution of insecticide residues in soil 7 to 11 years after soil treatment. Pesticides monitoring ‘
journal , 5:218 (1 97 1). 5. Nash , R. G. & Harris , W. G. Chlorinated hydrocarbon insecticide residues in crops and soi l. J ournal
0/
environmental quality , 2:269 (1 973). 6. Gannon , N. & Decker , G. C. The conversion of aldrin to die Idrin in plants. J ournal
0/
economið
entomology , 51:8 (1 958). 7. Davidow , B. & Radomski , heptachlor. Journal
J. L. Isolation of an
epoxide metabolite from fat tissues of dogs fed
0/
pharmacology and
experimental therapeutics , 107:259 (1 953). 8. Benson , W. R. et a l. Photolysis of solid and dissolved dieldrin. Journal
0/
agricu 1t ural and
/ood
chemistry , 19:66 (1 97 1). 9. Khan , M. H. et a l. Insect metabolism of photoaldrin and photodieldrin. Science , 164:318 (1 969).
10. Georgackakis , E. & Khan , M. A. Q. Toxicity of the photoisomers of cyclodiene insecticides to freshwater animals. Nature , 233:120 (1 971). 1 1. Khan , M. A. Q. et a l. Toxicity-metabolism relationship of the photoisomers of certain chlorinated cyclodiene insecticide chemicals. Archives
0/
environmental
contamination and toxicology , 1:159 (1 973). 12. Chlordane and I.eptachlor in relation to man and
t he envi ronment. Washington , D C , US Environmental Protection Agency , 1976 (E P A 540/476005).
13. Breidenbach , A. W. et a I. Chlorinated hydrocarbon qAW
pesticides in major river basins. 1957-65. Public
health reports , 82:139
(1 96 7>.
14. Johnson , R. D. & Manske , D. D. Pesticide and other chemical residues in total diet samples (X I).
Pesticid es monitoring journal , 11: 116 (1 977). 15. Henderson , C. et a l. Organochlorine insecticide residues in fish (National Pesticide Monitoring Program). Pesticides monitoring journal , 3:145 (1 969) •
16. Savage , E. P. National study to determiné levels
0/ chlorinated hydrocarbon insecticides Agency , 1976 (E P A/540/9-78/005).
in humall milk. Washingdon , DC , US Environmental Protection
17. Evaluations
0/
some pesticide residues in /ood.
Geneva , World Health Organization , 1967. 18. Radonski , J. L. & Davidow , B. The metabolic of heptachlor. its estimation. storage and toxicity.
Journal
0/
pharmacology and experimental (1 953).
therapeutics , 107: 266
19. Kacew , S. & Singhal. R. L. The influence of p , pDDT , and chlordane , heptachlor and endrin on hepatic and renal carbohydrate metabolism and cyclic AMP-adenyl cyclase system. Li/e sciences , 13:1363 (1 973).
20. Kacew , S. & Singhal , R. L. Effect of certain halogenated hydrocarbon insecticides on cyclic adenosine 3' , 5'-monophosphate- 3 H formation by rat
2 1. -
0/ pharmaco!ogy and experimental therapeutics , 188:265 (1 974). Singhal , R. L. & Kacew , S. The role of cyclic kidney cortex. Journa/
AMP
306 一
in chlorinated hydrocarboninduced toxicity.
Federation proceedings , 35:2618 (1 976). 22. 1971 evaluations 01 some pesticide residues in
lood. (AGP
,
1971/M/9/1
, WHO
Pesticide Residues
Series , No. 1 , p. 314). 23. Mestitzova , M. On reproduction studies on the occurrence of cataracts in ra ts after long-term feeding of the insecticide heptachlor. Experientia. 23:42 (1 967). 24. Cerey. K. et a1. Effect of heptachlor on dominant letbality and bone marrow in rats. Mutation
research , 21.26 (1 973). 25. Ahmed , F. E. et a l. Pesticide-induced DNA damage and its repair in cultured buman cells. Mutation
r ,. search , 42: 161 (1 977). 26. Marshall , T. C. et a l. Screening of pesticides for mutagenic potential using Salmonella typhimurium mutants. Jou 俨 nal
01 agricultural and lood
chemistry , 24.560(1976). 27. Epstein , S. S. Carcinogenicity of heptachlor and chlordane. Science 01 the total environmen t. 6: 103 (1976) •
4.6 4.6.1 概述
林丹
林丹 (γ-六氯环己院〉又称为 γ-HCH 或γ-BHC ，是一
种白色固体，溶点 112.5 0C ，在水中易恪解(1 0 毫克/升) (1),
但是更易溶于有机i窑剂。它是一种广谱杀虫剂，属于环状氯 。。
代碳氢化合物类，称为有机氯杀虫剂。林丹广泛地应用于对 动物、建筑物、人〈外寄生物)、衣服、7]<. (灭蚊)、植物、种 子和土壤的处理 (2)。
林丹能被土壤微生物缓慢地降解 (3) ，并能被微生物和植 物异构化为 α 和/或 δ 异构体 ω 。
4.6.2 4.6.2.1
接触途径 水
水的污染来自于为控制蚊虫在水中直接应用的工业品六 氯环己;民 (HCH) 或林丹，在农业和森林中应用的 HCH 、
以及来源于工厂废水的偶然污染 ω 。 对美国出厂饮水的一项调查表明，林丹的最高浓度是 0.1 微克/升 (4) 。林丹是地面水的一种普遍污染物，浓度达
100 毫微克/升 (5 ， 6) ，这主要是由于在雨水中的挥发和沉
降 (7) 。在联邦德国，林丹存在于全部采'集的地面水样品中， 其浓度范围为 0.005""7.1 微克/升。
4.6.2.2
食物
据报道，林丹的每日摄入量为 1""5 微克/公斤体重，而 所有其他 HCH 异构体的摄入量为 1""3 微克/公斤体重 (8) 。 人膳食中 HCH 残留量的主要来源为奶、蛋及其他奶制品 (2) 。 美国环保局 (2) 估计，林丹生物富集系数的加权平均值为
780 ，这是根据测定翻车鱼已达到恒定状态时的生物富集而 估计的。
4.6.2.3
空气
在伦敦中心和郊区的空气中均已检出微量的 HCH
(2),
估计通过吸入而被吸收的林丹剂量为 0.002 微克/公斤体重/ 天 E的。
-308-
4.6.3
代谢
脂类载体能加快林丹的吸收速度。与其他有机氯农药相 比， HCH 和林丹通常易溶手水，所以能迅速被吸收和排 泄 (2， 10) 。林丹能经口和经皮被吸收 (2) 。
给实验动物林丹后，在动物脑中检出的林丹浓度高于其 他器官(1! -13) 。当腹腔注射 I.C-林丹后，结果至少给予剂量
的 75% 毫无例外地出现在皮肤、肌肉和脂肪组织中(14)。林
丹可通过胎盘进入人的胎儿 z 发现在其皮肤巾的浓度高于脑: 但却从不超过成人器官的相应值(1 5-16)。
林丹被代谢为一种中间体六氯环己烧，并进一步被降解
为 2 ， 3 ， 4 ， 5 ， 6-五氯-2一环己皖-1-盼、两个四氯酷和三个三 氯盼(1 7) 。这些化合物通常以结合物的形式见于尿中 E1830 无
论是游离型或结合型的氯酷，其毒性均明显低于原型化合 物 EI930
4.6.4 重 (20) 。
对健康的影响
林丹对大鼠的经口半数致死量为 125"""230 毫克/公斤体 在大鼠的慢性毒性研究中，动物接受溶于油中的林丹， 结果在高剂量时见到肝细胞肥大〈脂肪变性和坏死〉和肾脏病 变 (21-23) 。每天给大鼠吸入林丹 (0.78 毫克/立方米 )7 小时，
每周 5 天，共 180 天，观察到大鼠肝细胞增大，但却未见毒 性症状或其他异常 (24) 。给大鼠喂词含 10 毫克/公斤林丹的
饲料 1 或 2 年，结果在实验开始后 5 个月出现体重降低以及 尿、血和组织中的抗坏血酸水平发生变化 (25) 。从饲料中慢性
给予林丹的狗呈现肝轻度增大 E263. nwv
据报道，长期或不恰当地使用含 10 克/公斤林丹的一种
制剂治疗人的所庵，可引起中枢神经系统兴奋和其他副作用 〈恶心、呕吐、抽搞、呼吸减弱并伴随紫绪和血液疾患) (27) 。
接触工业品 HCH 的生产工人出现头痛、头晕以及皮肤、眼 相呼吸道粘膜的剌激症状。在某些情况下，还出现糖类和脂
类代谢的明显障碍和下丘脑-垂体-肾上腺系统的功能失 调 E2Bp2930 从事职业接触 HCH 达 11--23 年的工人出现中毒 性肝炎的生化改变 (30) 。
联合国粮农组织/世界卫生组织联席会议估计，林丹的最 大日许量为 10 微克/公斤体重 (3 1)。
4.6.4.1
致畸性
每天给娃振大鼠含林丹的词料，其剂量为 12 或 25 毫克/ 公斤体重，结果未见致畸作用 (32) 。
4.6.4.2
J才其他生殖功能的影响
在大鼠的四代慢性林丹喂词实验中观察到，平均旺报周 期延长，出生的仔鼠数减少，死胎比例增加以及第二代 (F 2 )
和第三代雌鼠性成熟延迟。此外，还看到某些第一代 (F 1 ) 和第二代 (F 2 ) 动物出现瘟孪性截瘫 E2530
在娃振期经词料给大鼠和家兔林丹引起着床后胚胎死亡 率增加 (32 ， 33) 。
4.6.4.3
致突变性
有关林丹致突变性的证据是不一致的。已报道，当林丹 浓度为 0.1--10 克/升的条件下培养时，人淋巴细胞的有丝 分裂活性和核型发生某些改变 (34) 。而在一项显性致死试验 a
或一项宿主问介试验 (35) 中，林丹却未表现有致突变性。但在 ·美国环保局， BHC-林丹，华盛顿地区， 基准和评价研究室(未发袤的报
告)。 nυ
使用鼠伤寒沙门氏菌和代谢活化系统的微生物试验、大鼠的
宿主间介试验和显性致死试验中表明林丹是致突变的。其他 的报告表明，林丹无明显的致突变活性 ω 。
4.6.4.4
致癌性
已报道，给不同品系的雄性或雌性小鼠喂词林丹 (γ HCH) ，引起肝脏肿瘤发生率增加 (36"'40) 。但是流行病学资 料尚不足以评价 (40 ， 4 1)。
4.6.4.5
建议值
基于联合国粮农组织/世界卫生组织会议对一个 70 公斤 体重的人所建议的林丹最大日许量 (3 1)，以及假设日许量的
1% 可能来自于水和每人每天摄入 2 升水，经计算所得建议
值为 3 微克/升。 参考文献
1. Ulma.ll n'，冒.， ed. Lind ane , monograph 0/ an insecticide. Freiburg , Verlag K. Schillinger , 1972. 2. Hexachlorocyclehexane , ambient water quality
criteria. Washington , DC , US Environmental Protection Agency , 1979. 3. Mathur , S. P. & Saha , J. G. Microbial degradation of lindane-C-14 in a flooded sandy loam soi l. .SoiJ '
science , 120:301 (1 975). 4. US Environmental Protection Agency. Preliminar y
assessment 0/ suspected carcinogens in drinking-water. Report to Congress. Washington , DC , 1975 , p. H-4. 5. Cro11 B. T. Organo-chlorine insecticides in water-Part
1. Water treatment examination , 181255-274 (1 969). ← 311~
6. Greve , P. A. Potentia lI y h.zardous substances In surface wateu. 1. Pesticides in the R. Rhine.
Science
0/ the total envlronment.
1'173-180 (1972).
7. Tarrant , K. R. &Tatton , J. O'G. Organo-chlorine pesticides in rainwater in the British Isle8. N ature , 219'725-727 (1968).
8. Duggan , R. E. & Duggan , M. B. Residues of pesticides in milk , meat and foods. In. Edwards , L. A. , ed. ,
Environmental pollution /rom pesticides. London , 1973 , p. 334. 9. Barney , J. E. Pesticide pollution of the air studied.
Chemical and engineering news , 47'42
(1 969).
10. Herbst , M. & Bodenstein , G. Toxicology of lindane , In , Ulmann , E. , ed. Lindanea monograph 23. 11. Lang , E. P. Tissue distribution of a toxicant following oral ingestion of the gammaisomer of benzene hexachloride by rats. J ournal
0/
an
insecticide. Freiburg , Verlag K. Schillinger , 1972 , p.
0/ pharmacology and (1 948).
experimental therapeutics , 93'277
12. Davidow , B. & Frawley , J. P. Tissue distribution , accumulation and elimination of isomers of benzene hexachloride. Proceedings
0/ the Society /or (1 951).
Experimental Biology and M edicine , 76'780
13. Huntingdon Research Centre. In , Ulmann , E. , ed.
Lindane. monograph 0/ an insecticide. Freiburg , Verlag K. Sch iIl inger , 1972 , p. 97. 14. Koransky , S. et a l. Absorption , distribution and elimination of alpha- and betabenzene hexachloride.
Archiv /ilr exþerimentelle Pathologie und -312-
Pharmakologi e , 244: 564 (1 963). 15. Poradovsky , R. et a I. Transplacental permeation of pesticides during normal pregnancy. Ceskoslovenska
gynekologie , 42:405 (1 977). 16. Nichimura , H. et a l. Levels of polychlorinated biphenyls and organochlorine insecticides in human embryos and fetuses. Pediatrician , 6145 (1 977). 17. Chadwick , R. W. et a l. Dehydrogenation , a previously unreported pathway of lindane metabolism in mammals. Pesticide biochemistry and physiology , 6:575 (1 975). 18. Chadwick , R. W. & Freal , J. J. The identification of five unreported lindane metabolites recovered from rat urine. Bull etin 0/ envi ronmental contamination and toxicology , 7:137 (1 972). 19. National Research Counci l. Drinking water and
health. Washington , DC , National Academy of Sciences , 1977 , p. 939. 20. 1966 Evaluations pp. 126-147. 2 1. Fitzhugh , O. G. et al. Chronic toxicities of benzene hexachloride , and its alpha , beta , and gamma isomer.
0/ some pesticide residues in /ood. Geneva , WHO , 1967 (WHO/Food Add. /67.32)
Journal 0/ pharmacology and experimental therapeutics , 100: 59 (1 950). 22. Lehman , A. J. Chemicals in food: A report to the Association of Food and Drug Officials. US
Association 0/ Food and Drug Officials quarterly bulletin , 16:85 (1 952). 23. Lehman , A.
J. Chemicals in food l A report to the -313 一
Association of Food and Drug Officials on current development. Part II , Pesticides. Section V: Pathology.
US Association
0/
Food and Drug O//icials
quarterly bulletin , 16:126 (1 952). 24. Heyroth , F. F.Jn , Ltlland , S. 1., Chemical Specialities
M anu/acturers Associ ation. Proceed ings annual meeting , 6: 110 (1 952).
0/
the
25. Petrescu , S. et a 1. Studies on the effects of long term administration of chlorinated organic pesticides (lindane , DDT) on laboratory white rats. Revue
médico-chirurgicale
(J assy)
,
78'831 (1 974).
26. Rivett , K. F. et a 1. Effects of feeding lindane to dogs for periods of up to 2 years. Toxicology , 9:237 (1 978). 27. Lee , B. et a I. Suspected reactions to gamma benzene hexachloride. J ournal
0/
the American M edical
Association , 236:2846 (1 976). 28. Kazahevich , R. L. State of the nervous system in persons with a prolonged professional contact with hexachlorocyclohexane and products of its synthesis.
Vrachebnoe delo , 2:129 (1 974). 29. Besuglyi , V. P. et a 1. State of health句 of
persons
having prolonged occupational contact with hexachlorocyclohexane. Zd ravookhranenie Belorussi i , 19: 49 (1 973). 30. Sasinovich , L. M. et a I. Toxic hepatitis due to prolonged expo~ure to BHC. Vrachebnoe delo , 10:133 (1 974) .
31. Pesticide residues in /ood. Report 1977. FAO Plant production and protection paper , Rome , 1978. 32. Mametkuliev , C. H. Study of embryotoxic and 314 一
teratogenic properties of the gamma isomer of HCH in experiments with rats. Zdravookhranenie
Turkmenistana , 20:28 (1 978). 33. Palmer , A. K. et a l. Effect of lindane on pregnancy in the rabbit and rat. Toxic%.gy , 9:239 (1 978).
34. Tsoneva-Maneva , M. T. et a l. Influence of diazinon and lindane on the mitotic activity and the karyotype of human lymphocytes cultivated in vitro.
Bib/iotheca haemat%gia , 38:344 (1 971). 35. Buselmair , W. et a l. Comparative investigation on the mutagenicity of pesticides in mammalian test systems.
M utation research , 21: 25 (1 973). 36. Goto , M. et a l. Ecological chemistry. Toxizitat von a-HCH in mausen. Chemosphere , 1 :153 (1 972).
37. Hanada , M. et a l. Induction of hepatoma in mice by benzene hexachloride. J apanese journa/
0/
cancer /or
research , 64 511 (1 973). ' 38. National Cancer Jnstitute. Bioassay (Technical Report Series , No. 14
0/
/i ndane
possi ble carcinogenicit y. Washington , D C, 1977
, Department
of
Health , Education , and Welfare Publication No.
(NIH) 77-814). Federa/ register , 42: (1 977). 39. Thorpe , E. & Walker , A. 1. The toxicology of dieldrin (HEOD). 11. In mice with dieldrin , DDT , phenobarbitone , beta-BCH , and gamma-BCH. Food
and cosmetics toxicology , 11 :433 (1 973). 40. Some ha/ogenated hyd rocarbons. Lyo 日，
International
Agency for Research on Cancer , 1979 (I ARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo l. 20). qo FD
41. Eriksson , M. et a I. Soft-tissue sarcomas and exposure to chemical substancesa a casereferent study. British
journal 01 industrial medicine , 38:27 (1 98 1).
4.7 4.7.1 概述
甲氧滴滴涕
甲氧滴滴涕的化学名称为 1 ，
1'(2 , 2, 2-三氯亚乙基)双
(4- 甲氧基苯〉。实际上，它不溶于水，易 j容于大多数的芳 香族有机海剂中。工业品约含 88% 的 p ， p'_异构体，其余部 分为 0 ， p'- 异构体。
甲氧滴滴涕是被用于处理农业庄稼和牲畜的-种杀虫 剂。目前已有关于该化合物的综述 ElH43 ，并将其有关资料总
结如下。
4.7.2 4.7.2.1
接触途径
水
甲氧滴滴涕在水中的半减期约为 46 天。其在河水中的残 留量为 2.9........89.1 微克/升，在湖水中达 0.1 微克/升，在生
物污水处理场的排出水中达到 106 微克/升，在密执安湖的
支流中，为 2.9........89.1 毫微克/升，而对以密西西比河和密 苏里河为水源的 500 个出厂自来水样品的分析，结果却未险 出甲氧滴滴涕。 食物
4.7.2.2
根据对食物中残留量的测定结果，美国在 1965........J970
年计算的平均每日摄入量为 0.5 微克/天。甲氧滴滴涕很少 贮存在组织中和从奶中排泄 α ， 4) 。 一 316 一
4.7.3
代谢
甲氧滴滴涕在大鼠体内被肝脏迅速代谢，其代谢产物主 要从粪中，其次从尿中排出。经口给予小鼠标记的甲氧滴滴 涕，在 24 小时内排出 98 %。甲氧滴滴涕主要是通过甲隧基
团的水解而降解，从而产生带有极性的盼，能被迅速地排出。 甲氧滴滴涕能在大鼠脂肪组织中贮存，并与剂量相关。 当在饲件中的含量为 500 毫克/公斤时，在 4 周内达到平衡 s 停止接触后 2 周内，从脂肪组织中消失。
从动物实验可以得出结论，甲氧滴滴涕代谢的速度快， 并且完全，从而可以解释它的贮存少和蓄积性低。， 2， 43.
4.7.4
对健康的影响
甲氧滴滴涕是一种急性毒性相当低的化合物，大鼠的经
口半数致死量为 3460 毫克/公斤体重。 甲氧滴滴涕对大鼠未呈现致畸性。在细菌、酵母或果蝇 试验中无致突变性。在小鼠的细胞遗传学和显性致死试验中 也呈现阴性。 对经口给小鼠和大鼠(数项实验〉甲氧滴滴 j弟的致癌性
进行了研究，结果在小鼠的实验中获得阴性结果，在大鼠的 三项试验中未能确证甲氧滴滴涕能引起肝细胞癌。总之，现 有资料未能提供证据表明甲氧滴滴涕对实验动物致癌 (4) 。
1977 年再一次确证了 1965 年确定的人的日许量为 0.1 毫克/公斤体重口， 2) 。基于此，一个 70 公斤体重的人每日最 大摄入量为 7 毫克。假如 1% 来源于饮水以及每人每日水的
摄入量为 2 升，则推荐在饮水中的建议值为 30 微克/升。
-317 一
参考文献
1. Evaluation
0/
the toxicity
0/
pesticide residues in
/ood. (F AO/PL/1965'10/ 1J WHO/Food Add. /27.6 日. 2. Pesticide residues in /ood. Report 1977. FAO Plant
Production and Protection Paper , Rome , 1978. 3. Some organochlorine pesticides. Lyon , International
Agency for Research on Cancer , 1974 (l ARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo l. 5). 4. Some halogenated hyd rocarbons. Lyon , International
Agency for Research on Cancer , 1979 (l ARC Monographs on the evaluation of the carcinogenic risk of chemicals to humans , vo l. 20) .
4.8 4.8.1 帽述
2 ， 4-滴
2 ， 4一滴 (2 ， 4-二氯苯氧乙酸〉是用于控制阔叶植物的一 种除草剂和植物生长的调节剂。它的商业生产是通过酷的氧 化生成 2 ， 4-二氯酌，然后再与一氯醋酸反应生成 2 ， 4-1商。
商业配方一般是由这种酸的盐类或脂类组成。分析表明， 通常不含二曙英。 2 ， 4-滴在化学上十分稳定，但其醋类易于水解成游离酸。 在水中易遭破坏。在土壤中不常被检出，因为它能被土壤微 生物破坏，并且据报道，它也不蓄积。当在空中向邻近森林 中喷洒 2 ， 4-1商后，该除草剂能在溪流中被检出，而在处理前
的自来水水源中仅含有较低的浓度 (<0.1 微克/升〉。 -318 一
4.8.2 4.8.2.1
接触途径
饮用水 由于微生物的破坏作用能迅速降低 2 ， 4-滴在
还没有获得有关 2 ， 4-滴在饮用水中浓度水平的资料，估 计其浓度很低 s
污染的地面水中的浓度。
4.8.2.2
食物
巳表明，某些食物已受到低浓度 2 ， 4- 滴的污染 (0.021-0.16 毫克/公斤) (1)。
已有报道指出，应用 2 ， 4-滴处理植物能引起硝酸盐含量
增加。虽然在食用庄稼上使用这种除草剂引起人们的担心，
但是没有证据表明，以这种方式接触 2 ， 4-滴的食物是有毒 的。
4.8.2.3
职业接触
已经报道了在工业上接触该除草剂对健康是有害的 (2) 。
4.8.3
代谢
实际上， 2 ， 4-1商以未改变的形式从人和动物尿中迅速被
排泄。在 2 ， 4-7商被吸收的部位，它可分布在各种组织中，但 是不贮存。接受 1~10 毫克 2 ， 4- 滴的大鼠在 24 小时内几乎
能将其全部从尿和粪中排出，而在较高剂量时，在组织中能 出现一些蓄积。
摄入苯氧酸除草剂对肌肉功能的影响可能与干扰糖类代 谢有关。 2 ， 4-1商能从牛奶中排出，这些牛是在用该除草剂或其醋 类处理过的牧场中生活的。
nwu
4.8.4
对健康的影响
由于使用或生产而接触 2 ， 4- 滴的人主诉易疲劳、头痛、
肝区痛、食欲丧失等，但是这些都是自觉症状，也曾有高血 压和肝功异常的病例。
接触 2 ， 4-滴剂量为 0.43........ 0.57 毫克/公斤体重/天达半 年至 22 年的工人与未接触的人群之间未呈现差异。
关于该化合物致癌性的研究尚无法定论，因为资料不充 分或所用动物数少。但是现有资料表明它不是一种潜在致癌 物。在关于恶性淋巴瘤与接触氯盼和苯氧酸之间关系的病例一 对照研究中，据说 7 个病例和 1 个对照仅仅接触 2 ， 4-滴(相 对危险性为 14.6 ，其 95% 可信限为 2.9........29.9)(3) 。
估计 2 ， 4- 滴的半数致死量大于 90 毫克/公斤体重。
4.8.4.1
建议值
联合国粮农组织/世界卫生组织确定 2 ， 4一滴的日许量为 0.3 毫克/公斤体重 (4 ， 5) 。根据毒性资料， 2 ， 4- 滴在饮水中的
建说值为 0.1 毫克/升。但是，在浓度为 0.05 毫克/升时，某
些人可能察觉到它的臭和睐。 参考文献 1. National Research Counci l. Drinking water and
health. Washington , DC , National Academy Press , 1977. 2. Some /umigants , the herbicides 2, 4-D and 2, 4, 5-T , chlorinated dibenzodioxins and misce//aneous industrial chemicals. Lyon , International Agency for Research on Cancer , 1977 (IARC Monographs on the evaluation of carcinogenic risk of chemicals to human 日，
vo l.
15).
一 320 一
3. Hardell , L. et a l. Malignant lymphoma and exposure to chemicals , especially organic solvents , chl@rophenols and phenoxy acids , a case-control study. Bri tish
journal 01 cancer , 431169 (1981). 4. 1974 Evaluations 01 some pesticide residues ín
lood. Geneva , World Health Organization , 1975 (Pesticide Residues Series , No. 4).
5. 1975 Evaluations 01 some pesticide residues ín
lood. Geneva , W or ld Health Organization , 1976 (Pesticide Residues Series , No. 5).
-321 一
5. 氯苯 一氯苯广泛用作溶剂和生产数种化学物质，如杀虫剂和
盼。二氯苯是染料的重要中间体。 1 ， 2-二氯苯用作溶剂和农
药 J 1 ， 4- 二氯苯用作驱蛙虫剂和除臭剂。 1 ， 2 ， 4-三氯苯用作 溶剂、电解质液、传热介质和杀虫剂。在四氯苯中，
1, 2, 4,
5- 四氯苯用作化学合成(例如生产 2 ， 4 ， 5-三氯酷)的中间 体(1)。五氯苯未获得广泛应用。六氯苯已在农药一节中论
述，故不在本节中讨论。某些低氯苯是水加氯消毒的副产品。 一氯苯在地下水、地面水和饮水中检出，其浓度达 10 微克/升 (1 ， 2) 。二氯苯常在原水水源中发现，浓度为 1.........10
微克/升或以上 (3) 。在水环境中的主要二氯苯是 1 ， 2-二氯苯
和 1 ， 4-二氧苯。在饮水中，这些化合物的浓度为 0.01.........1 微 克/升。在三氯苯中， 1 ， 2 ， 4-三氯苯最常被检出，在饮水中 的浓度波动在 0.01.........1 微克/升。飞尚缺乏有关其他三氯苯
和四氯苯在水中存在的资料，说明其浓度可能低于 0.1 微克/ 升，这个浓度是所用分析方法的检出限。 根据预计的氯苯在饮水中的最高浓度以及毒性资科和嗅 阔浓度，可以选定需要确定建议值的化合物 (4月。该资料汇 总于表 2 。 鉴于从感官和毒理学特性来考虑，这些化合物在饮水中 检出的浓度很低以及缺乏毒理学资料，仅选择下列三种化合 物作进一步的考虑 z 氯苯 J 1 ， 2-二氯苯 J 1 ， 4-二氧苯。
如表 2 所示，氯苯的感宫和毒理学限值均在同一数量级。 qaM 90 nrid
表 2
从毒理学考虑所选择的氯苯基准值 ←京示
最高浓度
一一-一一诙主f二
嗅阀浓度元有害作用浓
化合物 氯苯 1.2-二氯苯 1.3- 二氯苯
(微克/升)
(傲克/升)
(微克/升)度(微克/升}
10 10 10
10 1 0.1
20'" 100 2'" 10 20 0.3"'30 10 5"'30 50 20 400 130
5"'50 5"'50 5"'50 5-50
1. ←二氯苯 1.2.3- 三氯苯
‘10
1.2.4- 三氯苯
1. 3. 5-:?' 氯苯 1.2.3.4- 四氯苯 1. 2.3.5- 四氯苯
1.2.4.5- 四氯苯
* * * * * 2-20
一无资料，表明浓度可能低于 0.1 微克/升。 *无慢性毒性资料。
对二氯苯而言，嗅阔稍低于根据毒理学资料确定的限值。
5.1 5. 1. 1 概述
氯苯(一氯苯)a
氯苯(一氯苯〉广泛用作溶剂和生产染料、农药和其他 化学物质的中间体(1)。它也可在水加氯消毒中生成。
5. 1. 2 5. 1. 2.1
接触途径 水
据报道，氯苯在地下水、地面水和饮水中的浓度为 ‘关于一氯苯的参考文献与二氯苯的参考文献→并列于第 5.2 节后面。
0.005~10 微克/升。， 2) 。假设水的摄入量为 2 升/天，吸收
效率为 100% ，则每日摄入量波动在 0.01~20 微克之间。
5. 1. 2.2
空气
尚未见到在空气中检出该化合物的报道，经空气接触氯 苯的唯一资料来源于在工作环境的职业接触。所报道的职业 接触浓度波动于 O.004~0.3 毫克/升之间。2.
5. 1. 2.3
食物
没有关于从食物中摄入氯苯的资料.
1. 1. 3
代蹦
在哺乳动物体内氯苯可能经环氧化的方式被代谢为二盼
衍生物，继而，这些氯酣以结合物的形式被排出，氯苯也可 能被代谢为 p-氯苯硫醇尿酸 (2)。
5. 1. 4 5. 1. 4.1 剂 E130
对健康的影响 对人体的观察
氯苯剌激呼吸系统，而且是一种中枢神经系统的抑制
巳报道了几例由于吸入氯苯而发生的中毒。但是几乎没 有有价值的关于人体单纯接触氯苯的资料，只有接触氯苯合 并接触其他物质的资料。
5. 1. 4.2
对其他种属动物的观察
虽然缺乏可接受的慢性毒性研究资料，但是有足够证据
表明氯苯对靶器官有毒性，而且这种作用与剂量相关。 在狗的三个月实验中揭示，无有害作用水平为 27.25 毫
克/公斤体重/天。在大鼠的研究中表明，给动物剂量为 12.5
和 14.5 毫克/公斤体重/天，分别达三和六个月，结果未见有 巧，" ，也
害作用。对大鼠的另一项研究表明，给予动物剂量为 0.001
毫克/公斤体重/天共七个月，未观察到奇害作用。30
5. 1. 4.3
致突变性 致畸性 致癌性
无资料
5. 1. 4.4 5. 1. 4.5 5. 1. 4.6
无资料 在文献中没有资料表明一氯苯是致癌的。 建议值
在大鼠实验中未检出有害作用的剂量范围为 0.001 ~ 14.5 毫克/公斤体重/天。选择从大鼠短期实验研究中得出的 较高的无有害作用剂量为 14.5 毫克/公斤体重/天，应用较大
的安全系数 1000"'10000 ，则可获得暂行日许量为 0.0015'" 0.015 毫克/公斤体重。对于一个 70 公斤体重的人，该值
代表摄入量 0.1"'1 毫克/天，将该值的 10% 限定为从水中摄 入的量，则在饮水中的暂行毒理学限值为 5-:---50 微克/升二 因为在水中一氯苯的嗅阔浓度为 30 微克/升，该值接近 于根据对健康的影响而计算的数值，所以推荐一氯苯在饮7其 中的建议值为 10% 的嗅阔值，即 3 微克/升。
5.2 5.2.1 概述
二氯苯
二氯苯是一组含有三个异构体的卤代芳香族化合物。 1 ， 2-和 1 ， 3-二氯苯在常温下为液体， 1 ， 4-二氯苯为固体，它
们均具有相当的挥发性。 1 ， 2-二氯苯的主要用途是作为生产甲苯二异氟酸盐的加 -325 一
工溶剂以及合成染料、除草剂和脱脂剂的中间体，但是主要
是作为生产农药的中间体。 1 ， 4-二氯苯主要用作空气除臭剂、 农药和驱蛙虫剂。 )0 1 ， 3- 二氯苯可能是作为 1 ， 2-或 1 ， 4-二
氯苯的一种污染物而产生的，但是关于其商业生产和用途尚 无资料。 1 ， 2 和 1 ， 4-二氯苯几乎完全是生产一氯苯的副产品。
5.2.2
接触途径
二氯苯的生产、使用、运输和废弃已经引起了广泛的扩 散和对环境的污染，在河流、地下水、城市和工业排放物、
饮水、空气和土壤中均巳被检出。 5.2.2.1 水
1 ， 2 和 1 ， 4-二氯苯常见于处理前的自来水水源中，浓度 为 1.......10 微克/升。在饮水中的浓度为 0.001.......1 微克/升 (5). 。
假定每人水的摄入量为 2 升/天，而且吸收率为 100% ， 则每人从水中的摄入量波动在 6 X 10- 8 微克(在饮水中的中
值为 3 毫微克/升〉至 6 微克(报道的总二氯苯最高为 3 微 克/升〉。
5.2.2.2
空气
关于二氯苯造成空气污染的资料很有限。在加利福尼亚 室外空气中测定的结果表明， 1 ， 2- 二氯苯的浓度波动在 0.002.......50 毫克/立方米，未检出 1 ， 4- 二氯苯。在东京住宅
内 1 ， 4- 二氯苯的浓度高得多，其范围为 105 微克/立方米 (卧室〉至 1700 微克/立方米〈贮藏室〉。假定每天的呼吸 量为 20 立方米空气〈成;年男人) ，吸收率为吸入量的 50% ， 那么每天的摄入量波动在 0.02 毫克〈根据已报道的最低郊 'Campbell , 1. 1.饮水中二氯苯的最大容许限值。美国环保局(为世界卫 生组织准备的未发表的文件， 一 326 一
1980).
区浓度计算)至 20 毫克(根据由于使用 1 ， 4-二氯苯所造成
在贮藏室空气中的浓度计算〉。
5.2.2.3
食物
食物也可能受到二氯苯的污染。由于在动物呼吸的空气
中存在着 1 ， 4-二氯苯，致使猪肉受到污染，据报道母鸡接触
空气中 1 ， 4-二氯苯的浓度为 20"""38 毫克/立方米曾导致蛋 的污染。在日本沿岸水域的鱼中也检出 1 ， 4-二氯苯。尚无资
料可供估计由于摄入食物而造成对二氯苯类的特殊接触。
5.2.3
代甜
二氯苯可通过肺、胃肠道和完整的皮肤吸收。因卤苯的 水溶性低而脂溶性高，故易于通过扩散而穿透大多数的细胞 膜，包括肺和胃肠道上皮、脑、肝实质、肾小管和胎盘。 经胃管给灰兔一次剂量的 1 ， 2-二氯苯和 1 ， 4-二氯苯，结 果表明 1 ， 2-二氯苯主要通过氧化被代谢为 3 ， 4-二氯酷，并且
与葡萄糖酷酸和硫酸结合，主要从尿中被排出。 1 ， 4-二氯苯 主要经过氧化被代谢为 2 ， 5-二氯酷，并且仅仅以葡萄糖苦酸
和硫酸乙醋的形式被排出。在人体中也表明， 2 ， 5- 二氯酣是
1 ， 4-二氯苯的主要代谢物。
5.2.4 5.2.4.1
对健康的影响
对人体的观察
自 1939 年以来大多数已报道的人体二氧苯中毒的病例
(
22 例中的 16 例〉主要是由于长期吸入其蒸气所致。而有
些是由于食入 (22 例中 3 例〉和经皮吸收 (22 例中 3 例〉造 成的。大部分中毒是由于职业性的接触，但是有一些是由于
在家中使用或误用二氯苯制品而发生的。在多数的病例报告 -321-嗣
(22 例中的 15 例〉中，主要是接触含 1 ， 4-二氯苯的制剂， 其余病例主要是接触]， 2-二氯苯;还有几例是接触包括 1 ， 3一
二氯苯在内的混合物。靶器官系统或组织是一个或一个以上
的下列器官 z 肝、血液〈或网状内皮系统，包括骨髓和/或 免疫成分〉、中枢神经系统、呼吸系统和包膜。在这些病例 报告中的临床表现表明，二氯苯影响着广泛的靶器官。例如，
在 22 例已报道的临床病例中，至少 17 例出现一般的毒性或
兴奋症状〈例如疲劳或衰弱、厌食、体重降低、恶心、头痛、 兴奋或不适)， 17 例表现的症状和体症涉及到循环系统，包括
血液和/或骨髓或其他网状内皮成分〈例如贫血、白血病、自 细胞减少或自细胞增多、多核自细胞增多、骨髓增生、成自
细胞增多、出血倾向、脾大和黄瘟〉。
5.2.4.2
对其他动物种属的观察
给 10 只豚鼠灌胃一次剂量 (800 毫克/公斤体重〉的 1 ， 2二氯苯〈以 50% 浓度溶于橄榄油中〉引起动物体重降低，但
是所有动物均存活。剂量为 2000 毫克/公斤体重时全部受试 动物死亡。将 1 ， 2-二氯苯 i容于经阿拉伯树胶乳化的橄榄油 中，以胃管给各组大白鼠分别灌入剂量为 18.8 、 188 和 276
毫克/公斤体重/天的二氯苯，每周五次，在 192 天中总共给 予 138 次剂量，结果在高剂量组的阳性所见为肝和肾重量增
加、脾重量降低以及显微镜检查出现轻度至中度的肝浊肿。
在中剂量组，动物的肝和肾重量轻度增加。在低剂量组未见 有害作用。将两滴未稀释的 1 ， 2- 二氯苯滴入家兔眼中引起疼 痛和结合膜充血，但在一周内完全恢复。
在一项慢性毒性试验中，给大鼠 1 ， 2- 二氯苯的每日剂量 分别为 0.001 、 0.01 和 0.1 毫克/公斤体重，结果 9 个月后
高剂量组动物出现中枢神经系统高级皮质功能紊乱、血红蛋 一 328 一
自降低、血小板增多、嗜中性白细胞减少和骨髓有丝分裂活
性受到抑制。 0.1 毫克/公斤体重组的剂量为"阔剂量"，低 剂量 (0.001 毫克/公斤〉组的剂量为"阔下剂量"。 用胃管给雌性大鼠灌入 1 ， 4-二氯苯的油 j容液(用阿拉伯
树胶乳化)，每周 5 次，在 192 天中共给予 138 次剂量。结
果在高剂量组( 360 毫克/公斤体重/天〉动物出现肝和肾重 量增加、肝硬变和局灶性坏死。在中剂量组(1 88 毫克/公斤
体重/天〉观察到肝和肾重量增加。在低剂量组(1 8.8 毫克/ 公斤体重/天〉未见有害作用。
5.2.4.3 5.2.4.4 5.2.4.5
致突变性
现有资料不足以得出结论。 致畸性 致癌性 无资料
还没有关于二氯苯对动物特殊致癌试验的报告以及人的 有关流行病学研究资料。 虽然没有关于二氯苯致癌性的直接布力证据，但是有一
些资料表明，在掌握较足够的资料之前应将其视为可疑致癌 物。现有大鼠的资料未证实 1 ， 4-二氯苯的致癌性。也没有关
于 1 ， 2-二氯苯的充分资料。流行病学研究未能提供充分资科 以评价二氯苯对人体的致癌性 (6)。
5.2.4.6
建议值 ,...."
巳发表的资料表明，无有害作用剂量的范围为 0.001 同的结果。
13.4 毫克/公斤体重/天。因此使用不同数值可以得出很不相
使用从大鼠短期实验中所获得的未检出奇害作用的剂量
为 13.4 毫克/公斤体重/天，和安全系数 1000"""'10000 ，经计 -329 一
算得到 1 ， 2 和 1 ， 4- 二氯苯的暂行日许量为 0.00134 "-'0.0134 毫克/公斤体重。对于一个 70 公斤体重的成人而言，这意味
着摄入量为 0.1 ......， 1 毫克/天，假定该剂量的 10% 来自水的摄 入，而且每天水的摄入量为 2 升，则依据毒理学资料其在饮 水中的暂行限值为 5"""'50 微克/升。
该值超过嗅阔值， 1 ， 2-二氯苯的嗅阔值约为 3 微克/升。 故推荐 1 ， 2- 二氯苯嗅阔值的 10% 为合理的建议值，即 0.3 微克/升。同样， 1 ， 4-二氯苯的嗅阑值仅为 1 微克/升，其建
议值为 0.1 微克/升。 参考文献
1. Toxicological appraisal
0/ halogenated aromatic
compounds /o l/ owing groundwater pollution. Copenhagen , WHO Regional Office for Europe , 1980.
2. Ambient water quality criteria /or chlorinated benezenes. Washington , DC , US Environmental Protection Agency , 1979 (EPA 440/5-80-028). 3. Van Gemert , J. L. & Nettenbreyer , A. H. Compilation
0/ odour threshold ualues in air and water. Leidschendam. The Netherlands. National Institute for Water Supply. 1977 (2 :l 60 AD). 4. Zoeteman. B. C. J. Sensor y assessment
0/
water
qua/ity. Oxford. Pergamon Press , 1980. 5. Ambient water qualit y criteri a /or d ichlorobenzenes. Washington. DC. US Environmental Protection Agency ,
1979 (440/5-80-039).
6. Some industrial chemicals and dyestu//. Lyon , International Agency for Research on Cancer. 1982 (l ARC Monographs on the evaluation of the
carcinogenic risk of chemicals to humans. vo I. 29).
6. 苯和低级烧基苯 苯和低级皖基苯〈例如甲苯和乙苯〉被广泛用于化学工
业，作为生产各种化学物质(诸如盼和环己烧〉的中间体。 低级皖基苯是汽油的成分，被用{乍油漆和涂料的溶剂。苯和 低级皖基苯在地下水中的浓度通常高于地面水，这是由于在
地下水中妨碍了蒸发过程所致。现已发现因为化学废弃物的 渗漏和堆放，致使这些化学物质在地下水中的浓度可达数毫 克/升。在饮水中一般不超过 1 微克/升。但是，除苯以外， 目前在饮水中已发现的皖基苯浓度似乎尚未掏成潜在的健康
危害。
6.1
概述
苯和甲苯主要来自石油加工以及生产煤气和焦炭的副产
品。大量被应用于化学工业，三种主要用途是生产苯乙烯、 对异丙基苯(用于生产酷和丙嗣〉和环己烧(用于生产尼龙)。
而大量的甲苯是用于生产苯。小量(但是数量仍然可观〉的
苯和甲苯被用于各种工业(例如塑料、油潘、洗涤剂和汽油 添加剂)，或者用于许多化学合成的中间体或者作为洛剂。
其他烧基苯〈乙苯、二甲苯)也被广泛用作海剂或化学中间 体。
6.2
接触途径
苯和低级皖基苯具有挥发性，在环境中比较不易发生反
应。由于空气的运动使其广泛分布在环境中，并且通过雨和 从水表面的蒸发使其在空气和水体之间不断地再循环。最终
可能因生物的和微生物的氧化而被降解。
6.2.1
空气
在城市空气中苯和甲苯的浓度通常为 100 微克/立方米 的水平，其主要来自与汽油有关的排放〈汽车废气、汽油加
工)，也来自于工业活动所造成的洛剂丢失和排放。对于世 界上很多城市空气中的平均浓度进行了汇总，结果表明苯的 浓度通常低于 90 微克/立方米 E13 ，甲苯为 112.5--150 微克/ 立方米 (2) 。根据计算，美国的平均接触水平为 9.5 微克/立方
米〈其范围从小于 3.5 微克/立方米至 1 毫克/立方米) (3)' 0
8.2.2
盒物
几乎没有关于苯或甲苯在食物中含量的资料，虽然一些
资料表明苯在天然情况下存在于某些水果、鱼、蔬菜、坚果、 乳制品、饮料和蛋中 (3) ，甲苯存在于邻近石油化工区的鱼 中(4)。
6.2.3
7](
水中苯和甲苯的主要来源是大气沉降〈通过雨和雪〉和 化学工厂排放的废水(少量来源于城市泾流和污水厂〉阳。
据报道，在化学工厂的废水中苯的浓度达 179 微克/升，而在 出厂的饮用水中通常低得多 F 在 4 个美国城市饮水中的浓度 为 0.1--0.3 微克/升 E63 ，而在加拿大的浓度低于 0.01 微克/ ·在原来的文献中，浓度以 ppb 表示，在此采用下列换算系数，
lppb 苯 ==3.0 微克/立方米 I lppb 甲苯 =3.75 徽克/立方米。 。。
升(1)。甲苯的浓度达 19 微克/升 (2) 。可以预料，挥发性碳氢
化合物，例如苯和甲苯，能迅速从水体挥发进入大气(苯租 甲苯在 25 0C 时的半减期分别为 37.3 和 30.6 分钟) (2) ，这一
事实可以解释为什么在地下水中的浓度有时比在地面水中的
浓度高得多。也一项有价值的观票是出厂水中的苯可能来源于 在水处理时所使用的无烟煤媳料 (7) 。
6.2.4
职业接触
在生产和使用苯的工厂空气中，苯的浓度通常比一般大 气环境中高 3 个数量级，其浓度范围一般为 0.3""-'9 毫克/立
方米(时间加权平均值) (1)。对于接触苯浓度为 3.0 毫克/立 方米(时间加权平均值〉的个人来说，每年接受的剂量约比 非职业接触者高 30 倍(1)。
6.2.5
对于人接触量的估计
接触本底浓度(例如，非职业接触〉的城市居民苯的总 吸收量估计约为 125 毫克/年，其中 90 毫克来自食物 (D 。但
是应该注意，关于苯在食物中含量水平的资料极为有限，该 本底值仅仅是近似的参考值，而且也并未考虑消费者使用含
苯产品的个别情况。然而确实表明了，与从食物和空气中的
摄入量比较，苯在饮水中的浓度通常很低，可能仅具有有限 的意义。同样的结论适用于甲苯。
6.3
代谢
苯在人和动物体内的代谢和排出途径是中日似的 (8)。不管
给药途径如何，主要是以未改变的苯从呼出气中排出。结合型
的代谢产物，典型的是大量的酷，伴随以小量的邻苯二酷、 -333 一
醒醇和是基酿醇，从尿中排泄。肝脏是氧化和结合作用的主
要部位。 甲苯被迅速和广泛地代谢为马尿酸，从尿中排出，给予 剂量的大部分在 12 小时内几乎全部以未改变的甲苯形式从 呼出气中排出或以马尿酸形式排出 (2) 。
6.4
对健康的影响
目前已有许多关于苯毒性的综述 (1 月叫1)。急性接触苯引
起中枢神经系统的抑制。虽然大多数关于苯毒性的研究工作 是采用吸入途径，但是有限的动物实验表明，其它中毒途径
也得到类似的结果。 慢性接触苯引起造血组织的改变，表现为贫血和白细胞 减少。流行病学研究和数个病例报告表明，接触苯和白血病
有关。根据这一证据，国际癌症研究机构工作组将苯归入人 体致癌物 (12) 。
动物实验表明，接触苯能影响免疫防御机制，在很多项 研究中还可能引起受试动物和培养细胞的染色体损伤。 急性吸入甲苯引起中枢神经系统抑制J (6) 。但是，未导致
不可逆性的组织损伤，并且认为其主要代谢物苯甲酸是相对
无毒的。虽然大多数的毒性研究采用吸入途径中毒，然而也 进行了大鼠的慢性经口实验。在最高剂量组，给予动物甲苯
的剂量为 590 毫克/公斤体重，每周 5 次，共持续 193 天，未观 察到有害作用 (5)。甲苯未呈现致畸性、致突变性或致癌性 (2) ,
但是在一项小鼠的研究中报道了当使用大剂量时出现致畸作 用 (13)。在许多长期职业接触甲苯的研究中未检出血液系统
的改变或肝损伤 (6) 。上述资料表明，甲苯是相对低毒的，经
计算水中苯的最高容许浓度为 14.3 毫克/升 E230 鉴于该值远
高于在水中所检出的法度，所以决定不推荐甲苯的建议值。
而根据终生和在最近十年内增加癌症危险性为十万分之一是 可以接受的，推荐苯在饮水中的建议值为 10 微克/升。 参考文献 1. Hollid 町， M. et a l. Benzene. Human health
implications
0/
benzene at leuels /ound in the
Canadian environment and workplace. Ottawa , Health and We lfare Canada , 1978 (Environmental Health Directorate Report No. 79-EHD-40).
2. Ambient water quality criteria /or toluene. Washington , DC , US Environmental Protection Agency , 1980 (EPA-440/5-80-7 日.
3. Mara , S. J. & Lee , S. S. Assessment
0/
human
exposure to otmospheric benzene. Washington , DC , US Environmental Protection Agency , 1978 (EPA-450/ 3-78-03 1). 4. Ogata , M. & Miyake , Y. Identification of substances in petroleum causing objectionable odour in fish. Water
research , 7.1493 (1 973). 5. Wolfe , M. A. et a l. Toxicological studies of certain alky lated benzenes and benzene. Archives
0/
industrial health , 14'387 (1 956). 6. National Research Counci l. Drinking water and
health. Washington , DC , National Academy of Sciences. 1977. 7. Smillie , R. D. et a l. Low molecular weight hydrocar bons in drinking water. J ournal enui γ onmenfa! οf
health ,
A13 :t 87
(1 978).
ea pa
8. Rusch , G. M. et a I. Benzene metabolism in benzene
toxicity. a critical evaluation. Washington , DC , American Petroleum Institute , 1977 , pp 23-36. 9. Laskin , A. & Goldstein , B. D. , ed. Benzene toxicit y. a critical evaluation. Washington , DC , American Petroleum Institute , 1977 , 147 pp. 10. National Research Council H ealth ef fects Academy of Sciences , 1976 , 23 pp.
0/
benzene. a review. Washington , DC , National 11. Some antithyroid and related substances , nitrofurans and industrial chemicals. Lyon , International Agency for Research on Cancer , 1974 (I ARC Monographs on th <.l evaluation of the carcinogenic risk of chemicals to humans , vo I. 7).
12. Chemicals and industrial þrocesses associated with cancer in humans. Lyon , International Agency for Research on Cancer , 1979 (I ARC Monographs on the evaluation of carcinogenic risk of chemicals to man , Supp I. 1). 13. Nawrot , P. S. & Staples , R. E. Embryo-foetal toxicity and teratogenicity of benzene and toluene in the mouse. Teratology , 19:41A (1 979).
-336-
7. 盼和氯盼 氯盼用作杀虫剂，并且可能当含酷的水加氯消毒后被发
现。众所周知，氯盼的嗅初味阐浓度很低，大多数具有臭味 的氯酷化合物〈一氯酷和二氯盼〉的昧阔浓度低至 1 微克/ 升，因此基于感官影响的原因，单个的氯盼在饮水中的浓度 通常不应超过 0.1 微克/升，酷和五氯酣例外，其味阔浓度
为 100 微克/升。在不加氯消毒的情况下，盼在饮水中的浓 度可以达到 100 微克/升。 在不能符合建议值 0.1 微克/升的情况下，应该注意到 某些氯曲在稍高的浓度时具有毒作用。 不经特殊浓缩处理而用分光光度法测定酌时，仅能检出
浓度大于 1 微克/升的氯酌，该浓度高于数种氯酷的味阐浓
度。如果怀疑氯酣引起了昧的问题，就需要由一组人员进行 直接的味觉评价和用色谱法进行分析测定。
控制饮水免受氯盼污染的最好办法是防止水源受到酷和 氯代酌农药的污染。如果原水中含有高浓度的酌，就应在加
氯消毒前尽可能降低酷的浓度。一且在水中存在低级取代的 氯酌时，则需用氧化法去除 F 而如果存在高级取代的氯酣时，
就只有用活性炭才能将其有效地去除。
7.1 7. 1. 1 概述
具有毒理学意义的氯盼
已知氯盼存在于饮水中，这是由于原水水源受到污染或 一 337 一
含酣类化合物的水经加氯消毒而产生的。在原水中含有盼可 能来源于煤炭蒸锢厂、石油化工厂和酌作为中间体的其他工
厂排出的废水。酣也存在于城市废水中 o 当含酷的水加氯消 毒时，其主要反应产物为 2-氯盼、 4-氯酷、 2 ， 4一二氯酷和 2 ， 4 ， 6-三氯盼。 2 ， 4-;-二氯酣在商业上是作为生产除草剂 2 ， 4j商、有关杀虫剂和五氯盼的一种中间体而进行生产的。五氯 酣用作木材防腐剂 J 2 ， 4 ， 5-三氯酣是一种杀真菌剂 J 2, 4 , 6三氯酣是一种防腐剂，也是杀虫剂林丹的主要代谢物。 2 ， 3 ，
量， 6- 四氯盼主要用作杀虫剂和木材防腐剂. 7. 1. 2 存在
被污染的原水，包括地下水，可能含有 1--10 微克/升的
酣、一氯盼和二氯盼。在饮水中也报道了类似的浓度。三氯 酷和四氯盼在原水中检出的浓度为 1--10 微克/升，有时更高 二些，而在饮水中的浓度通常低 1--2 个数量级。鉴于其在
水中出现的频率很低，故除被列入表 3 的氯酣以外，不再进 一步考虑其他氯盼对公众健康的影响 (1 叫飞
7. 1. 3
根据健康影响对氯酣的初步筛选
根据在饮水中预计的最高浓度以及有关毒性和对感官影
响的现有文献资料，显然，有几种氯酣基于毒性确定的限值 远高于基于昧和嗅作用而确定的限值。为了能够主要基于毒 理学考虑选择化合物，现将有关资料汇总于表 3(5-7) 。如表
3 所示，对某些氯酣而言，基于毒性考虑所计算的限值远高 于感宫的考虑。而 2 ， 4 ， 6-三氯酣和五氯盼例外，是基于毒理
学资料计算得出的限值。 对于其他氯酷和酷本身的限值，建议根据感官考虑、来确 。。
定。 表 3 从毒理学考虑确定氯盼限值的基准资料 最高检出浓度 一一一一一一一二二嗅闽浓度·
基 J佳浓度 味阅浓度 a-→←-一←→一
原水饮水 化合物 盼
毒性敖癌性
(徽克/升) (徽克/升) (徽克/升) (微克/升) (徽克/升) (徽克/升)
100 10 10 10 10 <0.1 I
1000
100
3000
2-氯酷 j- 氯盼
2 ， 4-二氯盼
10 10 100 100
3000
2. 。二氯盼
2.4.5' 三氯酌
2600
一_b
2.4.6- =:::氯i[，'} 2.3.4.6- 四氯盼 五氯盼 一资料不充分。
12
<0.1 10
<0.1
1000 1000 100 21
a 值。
在文献中所报道的阀值很不相同，巳有一些报道高于或低于所列出的数 2.3.7 ， 8- 四氯二苯 -p-二 R坚决 (TCDD) 是工业品 2.4.5-三氯盼她一种杂
b
质。在环境中它们的形为不同，因此应分别对待。
7.1.4
对感富的考虑
面临这样一个问题 z 即在几乎不知道某些氯百分毒性的情 况下，根据一氯酣和二氯盼的感宫阔浓度很低的事实是否可 以确保这种"无味饮水"的安全性，根据文献中已具有完善 记载的某些氯酷的毒理学资料，无臭和无昧的水不可能因氯
盼而造成对健康的任何直接危害，但是必须认识到，一般来 说，没有任何不良的臭和味并不能确保饮水的安全。当饮水 具有明显的味和臭时，应该调查是否可能存在着氯酷类化合
物。 一 339 ~
7.2 7.2.1 概述
2 ， 4 ， 6-三氯盼
2 ， 4 ， 6-三氯盼是一种黄色固体，熔点 69 .5 0C ，沸点
246 "C，微溶于7K(低于 0.1 克 /100 毫升) (8) ，而j容于有机溶剂。 它用作木材防腐剂、杀细菌剂和杀真菌剂。 2 ， 4 ， 5-三氯 盼是在含低浓度酷的水加氯消毒时生成的产物之一。
7.2.2 7.2.2.1
接触途径 1
水
2 ， 4 ， 6-三氯酣在河水中的检出浓度达 1 微克/升。 1978 年在荷兰莱茵河中的浓度范围为 0.04--0.63 微克/升，
Wegman a 还发现含有类似浓度的 2 ， 4 ， 5-三氯酷。在饮水中 的浓度通常较低，但资料很少。当含盼的水加氯时，可导致 2 ， 4 ， 6一三氯盼的浓度达到数微克/升的水平。估计一个 70 公
斤体重的成人从含 2 ， 4 ， 6-三氯盼浓度为 1 微克/升的 2 升水 中每天最大的摄入量为 0.00003 毫克/公斤体重。
7.2.2.2
食物
另一个摄入途径可能来源于乳制品，这是由于在乳制品 工业中广泛使用氯酷类杀菌剂的缘故 (9)。已表明 1 ， 3 ， 5-三 氯苯被代谢为 2 ， 4 ，←三氯酣 (10)。五氯环己烧在玉米和豌豆 属植物中也被转化为 2 ， 4 ， 6一三氯盼(11)。还可能由于食用鱼
和贝类而摄入。但是没有关于吸入接触的定量资料。目前由 于缺乏其他接触途径的定量资料，所以假定水是主要的摄入 • Wegman , R. C. 兰， 1980. C. 固立公共卫生研究所未发表资料，
Bilthoven ,
荷
一 340 一
途径。
7.2.3
代谢
2 ， 4 ， 6-三氯酣主要经尿被迅速从体内排出 (12) 。而关于
其代谢的进一步资料了解甚少。
7.2.4
对健康的影响
2 ， 4 ， 6-三氯盼的橄榄油溶液对大鼠腹腔注射的半数致死 量为 276 毫克/公斤体重 (13) 。与其他氯盼一样， 2 ， 4 ， 6-三氯
酣能使体温增加，并且在高剂量时能引起动物抽擂。
7.2.4.1
致突变性
2 ， 4 ， 6-三氯酣能增加啤酒酵母的突变率 E143，但是在沙
门氏菌/哺乳动物微粒体测试系统的 Ames 试验中，无论 代谢活化系统存在与否，均未见其致突变性 (15) 。
7.2.4.2
致癌性
美国国立肿瘤研究所 (16) 在 F344 大鼠和 B6C3-F1 小
鼠的实验中发现，该化合物引起肿瘤发生率增高。在雄性大
鼠的实验中观察到淋巴瘤和白血病发生率增高，并且其发生 率与剂量相关。在雄性和雌性大鼠中出现外周血白细胞增多
和单核细胞增多以及骨髓增生。在雄性和雌性小鼠中发现肝 细胞癌和腺瘤发生率明显增高，并与剂量相关。
流行病学资料不足以评价 2 ， 4 ， 5-和 2 ， 4 ， 6一三氯盼的致 癌性 (8 ， 17) 。
假定终生引起癌症的危害性为 1/100000 ， 采用线性多阶 外推模式计算所得的建议值为 12 微克/升，而味阔浓度仅为
1 微克/升。为提供可以接受的，而且味道良好的饮用水，故 推荐其建议值为 0.1 微克/升. 一 341-
7.3 7.3.1 概述
五氯盼
五氯酣广泛用作杀真菌剂和木材防腐剂，估计世界产量 为 4 万吨/年。它的熔点 190 'C，沸点 310 'C， 20 'C时在水中
的海解度为 14 毫克/升。
7.3.2
接触途径
受污染的地面水含五氯盼的浓度可达 10 微克/升。在 1978 年荷兰的莱茵河中浓度范围为 0.15--- 1. 5 微克/升。在
饮水中一般远低于 1 微克/升。已报道在蠕虫中的浓度为 0.2 毫克/公斤体重，在哺乳动物中为 0.1--- 1. 3 毫克/公斤体重。
7.3.3
代谢
五氯盼可经胃肠道很好地被吸收，也可经皮吸收。被人
体吸收的大部分五氯盼是以未改变的形式从尿中排出。在大
鼠和小鼠体内大约全部剂量的 20% 经脱氯生成四氯氢醒和 三氯-1-氢醋。
7.3.4
对健康的影响
五氯酌对大鼠的急性经口半数致死量为 27 毫克/公斤体
重。中毒的临床表现为多汗、口渴、体温增高、脉搏和呼吸 加快以及最终心动停止。
据报道，在使用五氯盼的木材工人中出现数例何杰金氏 病和白血病 El83，但是流行病学调查不足以进行评价(1 9 ， 20) 。
至少已有 30 例五氯盼中毒死亡的病例报道。在家兔实验 ← 342 一
中观察到五氯酌引起氯座疮，还可引起实验动物的肾和肝损
伤，对职业接触者也出现类似表现.已经表明五氯酌对实验 动物具有胚胎毒性和胎儿毒性。在酵母试验中引起突变率的
增加，但是在 Ames 试验中未呈现致突变性。在动物实验 中未证实具有致性，但是应该认识到现有的研究尚不够充 分∞
美国国家研究委员会。1)已经计算了五氯盼的日许量为 3 微克/公斤体重/天。假定一个 70 公斤体重的人每日饮用 2 升水，而且将日许量的 10% 限定为饮水，则计算所得在饮 水中的建议值为 10 微克/升。 参考文献
1. Ambl'ent water quality criteria /or phenot. Washington , DC , US Environmental Protection Agency , 1980 (E P A 440/5-80-066). 2. Ambienl water quality criteria /or chlorinated
phenols. Washington , DC , US Environmental Protection Agency , 1980 (EPA 440/5-80-032).
3. Ambient water qual it y criteri a /or 2 , 4dichloropheno/. Washington , DC , US Environmental Protection Agency , 1980 (EPA 440/5-80-042).
4. A mbi ent water qualit y criteri a /or pentachlorophenol. Washington , DC , US Environmental Protection Agency , 1980 (E P A 440/5-80-06 日. 5. Van Gemert , L. J. & Nettenbreijer , A. H. Compilation
0/ odour threshold values Water Supply , 1977.
in air and water.
Leidschendam , The Netherlands , National Institute for 6. Zoeteman , B. C. J. Sensor y assessment
0/ water duz
，a 。
qualit y. Oxford , Pergamon Press , 1980. 7. Dietz , F. and Traud , J. Geruchs- und GeschmacksSchwellenkonzentrationen von Phenolkorpern. GW F-
Wasser/abwasser , 119:H6 318 (1 978). 8. Some halogenated hydrocarbons. Lyon , International Agency for Research on Cancer , 1979 (l ARC Monographs on the evaluation of the carcinogenic risk 。f
chemicals to humans. vo 1. 20).
9. Stannard, D. J. & Scotter , A. The determination of phenol residues in dairy products. N ew Zealand
journal (1 97 7) •
01
dairy science and technology , 12:140
10. Kohli , J. et a 1. The metabolism of higher chlorinated benzene isomers. Canadian journal
01
biochemistry ,
54: 203 (1 976). 11. Moza , P. et a 1. Beitrage zur okologischen chemie LXXXIX Orientierende versuche zum metabolisms Von-pentachlorocyklohex-l-en in hoheren pflanzen in hydrokultur. Chemosphere , 6:255 (1 974).
12. Korte , F. et a 1. Ecotoxicologic profile analysis , a concept for establishing ecotoxicologic priority list for chemicals. Chemosphere , 7:79 (1 978).
13. F. arquharson , M. E. et a 1. The biological action of chlorophenols. British journal
01
pharmacology.
13120 (1 958). 14. Fahrig , R. et a 1. Genetic activity of chlorophenols and chlorophenol impurities. In , Rao , K. R. , ed. ,
Pentachlorophenol , chemistry , pharmacology and enuironmental toxicology. New York , Plenum Press , 1978. -344-
15. Rasanen , L. et a l. The mutagenicity of MCPA and its soil metabolite8 , chlorinated phenols , catechols and some widely used slimicides in Finland. Bull etin (1 977) .
0/
environmental contamination and toxico!ogy , 18:565 16. National Cancer Institute. 1979 Bioassay
0/
2, 4, 6-trichlorophenol /or possi bl e carcinogenicit y. Washington , DC , US Department of Health , Education and Welfare , 1979 (Technical Service Repor t Series, No. 155).
17. Theiss , J. C. et a l. In. Long-term hazards
01
pol ychlorinated d ibenzodioxins and pol ychlorinated dibenzo/urans. Lyon , International Agency for Research on Cancer 1978 (IARC Internal Technical Report No. 78/00 1).
18. Green , M. H. Familial and sporadic Hodgkin's disease associated with occupational wood exposure. Lancet.
2'626 (1 978). 19. Hardell , L. & Sandström , A. Case-control studYI Soft tissue aarcomas and exposure to phenoxyacetic acids or chlorophenols. British journal
0/
cancer ,
39:711 (1 979). 20. Eriksson , M. et a I. Soft-tissue sarcomas and exposure to chemical 5ubstancesl a case reference study
ind ustri al med icine , 38: 27 (1 981). 2 1. National Research Counci I. Drinking water and health. Washington , DC , National Academy of British journal
0/
Sciences, 1977.
-345 一
8. 兰卤甲烧 饮水中的三卤甲烧主要是用于对水进行氧化处理的化学 物质与存在于水中的天然布机物反应的产物，它的形成与氯
的使用特别有关。四种最常检出的三卤甲炕是氯仿、一澳二 氯甲;皖、二澳一氯甲:院和澳仿，这四种化合物在饮水中的总
浓度可能高达 1000 微克/升，但通常低于 100 微克/升。 已表明，氯仿可诱发两种实验动物癌症。但是目前才开
始对三种含误的三卤甲烧进行终生致癌实验。该实验类似于 在确定氯仿为致癌物时的实验条件。巳知在 Ames 沙门民
菌致突变试验中，这些三卤甲炕比氯仿活性更强。 重要的是认识到氯是一种有效的饮水消毒剂，显然，由
于消毒不完全会造成微生物引起的疾病危害。特别在发展中 国家尤其如此，估计每天因介水疾病而造成数千人死亡.而 氯是最方便和易于掌握的消毒剂，所以已被广泛使用。
8.1
概述
三卤甲;境是被卤素取代的，含→个碳原子的化合物，其
通式为 CHX 3 ， X 可能是氟、氯、澳和腆，或者是它们的混 合。就饮用水的污染而言，把讨论局限于该组中的四种化合 物是合宜的 s 即氯仿 (CHC1 3 ) 、一澳二氯甲屁 (CHBrCl z ) 、 二摸一氯甲炕 (CHBrzC l)和提仿 (CHBr 3 ) 。其中氯仿最 常见，现有资料几乎全部集中于它。
氯仿的最重要用途是作为生产一氯二氟甲烧的原料，后 -346 一
者用作致冷剂、气溶胶推进剂和合成聚四氟乙烯。氯仿还是
一种重要的洛剂和脱脂剂，也小量用于麻醉剂、擦剂、烫发 液、洗牙水、熏蒸剂以及止咳制剂的活性成分和防腐剂。， 2) 。
估计全世界 1973 年的产量为 245 X 10 8 公斤 E370 澳仿在工业上用作计量液 (gauge fluid) 、固体分离 的重液和合成其他化学物的中间体(1)。对于混合的卤素衍生 物还不知道其在商业上的应用。
8.2 8.2.1
接触途径
水
水中的三卤甲;境主要来自于氯(以及偶然存在的澳离子) 和天然存在的有机化合物之间的反应。许多国家的资料 (4-6) 表明，在经氯化消毒的出广水中三卤甲妮的浓度通常比原水 高得多，在原水中常常来检出。
表 4 列入了美国对自来水的调查结果，其他调查也得到 类似结果。
表 4
美国出厂水中三卤甲烧的含量 度范围 浓克 浓度中值 (毫克/升)
化合物
测定点数
(毫/升)
氯仿 一澳二氯甲烧 一氯二澳甲烧
80 78 72 26
0.0001~0.311
0.021 0.006 0.0012 <0.005
O. 0003~0.116 0.0004~O.110
澳仿
0.0008~0.092
研究表明，对于一定剂量的氯而言，其三卤甲;境形成的 一 347
一
速度(程度)随腐殖酸浓度、温度和 pH 的增高而增加 E630
已经表明，如果在配水系统中存在游离余氯，则三卤甲 烧可继续生成 (7) 。
在文献中尚未见到有关一澳二氯甲皖、一氯二澳甲;院和 澳仿生成的资料。
8.2.2
空气
已检出在农村大气中氯仿的浓度范围为 100 ...._， 180 毫微 克/立方米 (8)。在任何地点大气中氯仿的最高浓度范围为低 于 0.05 ...._， 73.5 微克/立方米，中值为 0.045 ...._， 5.0 微克/立方
米 (9)。在北半球和南半球采样点的地面水平的大气中氯仿 的浓度分别为 130 毫微克/立方米和低于 15 毫微克/立方 米(1 0) 。
8.2.3
食物
在 17 "C和 30 0C 晾干后，用含氯仿的气体熏蒸剂处理的 两个大麦样品，其氯仿的最初残留量分别为 123 和 132 毫克/
公斤。 60 天后，在 30 "C晾干的样品，已无氯仿残留，而在
17 "C晾干的样品中，氯仿残留量仍为 16 毫克/公斤。对于玉 米和高粱亦得到类似的结果。
1973 年在英国的食物中发现氯仿的浓度如下z 奶制品 1. 4 ...._， 33 毫克/公斤 F 肉 1 ...._， 4 毫克/公斤 F 油和脂肪 2---10 毫 克/公斤，饮料 0.4 ...._， 18 毫克/公斤，水果和蔬菜 2 ...._， 18 毫克/ 公斤(1 1) 。
8.3
代谢
接触含氯仿浓度为 13. Z""_'3 1. 8 克/立方米的空气 3---10 一 348-
分钟，可吸收 73% (12) 。
吸入的氯仿迅速进入血流并被转运至组织。已发现小鼠
的体脂是氯仿最重要的贮存部位 F 其次是脑、肺、肾、肌肉 和血液。氯仿的代谢在肝脏进行。全身放射自显影的研究表 明，放射活性逐渐从脂库转移至肝脏 (13) 。由于在脐带血中
氯仿的浓度高于母血，所以看来氯仿能穿过人体的胎盘屏 障 (14) 。
在人体，经口剂量 (7 毫克/公斤体重〉的 50.6% 被代 谢为二氧化碳，但是个体问差异甚大，而食入量的 68.3% 是 以原形被呼出 EI530 相当一部分被吸收的氯仿当其首次通过
肝和肺时被排出，而仅仅给予剂量 500 毫克的 50""-'65% 通过 一般循环到达身体的其他部分(1 6)。志愿者食入一次剂量 500
毫克的氯仿，在 8 小时内以原形呼出 18""-'67% (1 5) 。氯仿的
主要代谢物是通过肺〈以二氧化碳形式〉或肾〈以无机氯化 物形式〉被排出 E173.
8.4 8.4.1 毒性
对健康的影响
关于三卤甲烧对健康影响的讨论主要涉及到氯仿，由于 它过去被用作吸入麻醉剂以及在饮用水中它是三卤甲:境的主 要形式，因而曾对其进行了最广泛的研究。而其他三卤甲烧
的毒作用可能与氯仿类似。 氯仿是中枢神经系统的抑制剂，也影响肝和肾的功能。 它中毒的即刻作用是意识丧失，继而可能昏迷和死亡时， 19)。
接触后 24""-'48 小时见到肾损伤，肝损伤见于接触后 2........õ 天 s 因此，氯仿麻醉恢复后数天可能发生中毒症状(1 8) 。 -349 一电
认为对于一个 70 公斤体重的人，氯仿的致死中量约为 44 克或 630 毫克/公斤，但是，有人摄入 250 多克氯仿仍然 存活 (20) 。已发表的对人最低致死量为 210 毫克/公斤体 重 E213，摄入低至 440 毫克氯仿引起对胃的刺激和肠道蠕动增 加以及局部麻痹 (19) 。
除对氯仿致癌性的研究以外，几乎未研究真慢性毒性。
长期经口给予剂量为 0.4 毫克/公斤体重的氯仿，结果未引起
大自鼠出现任何观察指标的改变，对豚鼠的唯一作用是使肾 上腺维生素 C 量增加 (22) 。在两项长期的研究(包括 229 名
人体实验者)中，评价了将氯仿加入牙膏和嗽口水中的安全 性 (23) 。估计在 1~5 年多的时间内氯仿的每日摄入量为
0.34~0.96 毫克/公斤体重。该项研究结果表明，依据肝功
能试验未见到对肝脏毒性。一位 47 岁的人每天服用 12~20 盎司 (336~560 克〉含氧仿的镇咳剂达 10 年之久，所观察 到的唯一作用是可逆性的肝脏毒性。估计每日氯仿的剂量为 23~37 毫克/公斤体重 (24) 。
认为滇仿引起的毒性症状类似于氯仿。据报道，皮下给 小鼠澳仿的半数致死量为 1820 毫克/公斤体重 (25) 。而几乎
没有关于其他三卤甲院毒性作用的资料。巳确定一澳二氯甲
院对 A/st 雄性小鼠的最大耐受量是 100 毫克/公斤体重，该 剂量是在两周内经腹腔注射了 6 次。用同样方法得到澳仿的 最大耐受剂量也是 100 毫克/公斤体重 E2630
8.4.2
.癌性的研究
美国国立肿瘤研究所进行了一项研究 α73，将溶于玉米
油中的氯仿用管词洁给 Osborne-Mendel 大鼠和 B6C3-
F1 小鼠，分两个剂量水平，每周给予 5 次.给予雄性大鼠
90 或 180 毫克/公斤体重两个剂量水平共 78 周，雌性大鼠 在最初 22 周接受的剂量为 125 或 250 毫克/公斤，随后的剂 量与雄性式鼠相同， 111 周后处死动物，发现雄性动物肾上 皮肿瘤发生率具有统计学意义的明显增加(高剂量组为 24% ，低剂量组为 8%) 。但是未见到雌性动物的这种改变，
雌性大鼠甲状腺肿瘤发生率增加，然而被认为无显著意义。 雄性小鼠最初接受的剂量为 100 或 200 毫克/公斤体重，雌性 小鼠为 200 或 400 毫克/公斤体重， 18 周后其剂量分别改为
150 和 300 毫克/公斤体重和 250 和 500 毫克/公斤体重。结 果两性小鼠的肝细胞癌发生率非常明显的增加 z 在高剂量
组，其发生率对雄性小鼠为 98% ，对雌性小鼠为 95% ，而
低剂量组分别为 36% 和 80% 。在低剂量组未发生肝细胞癌 的雄性小鼠常常出现结节性增生。但是应该强调，美国国立
肿瘤研究所使用的剂量极高，动物体重降低 10% 以上，因此 可以认为该剂量高于真正的最大耐受量。
8.4.3
流行病学研究
Cantor 等 (28)观察了 16 种不同的癌症死亡率与饮水中
三卤甲庇浓度的关系，并且对氯仿和非氯仿成分分别进行研 究。有关接触的资料来自于 1975 年国家有机物调查和环保 局 V 区的调查.在研究中包括了 76 个县，其中 50% 以上的
人口供给经过测定的给水。最一致的发现是膀眈癌死亡率与 三卤甲烧浓度水平之间的相关。在两种性别中均观察到这种 相关，并且其相关的强度与所测定的给水供应的人口百分数
成比例。还注意到澳化三卤甲烧时的这种相关较氯仿强些。 但是，在 13 项流行病研究的综述中，国家科学院安全饮水 委员会的结论是 s 从巳报告的研究中不能得出因果关系 (29)。
-351-
Hogan 等 (30)使用同样的资料和应用各种统计学方法
以确定统计学模式的适用性。当使用加权回归方法进行分析
时，其结果类似于原来的研究。直肠癌和膀眈癌死亡率和饮 水中氯仿的浓度呈正相关。
8.4.4
建议值
氯仿对人体健康具有数种有害作用。但由于尚未进行充 分的定量研究，故对人体中直接观察到的大多数作用难以估 计摄入量的安全水平。就接近于饮水中所观察到的浓度时所
可能产生的潜在健康危害而言，最严重的是在实验动物中所
见到的致癌作用以及从饮水中接触较高浓度三卤甲;院的人所 出现的类似作用。
根据国立肿瘤研究所选用大鼠对氯仿的生物试验结果，
采用多阶线性外推模式估计了氯仿的安全水平。大鼠的资料 比 B6C3-F1 小鼠的资料可取，因为当使用肝脏毒物，如氯
仿时，对在小鼠这种动物模型中引起肝脏肿瘤的机制具有明 显的疑问。在国立肿瘤研究所之后获得的资料清楚地表明， 小鼠对氯仿引起的肝损伤较大鼠敏感得多 E313，这些损伤是
在使用的剂量低于国立肿瘤研究所使用的剂量情况下发生
的，表明，这是一种外生性机制。对这种作用不存在外推模 式。此外，大鼠对氯仿的代谢速度比小鼠更接近于在人体中
所观察到的结果。
基于上述考虑，推荐在饮用水中氯仿的建议值为 30 微 克/升。在每天平均水的摄入量为 2 升的情况下，该浓度导致
每 10 万人口中终生增加的癌症病例少于一例。但是应该指 出，由于消毒不充分而造成的危害远远大于因氯仿明显超过
建议值所带来的危害 o ，。
(秦伍慧
译〉
参考文献
1. Hardie , D. W. F. Chlor%rm. In: Kirk. R. E. & Othmer , D. T. , ed. Encyclopedia Publishers , 1964. p. 119. 2. l'\..t tional
0/
chemical
technology.2nd ed. Vo I. 5. New York , Interscience Institute for Occupational Safety and Health.
C 俨 iteri a
/or a recommend ed stand ard . Occupational
exposure to chlor%rm. Washington , DC , US Department of Health , Education and Welfare , 1974 , p. 75-114. 3. Pearson , C. R. & McConne lI, G. Chlorinated C, and Cz hydrocarbons in the marine environment.
Proceedings
0/
the Royal Society. Series B , 189:
305-332 (1 975). 4. Rook , J. J. Formation of haloforms during chlorination of natural waters. Water treament and examination , 23:234 (1 974). 5. Rook , J. J. Haloforms in drinking water. Journal
0/
the American Water Works Association , 68:186 (1 976) .
6. Stevens , A. A. et a l. Chlorination of organics in drinking water. In , Jolley , R.
L.,
ed. , Water
chlorination. Envi ronmental impact and health e//ects , Ann Arbor , MI , Ann Arbor Science , 1975. 7. Health & Welfare , Canada , National survey /or
halomethanes in d rinking water. Ottawa , Health & Welfare , Canada , 1977 (7-EHD-9). 8. Russell , J. W. & Shadoff , L. A. The S'
ampling and nd
determination of halocarbons in ambient air using concentration on porous polymer. J ournal 01
chromatology. 134:375 (1 97 7). 9. Lillian , D. et a l. Atmospheric fates of halogenated compounds. Environmental science and technology , 9:1042 (1975). 10. Cox , R. A. et a l. Photochemical oxidation of halocarbons in the troposphere. Atmospheric
environment , 10:305 (1976). 1 1. McConnell , G. et a l. Chlorinated hydrocarbons and the environment. Endeavour , 34:13 (1 975). 12. Lehmann , K. B. & Hasegawa , D. Studies of the absorption of chlorinated hydrocarbons in animals and humans. Archiv /ür Hygiene und Bakteriologie. 72:327 (1 910). 13. Cohen , E. N. & Hood , N. Application of low-temperature autoradiography to studies of the uptake and metabolism of volatile anesthetics in the mouse. I. Chloroform. Anesthesiology. 30:306 (1 969). 14. Dowty , B.
J. et al. The transplacental migration and
accumulation in bJood of volatile organic constituents.
15. Fry. R.
Pediatric research , 10:696 (1976). J. et a l. Pulmonary elimination of chloroform
and its metabolite in man. Archives internationales
de pharmacodynamie et de thérapie , 196:98 (1 972). 16. Chiou , W. L. Quantitation of hepatic and pulmonary first-pass effects and its implications in pharmacokinetic study. I. Pharmacokínetícs of chloroform in man. J ournal 0/ pharmacokinetics and biopharmaceutics. 3'193 (1 975).
17. Van Dyke , R. A. et a 1. A metabolism of volatile anesthestics. I. Conversion in vivo of several Il nesthetics
to uC0 2 and chloride. Biochemical
pharmacology , 13'1239 (1 964). 18. Whipple , G. H.
& Sperry , J. A. Chloroform 0
poisoning-liver necrosis and repair. Bulletin
f t he
Johns Hopkins University , 20'278 (1 909). 19. Secher , O. Physical and chemical data on anaesthetics.
Acta anaesthesiologica scandinavica , 42 (Supp l.), 1 (1 97 1) .
20. Gosselin , R. E. et aJ. C /i nical toxicology of
commercial products. 4th ed. Baltimore , MD , The Williams and Wilkins Co. , 1976. 2 1. Dreisbach , R. H. Handbook of poisoning. Los Altos , CA , Lange Medical Publications , 1974. p. 275. 22. MikIashevskii , V. E. et a I. Toxicity of chloroform administered perorally. Gigiena i sanitarija , 31 :320 (1 966) •
23. De Salva , S. et a I. Long-term safety studies of a chIoroform containing dentifrice and mouth-rinse in man. Food and cosmetics toxicology , 13:529 (1 975). 24. Wallace , C. J. Hepatitis and nephrosis due to cough syrup containing chloroform. California medicine. 73 =4 42 (1 950). 25. Kutob , S. D. & Plaa , G. L. A procedure for estimating the hepatotoxic potential of certain industrial soIvents. Toxicology and applied pharmacology , 4:354 (1 962). 26. Cardeihac , P. T. & Nair , K. P. C. Inhihition hy castration of aflatoxin induced hepatorna in carbon -355-
tetrachloride-treated rats. Toxicology and applied
pharmacology , 26:393
(1 973).
27. National Cancer Institute. Carcinogenesis bioassay
0/
chlo 俨%rm.
Bethesda , MD , National Cancer
Institute , 1976. 28. Cantor , K. P. Association of cancer mortality rates and trihalomethane level in municipal drinking water supplies. (Abstract) American journal
0/
epidemiology , 106:230 俨 egister ，
(1 977).
29. US Environmental Protection Agency. Federal 68698-68703 (1 979). 30. Hogan , M. D. et a I. Association between chloroform levels in finished drinking water supplies and various site-specific cancer mortality rates. J ournal (1979).
0/
environmental pathology and toxicology , 2:873
31.
Bull ,
R. J.
et a l. In depth biochemical , pharmacological
and metabolic studies of trihalomethanes in water.
P , oceedings
0/
the NCI/EPA/NIOSH Workshop
on Environmental atld Occupational Carcinogenesis , 1980 (in press).
第五章感官指标 和特性
1. 铝 1. 1 概述
自然界存在着丰富的铝化物，并且在水中常可被发现。铝
盐广泛用于水处理过程，以去除水中的颜色和混浊。与从食 物中摄入的生吕相比，从水中摄入的量很少。未发现摄入铝盐 对人具有任何有害作用，但是如果在出厂水中铝浓度超过
0.1 毫克/升，则将增加在配水系统饮水中产生颜色的机率，
因此，基于感宫的考虑，推荐在饮水中铝的建议值为 0.2 毫 克/升。虽然在该浓度下可能出现一些颜色，但是当使用铝
化物进行水处理时可能难以达到更低的浓度，所以该值是一 种折衷的考虑。在这种情况下，应特别注意维护配水系统。
1. 2
存在
铝广泛分布于自然界，是所有土壤、植物和动物组织的 一种成分 EIH430 由于铝在自然界的广泛分布以及人类的活
动，所以因天然存在或污染而使铝存在于空气、食物和水 qa Rdw
1. 2.1
和底泥的污染是铝进入水环境的主要途径。在水中铝的浓度 很不相同，在铝加工厂附近可能超过 10 毫克/升 E62. 在任何 -358 一
由!
水
工业废物、腐蚀、矿物和土壤的沥滤以及来自大气灰尘
特殊水体中的浓度是由 pH 、可能存在的自己位剂的种类和浓
度、矿物质成分的氧化状态以及在系统中的氧化还原能力所 决定的 p 很多酸性水在天然情况下含有高浓度的铝，设想是 通过渗漏过程而造成的。 在水处理过程中，特别是当水有颜色和混浊时，已广泛 使用铝盐(例如明矶或铝酸纳〉混凝的方法去除细颗粒矿物
质和有机物。因此，由于在水处理过程中使用这些铝盐而影
响饮水中铝的浓度。但是作为混凝剂使用的大部分铝能以不 海性铝盐的形式通过沉淀或过滤而被除去。然而为确保水的 感官良好和有效的消毒，仔细地进行这些操作过程对于成功 地进行水处理还是必要的 E730
尽管在水处理中，要使水中铝的残留量尽可能少，但是
常常会残留一些。据报道，在处理的水中浓度为 <0.01"""2 毫克/升 (2) 。应该指出，这些数值表明"明矶"的投句口量高
了 12 倍。在出厂水中铝的浓度大于 0.3 毫克/升时，通常反 应了水处理过程中混凝、沉淀或过滤阶段的失败。 在配水过程中，一部分铝可能沉淀下来，从而使其浓度 逐渐降低 (8) ，而蓄积在配水系统中，特别是在水流慢的部
位，并可与铁、锤、砂、有机物和微生物一起形成沉淀，当 水的流速改变时，这些沉淀容易被搅动，而使其出现在用户 的水龙头中，致使人们从感官上不能接受这种水 (9) 。在一般 情况下低浓度铁本来不引起问题，但当铝存在时则可能使水 明显着色。已表明 E103，如果铝在出厂水中的浓度超过 0.1 毫
克/升，则在配水系统中产生颜色的机率增加，因此用户的
抱怨增多。
-859-
1. 3
接触途径
从饮水中摄入的铝仅占人每日估计摄入量的很小部分。 大量的摄入来自食物，基于已发表的资料(1 1 甜 1 4) ，估计铝的
总摄入量为 88 毫克/人/日。每日摄入 2 升含铝量为1. 5 毫
克/升的水仅能摄入铝 3.0 毫克/人/日，即小于每日正常摄 入量的 4% 。
1. 4
对健康的影响
铝似乎并不是人的必需营养素。在正常情况下，铝盐不
能从食物和水中被吸收，但能与磷酸盐结合，从粪中排 出(1 5) 。长期大量使用氢氧化铝作为"抗酸剂"，可能引起磷
酸盐从身体中的过量丢失。除骨器外，摄入的铝并不大量地 蓄积在组织中 E183，而随灰尘吸入的铝化合物蓄积在肺 (17) 和 淋巴结中 (18) 。
按 2.5 毫克/公斤体重/日的剂量给大鼠 6 个月的铝，引 起最小的全身毒性和最低的性腺毒作用(1 9) 。大鼠饮用含铝 浓度为 5 毫克/升的饮水，未见对平均寿命、长寿、肿瘤发 生率和临床生化的影响 (20) 。
使用不同的哺乳动物和各种铝盐进行的一系列研究均未 发现致癌的证据 (2 1)。
铝与某些神经功能紊乱〈例如 dialysis 痴呆和早老性 痴呆〉有关 (20 ， 21) 。但是，是否铝的存在引起这种情况或铝
的存在只是其他因素作用的一种指标还不清楚. 参考文献
1. Cotton , F. A. -360-
& Wilkinson , G.
Advanced inorganic
chernistry , 3rd ed. New York. Wiley-Interscience , 1972 , pp. 261~262.
2. Sorenson , J. R. J. et a l. Aluminium in the env ironment and human health. Envi ronrnental health
perspectives , 8:3
(1 974).
3. Underwood , E. J. Trace elernents in hurnan and
anirnal nutrition , 3rd ed. New York , Academic Press Inc. , 197 1. 4. Saakashvili , T. G. & Kvirikaze , N. A. Content of certain trace elements in human biood. Trudy
lnstituia Urol. , Akadernia Nauk Gruzinskoi SSR , 1 :93 (1 962) J Chernica/ abs tr' acts , 61: 4784 (1 964). 5. Monier-Williams , G. W. A lurniniurn in /ood. London , Ministry of Health , 1935 (Reports on Public Health and Medical Subjects). 6. Sylvester , R. O. et a l. Factors involved in the location and operation of an aluminium reduction
0/ the 22nd lndustria/ Waste Conference. Lafayette , IN , Purdue University , 1987 , pp. 441~454. plant. Proceedings 7. Ridgw 町，
J. et a l. Water quality changes-chemical
and microbiological studies. In: Water d istri bution
system s • Medmenham , England , Water Research Centre , 1979. 8. Ainsworth , R. G. et a l. Deposits , corrosion products , and corrosion mechanisms in iron mains. Ina Research Centre. 1979. 9. Ainsworth , R. G. et a l. The introduction of ne ω
Water
distribution systetns. Medmenham. England , Water
water into old distribution systerns. Medmenham , -361 一
England , Water Research Centre , 1980 (TR 146).
10. Vozar , L. Content of aluminium in the diet and its biological action. Vo prosy pitani ja , 21' 28 (1 962). 11. Zook , E. G. & Lehmann , J. Total diet study , content 。f
ten minerals-aluminium , calcium , phosphorus ,
8odium. potassium , boron , copper , iron , manganese and magnesium. Journal 01 the Association 01 Official
Agricultural Chemists , 48'850
(1 9~5).
12. Gabovich , R. D. Contents of some trace elements in the food in certain cities and towns of the USSR.
Gigiena i sanitarija , 31141 (1 966). 13. Jaulnes , P. & Hamella , G. Présence et taux des oligo-éléments dans les aliments et les boissons de l' homme.
Annales de la nutrition et del' a1i ment ,
25' B133 (1 97 1). 14. Hamilton , E. 1. & Minski , M. J. Abundance of the chemical elements in man's diet and possible relations with environmental factors. Science 01 the total
environment , 11375 (1 973). 15. Thienes , C. H. & Haley , T. J. Clinical toxicology. Philadelphia , Lea & Febiger. 1972. pp. 169-170.
16. Deichmann , W. B. & Gerarde , H. W. Toxicology 01
drugs and chemicals. New York. Academic Press , 1969 , p. 88. 17. Ham iI ton , E. 1. et a I. Concentration and distribution 。f
80me stable elements in healthy human tissues environm~ntal
from the United Kingdom ,
study.
Science and the total envi ronment , l' 341 (1 973). 18. Krasovskii , G. N. et a I. Experimental study of biological effects of lead and aluminium following 一 362-
oral administration. Enuironmental hea/th
perspectiues. 30:47-51 (1979). 19. Mahurkar , S. D. et a l. Electroencephalogrlphic a"d radionuclide studies in dialysis dementia. Kidney
international , 13:306 (1 978). 20. Elliott , H. L. et a l. Aluminium toxicity during regular haemodialysis. British med ical journal , 1: 1101 (1 978). 2 1. Crapper , D. R. et a l. Brain aluminium distribution in Alzheimer's disease and experimental neurofibrillary degeneration. Science (Washington)
, 180: 511
(1 973).
2. 氯化物 2.1 概述
氧化物广泛分布在自然界，通常以铀 (NaC l)、饵 (KC l)和钙 (CaCl z ) 盐的形式存在。约构成陆界的
0.05% (1)。迄今，环境中最大量的氯化物存在于海洋中。 氯化物存在于天然水体中可能是由于盐类沉积物鸪溶 解 (2) 、为清除冰雪而向公路撒盐引起的污染。阳 7) 、化工废水 的排放 (8) 、油井的开采 (9) 、污水排放 (10) 、灌溉排7]( (11)、垃
圾渗漏的污染(1 2) 以及沿海地区海水的倒灌 EI30 每种来源均
可造成地面水和地下水的局部污染。氯离子易转移，但是最 终都转运到封闭的盆地或海中 (1) 。
2.2
存在
氧化物通常以低浓度存在于天然的地面水中。在未污染 的水中浓度常低于 10 毫克/升，并且可能往往低于 1 毫克/ 升 (11 ， 13) 。
在植物和动物性食品中，氯化物天然存在的浓度往往低 于 0.36 毫克/克(1 4) 。在加工或烹调时或在餐桌上加盐均能
明显增加食物中氧化物的浓度。
2.3
接触途径
估计氧化物从食物巾的每日摄入量是困难的，因为在餐 桌上广泛使用盐作为调味品。食用无盐膳食每天大约摄入 一 364 一
600 毫克氧化物(1 5 ， 16) 。但是由于向食物中加盐，每天氧化 物的平均摄入量为 6 克，可能高达 12 克 (17 ， 18) 。从饮水中氯 化物的平均摄入量约为 100 毫克/日 (10 ， 17， 18) 。从空气中的摄
入量可忽略不计。
2.4
对健康的影响
氧化物是人体内最丰富的阴离子，该离子及与其相结合 的阳离子一起，明显地构成了细胞外液的渗透活性; 88% 的 体内氯化物存在于细胞外液中(1 9) 。一个 70 公斤体重的正常 人约含 8 1. 7 克氯化物 (19) 和 45 升水。
体内水和电解质的平衡是通过调节膳食的总摄入及经肾 脏和肠道的排泄来维持的。正常人体对氧化物的吸收几乎是 完全的。大部分液体和电解质的吸收是发生在小肠的前半部 分 (20) 。每天液体的正常丢失约相当于1. 5~2 升水并随之排
出 4 克氯化锅，其中 90~95 %的氧化制从尿中， 4~8% 从 粪中，约 2% 从汗中排泄。每日氯化物最低的排泄量约 530 毫克 Cl930 基于这一估计，成人每日需要从膳食摄入的氧化
物为 9 毫克/公斤体重〈即对于一个 70 公斤体重的人为 630 毫克氯化物，相当于每人每日摄入的食盐量稍高于 1 克)。 对于 18 岁以下的孩子来说，每日从膳食中摄入 45 毫克/公 斤体重的氯化物是足够的(1 9) 。
饮水中氯化物的味阔浓度取决于与其相结合的阳离子，
但是通常其范围为 200~300 毫克/升。在水中氯化纳、氧化
伸手日氯化钙的味阔浓度分别为 210 、 310 和 222 毫克/ 升 (21 ， 22) 。当用含氧化物浓度为 400 毫克/升(以氯化的表
示〉或含氧化物浓度为 530 毫克/升(以氯化钙表示)的水 调制咖啡时，黑味道受到很大的影响 E232. ← 365 一
虽然从饮水中摄入的氧化物占每日摄入总量的很小部
分，但是因为在常规水处理过程中不能从水中去除氯离子， 所以从水的感官考虑，推荐在饮水中氯化物的建议值为 250
毫克/升。 参考文献
1. National Research Council of Canada Associate Committee on Scientific Criteria for Environmental Quality. Effects
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2. National Research Counci I. Nutrient and toxic
substances in water /or livestock and poultr y. Washington , DC , National Academy of Sciences , 1974. 3. Murray , D. M. & Ennst , V. F. W. An economic analysis of the environmental impact of highway de-icing salts. N ational technical in/ormation
service publication. 253:268 (1 976). 4. Pollock , J. J. & Toler. L. G. E//ects
0/ highway
de-icing salts on groundwater and wate 俨 supplies in Massachusett 宫. Washington , DC , Department of the Interior. 1972 (US Geological Survey). 5. Terry , R. C. Road salt , drinking water and sa/ety. Cambridge. MA. Ballinger , 1974. 6. Hutchinson , F. E. Effects of highway salting on the concentration of sodium chloride in private water supplies. Research in li/e sciellces. 15 (1 9ô9). 7. Ralston , 1. G. De-icing salts as a source 197 1. -366 一
0/ water
po l/ ution. Toronto. Ministry of the Environment.
8. Li ttle , A. D. lnorganic chemical pollution Protection Agency , 197 1.
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/reshwater. Washington , DC , US Environmental 9. Pettyjohn , W. A. Water pollution by oil-field brines ,md related industrial wastes in Ohio. Ohio jour:wl
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science , 71:257 (1 37 1).
10. Pettyjohn , W. A. Wate俨 quality in a stressed
environment. Minnesota , Burgess Publishing Co. , 1972. 11. Bond , R. G. & Straub. C. P. Handbook Rubber Co. , 1973. 12. Schneider , W. J. H ydrologic implications
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environmental control , Vo l. 3 , Cleveland , Chemical
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so/id-waste disposal. Washinghton , DC , Department of the Interior. 1970 , pp. F1-F10 (US Geological Survey Circular 601-F). 13. National water quality data bank. Ottawa , Inland
Waters Direc torate , Water Quality Branch , 1976. ,
14. Long , C. et a l. Biochemists' handbook. London , E. and F. N. Spon Ltd , 19tH. 15. Dahl , L. K. $alt and hypertension. American journal
01 clinical nutrition , 25: 231 (1 972). 16. Meneely , G. R. A review of sources of and the toxic effects of excess sodium chloride and the protective effect of extra potassium in the diet. Plant /oods
/or human nutrition , 23:3 (1 973). 17. Zoeteman , B. C. J. & Brinkman. F. J. J. Human intake of minerals from drinking water in the European Communities. In: Hardness and drinking
water and public health. Proceedings of the European Scientific Colloquium , Luxembourg. Oxford , -367 一
Pergamon Press , 1976 , p. 175. 18. Sodium , chlorides and conductivity in drinkin [J ω ater. Report on a W HO Workin[] Group. Cophenhagen , WHO Regional Office for
Europe , 1979 (EURO Reports and Studies , No. 2).
19. Dietary standard lor Canada. Ottawa , Health Protection Branch , Department of National Health and Welfare , 1975. 20. Sladen , C. E. Ahsorption of fluid and electrolytes in health and disease. In: McColl , I. & Sladen. G. E. , ed. 1 ntestine absor ption in man. London , Academic Press , 1975 , p. 5 1. 2 1. Whipple , G. C. The value
01 pure water.
New York ,
John Wiley and Sons , 1907. 22. Richter , C. P. & Maclean , A. Salt taste threshold of humans. American journal (1 939).
01 physiolo [J Y' 126:1
23. Lockhard , E. E. et a I. The effect of water impurities on the flavor of brewed coffee. Food research , 20: 598 (1955).
-368-
3. 颜色 3.1 概述
饮水中的颜色可能是由于存在下列物质 z 有色的有机 物，通常是腐擅质，金属，例如铁和髓，或高度带色的工业 废水，最常见的是纸浆和纸以及纺织业的废水。虽然饮水中
的颜色主要是对水的美学方面的影响，但是对感官方面的影 响可能被当做是对健康的影响。
经验表明，如果饮水中含有在感官上令人不快的颜色， 那么用户就可能寻找并不安全的其他水源。 大多数人对在一杯水中高于 15 真色度单位 (TCU) 的A
水能觉察出来，在水氯化前去除过多的颜色也能减少三卤甲 烧的产生及去除氯代有机物所产生的味道，所以限制自来水 中的颜色也就限制了与腐殖质结合或吸附在其上的有害物质 的浓度。 推荐饮水中颜色的建议值小于 15 真色度单位。
3.1.1
来源
饮水中出现颜色是由于吸收了某些波长的正常"白光"， 水中存在有色物质以及悬浮颗粒物引起的光线散射所 致口， 2)。在含悬浮物的水中测定的颜色称为"表观颜色"
(apparent colour) J 产生的颜色。， 43@
经离心或过滤从水样中去除颗植物
后测定的颜色为"真色"，这种颜色是在真溶液中由腐殖质 一般而言，水样的真色明显低于其表观颜
色 ω。
8. 1. 2
测定
水样水中的颜色是用肉眼与一系列含有已知量的氯铀酸 何和氯化钻 (ll) 的标准榕液进行比较而测定的。由于铀 钻标准法是测定水的天然颜色，所以当公共给水的颜色与标
准溶液比色有困难时可能表明水已受到污染。 一个真色度单位相当于在特定的实验条件下每升1. 0 毫 克铀(以氯铀酸盐离子的形式存在)的溶液所表示的颜色 ω 。
大多数用户能察觉在一杯水中含 15 真色度单位曲颜色，在 大量水中，例如在一个白色的浴缸中， 5 真色度单位是明显 的，、而几乎无人能察觉 3 真色度单位的颜色 (5) 。
天然地面水的颜色通常随 pH 增加而增加，一般称其为 4指示效应叫1)。因此，为考虑这种效应，广泛地建议测定样 品颜色时应记录 pH 值 (4) 。
3.2
存在
对水中颜色的抱怨在数量上通常接近于对昧和噢抱怨的
总和。天然水中的颜色主要是由于有机物质，特别是腐殖质 所致，腐殖质来源于植物的腐烂和经水提取到地面水中。铁
和瞌常常存在于地下水和某些地面水中并产生颜色。快在饮 水中的另一个重要来源是输送水的铁管被溶解。铁和锤能分
别引起红水平口黑水现象。从铜管溶解的铜可引起卫生器具着 蓝『绿色，此外在例外的情况下可使水呈淡蓝色。颜色深的
废水，特别是来自纸浆和纸以及纺织工业的废水能引起水出 现颜色的问题。
微生物引起的颜色问题能产生"红水"现象，这是由于
二价铁氧化为三价，铁从溶液中以氢氧化铁的形式沉淀下来 坪使水呈现特有的红色。在严重的情况下，这些"铁细菌"
可将配水管道堵塞。同样，细菌能把溶解的钮氧化为不溶性 氧化物而使饮水呈现黑色。以地下水为水源的给水比地面水 更经常出现颜色问题。 颜色的去除
3.2.1
在实践中，有时用化学氧化的方法来补充混凝和过惊能
最有效地去除水中的颜色。在瑞典约 200 个水厂于 1976 年
记录的资料表明，处理前的水中颜色范围为 <5 ......， 150 真笆度 单位，而处理后降到 <5 ......， 25 真色度单位朋。
3.3
对健康的影响
对自来水中颜色的限l值，习惯上是依据感官来考虑。但
是已经注意到，当供应有色的水时，可能使用户寻找无色， 然而也许不安全的其他饮水水源 (7) 。其他与健康有关的方面
包括颜色和产生某些氯代有机化合物之间的关系 F 干扰水的
处理和增加氯的消耗量。
对天然带色的有机化合物几乎未进行毒理学研究。在 90 天内雄性大鼠饮用含低灰的土壤灰黄霉酸市1] 剂为 10 、 100
和 1000 毫克/升的饮水，未见体重、食物和水的摄入量、器 官/体重比值以及组织形态学的明显改变 (G. C. Becking 和 A. P. Yagminas ，未发表的资料， 1978) 。每天给大鼠 经口灌入剂量为 1000 毫克/公斤相同的灰黄霉酸共 14 天生
动物未出现死亡。但是与对照动物相比，体重的增长速度减 慢以及某些肾脏酶的浓度轻度改变。迄今在一项最有关的研
究中，给予大鼠饮用含两种法度腐殖酸的饮水达 19~35 周，
作者得出结论 z 采用安全系数 100 ，含腐殖酸浓度为 2.5 毫 一 37J 一
克/升的饮水对人是安全的 (8) 。
关于微量金属及其腐殖酸盐络合物对人体健康影响的比 较毒性研究很少 (9) 。已表明，当铁、铅、坝、银、铜和铸为
其腐殖酸盐络合物时，对哺乳动物静脉注射的急性毒性明显 增加，但是经口摄入腐殖酸锚的毒性，至少比醋酸铅的毒性
小 60% (1 0) 。
已发现，在腐殖酸存在的情况下，透过肠道的
!离子物质(钙、镜、铁、锤、铮和硫酸盐〉的量增加 50""'"
100% (11)。遗憾的是，所研究的物质不包括有毒的微量金 属。迄今，关于哺乳动物对毒性有机物的腐擅酸盐络合物的
生物可利用性问题尚未见公开发表的资料。
3.4
其他方面
从土壤中分离的腐殖质(1 2 ， 13)对有机物的吸附是值得进
一步研究的问题，这些有机物在水中的含量可能超过其溶解 度 (14 ， 15) 。由于腐殖质有聚阴离子电解质的性质，所以对带
阳电荷无机离子〈例如重金属〉的溶解、运输和沉积起了必
要的作用。大部分金属在与水中腐殖质接触时发生络合。络 合物的形成将明显地增加金属的溶解度，例如，水中天然的 腐殖质可使铁的溶解度增高 10& 倍(1 6) 。
在一定条件下某些金属与腐殖质形成不溶性络合物，这 是在水处理中使用铁和铝盐的依据。
在水处理中有毒金属络合物的命运是特别值得关心的问
题。最普遍的看法是，虽然与水中悬浮固体络合的有毒金属 至少能去除一部分，但是珞解性微量金属在常规水处理过程 中也许仅能去除微量(1 7) 。
溶解性腐殖质能使饮水产生味道的看法还未得到确 t正(1 8) ，因为迄今尚未对这一问题进行研究。高度带色的污染
qo 9" 哼d
水体常常出现令人厌恶的昧，但是还不清楚这两者间相关的 程度。已知水中带色的有机物能剌激很多微生物的生长 E193 ，
其中某些微生物与水中出现气味直接相关。 腐蚀和结垢之间的关系以及水中腐殖酸的含量都是很复
杂又很重要的问题。小量的腐殖质(1 --2 毫克/升)有助于 在配水系统中碳酸钙保护层的沉滤 (20) 。当加石灰作为腐殖 性水处理后的纠正步骤时 E213 ，大量的腐殖酸可能在配水系统
中引起妨碍流速的"腐殖质泥"沉淀。含有很小量 i容解性腐 殖质的水对金属的腐蚀性较含大量腐殖质的水强 E22 ， 2330
由于腐殖酸及其某些金属络合物在自来水 pH 值的情况 下难洛于水，故可能造成水样的部分混i虫，而水中溶解性腐殖 质主要以肢体分散相存在，并且目测法测定的温度受呈肢体 大小的颗粒物影响 (2) ，所以水中这样的颜色将影响浊度值。
在配水系统中难以保持游离有效氯，这可能是由于在处 理的水中存在着带色有机物。虽然 1949 年以来就知道这一 事实 (18) ，但是直至 1974 年发现剧， 25) 在氯化水中的氯与溶
解性腐殖质反应能生成相当大量的氯仿和三卤甲皖后才进行
了详细的研究。三卤甲院是氯(偶尔存在漠和腆)与腐磕质 反应的产物，但是幸运的是，用于混凝的方法能从原水中去 除大部分有机物前体 (26) 。
颜色能干扰对水中许多成分的化学分析。例如在水和废 水标准检验法中要求，在比色分析时必须对颜色进行补偿或 去除颜色 (4)。
腐殖质与金属络合的性质能干扰非比色分析法。当使用 络合剂以及用有机溶剂进行提取而浓缩样品时，腐磕质能干 扰微量金属的分析 E273@
-37~~
参考文献
1. Black , A. P. & Christman , R. F. Characteristics of colored surface waters. J ournal
0/ the American
Water Works Association , 551753 (1 963). 2. Black. A. P. & Hannah , S. A. Measurement of low turbidities. Journal
0/
the American Water Works
Association , 57:901 (1 965). 3. Sawyer , C. N. & McCarty , P. L. Chemistry /or
sanitar y engineers , 2nd ed. Toronto , McGraw-Hill Book Company , 1967 , p. 299.
4. Standard methods /or the examination 0/ water and wastewater , 14th ed. Washington , DC , APHA , AWWA , WPCF , 1976. 5. Bean , E. L. Progress report on water quality criteria.
Journal
0/
the American Water Works Association ,
54: 1313 (1 962). 6. Anon. V A V AD 76 Vattenbeskaf fenhet 1976 CWater
quality 1976). Stockholm , Swedish Water Works Associa t ion , 1977.
7. Public health seruice drinking water standards. Rockville , MD. US Department of Health. Education and Welfare , 1962 , p. 21 (Public Health Service Publication No. 956). 8. Ja necek , J. & Chalupa , J. Biological effects of peat water humic acids on warmblooded organisms. Archiu
/Ur hydrobiologie , 65:515 (1 969). 9. Brown , V. M. et a l. Aspects of water quality and the toxicity of copper to rainbow trout. Water research , 8:797 (1 974). -374 一
10. Klocking , R. Influence of bumic acids on the toxicity of lead. Proceedings
0/
the European sociefy
0/
toxicology , 16 1258 (1 975). 11. Visser , S. A. Some biological effects of humic acid in tbe rat. Acta biologica et medica Germanica , 311569 (1 973) .
12. Flaig , W. et a l. In. Gieseking , J. E. , ed. Soil
components. Vo l. l. New York , Springer-Verlag , 1975. 13. Felbeck , G. T. Jr. In , McLaren A. D. & Skujins , J. , ed. Soil biochemistry. Vo l. 2. New York , Marcel Dekker , 1971 , pp. 54-56. 14. Kban , S. U. & Scbnitzer , M. Tbe retention of hydrophobic organic compounds by humic acid.
Geochimica cosmochimica acta , 361745 (1 972). 15. Hague , R. & Schmedding , D. Studies on the adsorption of selected polycblorinated biphenyl isomers on several surfaces. J ournal
0/
environmental science and health , 811 :1 29 (1 976). 16. Sbapiro , J. Effect of yellow organic acids on iron and otber metals in water. Journal
0/ the American
Water Works Association , 56:1062 (1964). 17. Committee on Water Quality Criteria , National Academy of Sciences. Water quality criterÎa 1972. Wasbington , DC , US Government Printing Office , 1973 , p. 51 (EPA-R-73-033). 18. McK 肘， J. E. & Wolf , H. W. , ed. Water qua1i ty
criteria. 2nd ed. Sacramento , CA , California State Water Quality Control Board , 1963 , p. 198 (Publication No. 3- A). 19. Prakasb , A. et a 1. Humic 8ubstances and aquatic -375 一
productivity. In , Povoledo , D. & Golterman , H. L., ed. humic substances 1972. Wageningen , The Netherlands , Pudoc , 1975 , pp. 259-268. 20. American Water Works Association. Water qua/ify
and treatment , 3rd ed. Toronto , McGraw-Hill Book Company , 1971 , p. 31 1. 2 1. Gjessing , E. T. Physical and chemical
characteristics 01 aquatic humus. Ann Arbor , MI , Ann Arbor Science , 1976. 22. Van Beneden , G. & Leclerc , E. Les matiêres humiques. Leur comportement dans les eaU :1 ou géneral , leur role dans la corrosion des metaux.
Techno/. water (Czech) , 8: 225 (1 964). 23. Moore , M. R. Plumbosolvency of waters. Nature , 243: 223 (1 973). 24. Rook , J. J. Formation of haloforms during chlorination of natural waters. J ourna/ 01 water
treatment and examination , 23:234 (1 974). 25. Bellar , T. A. et a l. The occurrence of organohalides in chlorinated drinking waters. Journal 01 the
American Water Works Association , 66:703 (1 974). 26. Stevens , A. A. & Symons , J. M. Measurement of trihalomethane and precursor concentration changes.
Journal 01 the American Water Works Association. 69:546 (1 977). 27. Pakalns , P. & Farrar , Y. J. The effect of surfactants on the extractioll-atomic absorption , spectrophotometric determination of copper , iron , manganese , lead , nickel , zinc , cadmium and cobalt.
Water research , 111145 (1977) , ← 376 一
4. 铜 4.1 概述
铜及其化合物普遍存在于环境中，因此经常在地面水中 发现。水中铜的性质取决于水的 pH 值和碳酸盐浓度以及溶 液中的其他阴离子(1)。
水处理过程通常能使微量金属从水中去除，但是在用户 龙头的饮水中铜浓度却可能高于水源水或高于进入给水系统 的巳处理的水。给水的各种理化特性影响铜从配水系统和家
庭管道中溶出。贮于铜容器中的水易于保持良好的细菌学质 量而不致使水质恶化 (2) 。海解的恫能使饮水着色和具有不良 味道 E330
4.2
存在
士壤中的铜含量取决于下列因素，如地理位置、是否靠 近工业区或使用化肥等。无机肥料、中铜的含量范围为 0.01 ......， 0.05 毫克/克 E430 而食物中的铜含量将因其生长的土壤中铜
的含量而异。通常在食物，如蔬菜、面粉、奶和肉制品 中，铜的含量小于 0.01 毫克/克\饮水中铜的浓度一般波 动在 0.01---0.5 毫克/升 (5) 。 a 国家卫生和幅利部 z 食品监测调查. FMOI 计划，渥太华， 加拿大，
1971 年 8 月 -1976 年 1 月(未发表资料).
-377 一
4.3
对健康的影响
铜是人体代谢的必要元素，在红细胞生成、组织中铁的 释放以及骨路、中枢神经系统和结缔组织的发育中具有一定 作用。通常铜与蛋白质结合。红细胞中的血铜蛋白和血浆中 的血浆铜蓝蛋白都含有铜，它作为其结构中不可分割的部
分，金属琉基组氨酸三甲内盐是铜贮备蛋白。已经分离出许
多含铜的酶，主要是细胞色素氧化酶、抗坏血酸氧化酶和尿 酸酶。 因为铜广泛分布于食物中，所以除使用特殊奶制品的婴
儿以外，人的膳食中不可能缺乏铜。在使用铁剂治疗婴儿营
养性贫血时，铜是一种有益的辅助治疗剂。但是，动物的许 多临床紊乱与缺铜有关 (6) 。 人摄入过量铜会导致对粘膜的严重剌激和腐蚀作用、广
泛的毛细管损伤、肝和肾损伤以及中枢神经系统先兴奋后抑 制。还可能发生严重的胃肠道剌激以及肝和肾的坏死性改 变。但是人和高等哺乳动物都很少发生铜中毒，这是由于铜
有强烈的催吐作用。在皮肤上使用铜盐具有腐蚀作用，并可 能引起丘彦水泡性湿彦，对眼的局部作用引起严重的炎症。
水中的铜有令人厌恶的涩味。味阔高于 5 毫克/升，但是当 铜在蒸锢水中浓度为 2.6 毫克/升时就可察觉出有味 (7) 。
4.4 生活用水。
其他方面
给水中的铜虽然不能梅成对健康的危害，但是可能影响 公共给水中存在铜能增加对铝和铮器具和设备的腐蚀作
用。当水中铜浓度超过1. 0 毫克/升时，衣服和管道说备可
能着色，所以铜在饮水中的建议值为1. 0 毫真/升。
参考文献 1. McKee , J. E. & Wolf , H. W. Water qua/ity criteria. Sacramento , CA , California State Water Quality Control Board , 1953. 2. Dhabadgaonkar , S. M. Metallic copper for disinfection of water in rural areas. Journal
'0/
Indiall Water
Works Association , 12143 (1 980). 3. Page , G. G. Contamination of drinking water by corrosion of copper tubes. New Zea[and journa[
0/
sciences , 16:349 (1 973). 4. Van Loon , J. C. & Li chwa , J. A study of the atomic absorption determination of some important heavy metals in fertilizers and domestic plant sludges.
Environment letters. 4:1 (1 973). 5. Zoeteman , B. C. J. & Brinkman , F. J. J. Inl Hardness
0/
drinking water and public health. Proceedings of
the European Scientific Colloquium , Luxembourg , 1975. Oxford. Pergamon Press. 1975 , p. 173. 6. WHO Technical Report Series. No. 532 , 1973 (Trace
elements in human nutritionl report of a WHO Expert Committee) . 7. Cohen , J. M. et a l. Taste threshold concentrations of metals in drinklng water. Journal
0/
the American
Water Works Association , 52:560 (1 960).
-379 一
5. 硬度 a 5.1 概述
水的硬度是衡量水与肥皂起反应能力的传统指标，硬水
则要求大量的肥皂才能产生泡沫。热水管、锅炉和其他家用 器具的结垢也是由于硬水造成的。水的硬度是由 i容解性多价 金属离子形成的。在淡水中，主要形成硬度的离子是钙和钱; 锦、铁、顿和主孟也起作用 E130 迄今，硬度通常是通过水样中 的多价金属离子和整合剂〈如 EDTA) 的反应进行测定的，并 且以等量碳酸锦浓度来表示口， 2)。也可以分别测定硬度中各
个成分的浓度来估计，并以等量的碳酸钙的和来表示，根据
等量碳酸钙的浓度可将饮水分类如下 z 软水 中等硬度水 硬水 高硬水
。 -60 毫克/升
60-120 毫克/升 120-180 毫克/升
180 毫克/升和以上
硬度也依据等量氧化钙
(Ca 0)
或氢氧化钙
(Ca(OH)2) 来分类。在国际单位制系统中建议将硬度表示 为每立方米中 Ca* 的克分子数 ω 。 尽管硬度是由阳离子形成的，但是也从碳酸盐(暂时 的〉和非一碳酸盐(永久的〉硬度方面进行了讨论 (4)。碳酸
盐硬度代表了珞解的碳酸盐和碳酸氢盐的含量，通过煮沸可 ·在第三章第 9 节(第 144 页}也讨论了与饮水硬度有关的问题。
•
380-
以将真去除或沉淀下来，这种硬度引起热水管和茶壶水垢的 沉积。非-碳酸盐硬度是，由于引起硬度的阳离子与硫酸盐、氯 化物~硝酸盐结合而形成的，称为"永久性硬度"，因为煮 沸不能将其去除。
碱度是水缓冲能力的指标，与硬度密切相关。在大多数 情况下，碱度是由阴离子或弱酸分子所形成的，主要是氢氧 化物、碳酸氢盐和碳酸盐 P 其他分子，如棚酸盐、磷酸盐、 硅酸盐和有机酸存在时也起到较小的作用，无论哪种分子构
成水的碱度，总是以等量碳酸钙表示。当地面水的碱度是由 于碳酸盐和/或碳酸氢盐的存在时，碱度值常常接近于硬度 值 E53.
5.2
存在
水中硬度的主要天然来源是沉积岩、土壤渗出和流出。 在正常情况下，硬水来源于厚的表土层和石灰石形成的区 域(4)。通常地下水较地面水硬度高。富含碳酸和溶解氧的地 下水往往能高度溶解土壤或岩石，而这些土壤和岩石含有大
量的矿物质方解石、石膏和白云石，因此硬度可达数千 毫克/升 (4 ，的。
硬度的两个主要工业来源是无机化学和采矿工业 (4， 7) 。 在建筑工业中，氧化钙用于砂浆、粉泥和灰浆，也用于生产
d纸浆和纸、精制糖、炼油、制革以及作为处理水和废水的化 学物 (8)。模用在纺织、制革和造纸工业的各种加工过程中。
镇合金广泛用于模具和钢模的浇铸、手提式工具、皮箱和一 般家用物品。钱盐用于生产金属镜、肥料、陶瓷制品、炸药 和医药 (9)。
-381-
5.3
对健康的影响
正如在 5.1 节中所述，构成7](硬度的主要因素是钙和模 离子。尚无证据表明，水中含有高浓度的钙或镜对健康具有 有害作用。 除高硬度的水给生活使用带来不便以外，其他的影响可 能是由于模和碳酸盐离子结合而引起缓泻作用。
饮水中钙离子的昧阔取决于存在的阴离子，其波动范围 为 100~300 毫克/升，钱离子的味阔小于该值 (10) 。为详细
了解水的硬度和心血管病的关系，应参考第三章关于饮水中
的无机成分对健康的影响〈见 147 页〉。没有推荐水中钙和模 的建议值，因为根据感宫的考虑推荐了总硬度的建议值。
5.4
其他方面
软水易于腐蚀管道，因此某些重金属，如铜、铮、铅和 钙就可能存在于饮水中 (11 帽 1 4)。而腐蚀作用和金属洛解的程
度与 pH 、碱度和溶解氧浓度有关。在某些给水中，腐蚀很 严重，必须对给水采取特殊的预防措施 (5) 。
在使用高硬度水的区域，由于沉积的结垢可能培塞家庭 内的管道(1 6'; 也能使厨房用具结垢并使肥皂的消耗量增加。
因此硬水会使用户感到厌恶和造成经济负担。公众对硬度的 接受程度在各居民区之间是不同的，这常常与多年来用户所
习惯的硬度有关，在许多居民区能耐受高于 500 毫克/升的 硬度。大约硬度为 100 毫克/升 (CaCû 3 ) 时，腐蚀和结垢
问题之间可基本平衡，但是基于感宫的考虑，推荐硬度的建 议值为 500 毫克/升 E173.
一 382 一
'考文献 1. Qua!ity criteria lor water , Washington , DC , US Environmental Protection Agency , 1976 (EP A-440/976-023). 2. Sekerka , I.
& Lechner , J.
F. Simultaneous
determination of total , non-carbonate and carbonate water hardnesses by direct potentiometry. Talanta , 22: 459 (1 975). 3. Glohmann , A. Harte des Wassers. In: Amavis , R. et aJ. , ed. Hardness 01 drinking water and public
health. Proceedings of the European Scientific Colloquium , Luxembourg , 1975. Oxford , Pergamon PreS5 , 1976 , p. 129. 4. Sawyer , C. N. & McCarty , P. L. Chemistry lor
sanitary engineers , 2nd ed. New York , McGraw-'Hill , 1967 (Series in sanitary science and water resources engineering) • 5. Thomas , J. F. J. lndustrial wate 俨 resOurces 01
Canada. Water Survey Report No. 1. Scope , Procedure and Interpretation of Survey Studies , Ottawa , Queen's Printer , 1953. 6. De Fulvio , S. & Olori , L. Definitions and classification of natura Il y 50ft and natura lI y hard waters. ChemicaI and physicaI characteristics of the water in some member states of the European communitY. Ina Amavis , R. et al 叫 ed. Hardness 01 d d nking water and pub !i c hea !t h. Proceedings of the ,
European Scientific Colloquium , Luxembourg , 1975. Oxford , Pergamon Press , 1976, p. 95. -383 一
7. Biesecker , 1. E. & George , J. R. Stream quality in Appalachia as related to coalmine drainage , 1965. In , Pettyjohn , W. A. , ed. Water quality in a stressed
environment. Minnesota , Burgess Publishing Company , 1972. 8. McQuarrie , M. C. Lime. In , McGraw-Hill
encyclopedia 01 science and technology. New York , McGraw-Hi Il, 1966. 9. Bech , A. V. , ed. The technology of magnesium and its a Il oys. In: McGraw-H iI/ encyc/opedia 01 science
and technology , New York , McGraw-H iIl, 1966. 10. Zoeteman , B. C. J. Sensory assessment 01 water
quality. Oxford , Pergamon Press , 1980. 1 1. Neri , L. C. Some data from Canada. In , Amavis R. , et al., ed. Hardness
0/ drinking water and public hea/th. Proceedings of the European Scientific Press , 1976 , p. 343.
Co Il oquium , Luxembourg 1975. Oxford , Pergamon
12. Sharrett , A. R. & Feinleib , M. Water constituents and trace elements in relation to cardiovascular diseases. Preventive medicine , 4'20 (1975).
13. Craun , G. F. & McCabe , L. J. Problems associated with metals in drinking water. J ournal 01 the
American Water Works Association , 67:593 (1 975). 14. Neri , L. C. & Hewitt , D. Review and implications of ongoing and pro jected research outside the European communitieö. In , Amllvis , R. et a l., ed. Hardness
0/
drinking water and public health. Proceedings of the European Scientific Co Il oquium , Luxembourg , 1975. Oxfo J' d , Pergamon Press , 1976 , p. 443. -884-
15.
Mulle 口，
E. D. & Ritter , J. A. Potable-water corrosion
contro l, J ournal
0/ the American Water Works
Association , 66:473 (1 974). 16. Coleman , R. L. Potential public health aspects of trace elements and drinking water quality. Annals
0/
the Oklahoma Academy
0/ Science , 5:57
(1 975).
17. Bean , E. L. Quality goals for potable water. Journal
0/
the 4merican Water Works Association , 60:1317
(1 968) •
-385 -
6. 硫化氢 6.1 概述
硫化氢是一种易燃的有毒气体，具有臭蛋昧(1)。梳化氢 以及碱金属和碱土金属的疏化物溶于水 E23 ，溶解的硫化物盐 类在水中解离为硫离子，与氢离子反应生成硫氢离子 (HS-)
或硫化氢 (H 1 S) 。这些成分的相对浓度与水的 pH 值有关 F 硫化氢浓度随着 pH 值的降低而增加口， 3) 。当 pH 值大于
10 时，就会存在高浓度的硫离子。
在良好曝气的水中，硫化氢被氧化为硫酸盐。经生物氧 化为元素硫 E32 ，并且在自然界"硫的循环"中硫化物是必不 可少的环节 (4)。
6.2
存在
在自然界中硫化物存在于矿石、石油和煤矿床中 (5) 。铜、
铅、铸、镇和其他开采的金属可能以简单或复杂的硫化物形 式存在。硫化铁常与这些矿石共生(1)。
硫化物也存在于石油和石油化工厂、化学工厂、煤气厂、 造纸厂、重水厂以及制革厂排出的工业废物中 (1 ， 3 ， 6 ， 7) 。院化 物是由硫酸盐-还原菌产生的(1， 3 ， 8-10)。这种细菌在配水系
统中生长可能是饮水产生臭和昧问题的主要原因。 在密西西比河水中硫化物的浓度约为 0.092 毫克/升，在 明尼苏达圣保罗的池塘和井水中 (1 1)分别含有 0.16 和 ← 586 一
0.19 毫克/升。 大气中硫化氧的天然浓度范围为1. 5 104
x 10-4~4.6 x
毫克/立方米 (6)。在工业区明显增高。
硫化物存在于很多生的和熟的食物中。细香葱和蒜含有 二甲硫和二甲三硫化合物 (12)。在洋葱头中 s- 甲基甲硫氨酸
的浓度小于 0.001 毫克/克，在西红柿中平均含量为 0.003 毫 克/克，在洋白菜中达 0.05 毫克/克 (13) 。
二甲硫在英国的淡色啤酒
(0.0002 .......， 0.0037 毫克/升〉 中是→种重要的调味
和欧洲啤酒 (0.003 .......， 0.114 毫克/升〉
品 (10 。也存在于蛋、乳和乳制品中 (13) 。
硫化氢在熟肉中的含量范围从 0.276 毫克/公斤〈牛肉 馅〉到 0.394 毫克/公斤(羊肉馅 )05) 。在热乳制品中硫化氢
的浓度从 0.80 毫克/升(脱脂牛奶，
含 0.1% 脂肪)至
1. 84 毫克/升〈奶油，含 30.5% 的脂肪) (16) 。
硫化氢和其他可溶性硫化物用于色素和染料的生产、制 革、纸浆和化学加工，也用于生产化妆品 07) 。含高浓度硫 化氢的泉水用作药浴和医药目的.
6.3
接触途径
关于食品中硫化物含量的资料还不充分，所以尚未估计 硫化物每日从膳食中的摄入量。对它的接触可能来源于食用
啤酒和谈色啤酒、海产品、熟肉、热奶和芦笋以及其他蔬菜。 在英国，从人造甜味剂、汽水、奶油和果子冻中二甲硫的 ?每日最大可能摄入量"估计为1. 7 毫克 (17) 。
若每日吸入 20 立方米空气，空气中硫化氢浓度为天然 坡度，则每日摄入量为 0.003-....0.01 毫克.
由于缺乏饮水中硫化物含量的资料，所以不能估计从饮 巧，
水中的摄入量。
6.4
对健康的影响
碱金属硫化物易于迅速地经肠道吸收 E1830 硫化氢在血
液和组织中转化为碱金属硫化物。硫化物经肾和肺排出，而 金属硫化物经肠排出 (8) 。硫化物也可能被氧化为硫酸盐和 硫代硫酸盐，并经肾脏排出 E1830
硫化氢能阻断某些酶系统的活性，其中某些酶系统直接
参与细胞氧化过程。已经报道了对脱氢酶(唬咱酸脱氢酶)、 磷酸醋酶 (ATP 酶〉、氧化酶(多巴氧化酶〉、联酸所酶、 二肤酶、苯眈胶酶和某些含铁酶的这种抑制作用 (6) 。其作用
机制还不完全清楚，然而认为这是由于形成金属硫化物，降 低了酶对阳离子的可利用性所致。 碱金属硫化物刺激粘膜上皮(1 8) 。当经口摄入时引起恶 心、呕吐和上腹部疼痛 0 ， 18)。口服 10.-..15 克硫化铀可以致 死 E1830 每日经口摄入剂量为 250 毫克/公斤体重的二甲硫达
14 周，未引起大鼠的任何毒性反应，该剂量相当于一个 70 公斤体重的成人每日摄入量为 17.5 克(1 7) 。
吸入硫化氢气可引起呼吸中枢麻痹而导致死亡(1 9) ，急性 吸入中毒阔值范围为 700'-"1000 毫克/立方米〈未注明接触 时间〉。吸入 1400'-"2100 毫克/立方米的硫化氢剌激神经系
统，并在 30 分钟内引起死亡 F 高于 2800 毫克/立方米导致 神经系统麻痹而发生即刻死亡。症状的发展包括突然疲劳、
眩晕、紧张焦虑、嗅觉功能丧失、袁竭、呼吸停止和死亡。 嗅觉的麻痹能妨碍察觉浓度超过 225 毫克/立方米的硫化 氢 (6) 。空气中浓度达 0.12 毫克/立方米时引起精神抑郁，
1.5~43 毫克/立方米引起结膜炎症和视力障碍阳。 70~700 -388 一
毫克/立方米引起慢性中毒，出现精神改变、眩晕、嗜睡、心 搏过速、咳嗽和呕吐白， 19) 。
水中的硫化氢引起令人庆恶的臭和昧，是造成用户抱怨
的主要原因。估计谣解的碗化氢嗅和味阔浓度范围为 0.05.........0.10 毫克/升 0 ， 9) 。硫化物的嗅和昧阔浓度约为 0.2 毫克/升 (1) 。因此人们不可能饮用有害浓度的硫化物，因为
令人厌恶的嗅和昧浓度远低于毒性浓度 F 所以选择的建议值 应为用户不能察觉的浓度。
6.5
其他方面
与溶解性铁结合的硫化氢在管道阳设备中能产生黑色沉
淀，使所洗的衣物被着上黑色。 参考文献
1. McKee , J. E , 。导 lity
& Wolf ,
H. W. Water qua 1i ty criteria ,
2nd ed. Sacramento , CA , Califor Jl ia State Water Control Board , 1963. pp. 156-7 , 271-2 , 277 , 335-336. 2. Senko , M. J.
principles and properties , 2nd ed. N E: w York , M,:Graw-H ilI Publishing
& Plane. R. A.
Chl~mical
Co. , 1974 , pp. (39-640.
3. Quality criteria for water. Wa!hington , DC , US Environmental Protection Agenc: r , 1976 , pp. 410-41~._ 、
4. Smith , R. L. Ecology and field biology , 2nd ed. New York , Harp)r & Row Publ i> hers , 1974 , pp. 88-89 , 123-128. 5. 民 ational
Research Council of Canada. Sulfur and its
inorganic derivafives in the Canadian environment. -389 一
Associate Committee on Scientific Criteria for Environmental Quality. OUawa , National Research Council , 1977. 6. Booras , S. G. Hydrogen sul/ide. health effects and
recommend ed ai r quality stand ar d. IIlinois Institute for Environmental Quality , 1974 (NTlS PB-233 843). 7. Colby , P. J. & Smith , L. L. , Jr. Survival of Walleye eggs and fry on paper fibre sludge deposits in Rainy River , Minnesota. Transactions
0/ the American
Fisheries Society , 96'278-296 (1 967). 8. Adelman , 1. R. & Smith , L. L. , Jr. Toxicity of hydrogen sulfide to goldfish (Carassius Auratus) as influenced by temperature , oxygen , and bioassay techniques. J ournal
0/ the Fisheries Research (1972). Yo 俨 k
Board
0/ Canada , 29:1309
9. Acree , T. E. & Splittstoesser , D. F. Prevention of H1 S in wine fermentation. New
/ood and li/e
sciences bulletin , 5'19 (1 972). 10. Kadota , H. & Ishida , Y. Production of volatile sulfur compounds in microorganisms. Annual review
0/
microbiology , 26 :1 27 (1972). 11. Broderius , S. J. & Smith , L. L. , Jr. Direct determination and calculation of aqueous hydrogen sulfide. Analytical chemist ，γ ， 49:424 (1 97 7).
12. Crewe , R. M. & Ross , F. P. Biosynthesis of alkyl 8ulphides hy an ant. N ature , 254: 448 (1 975). 13. Hattula , T. & Granroth , B. Formation of dimethyl sulphide from s-methylmethionine in onion seedlings (Allium Cepa). Journal
0/ the science 0/ /ood and
agri culture , 25'1517 (1 974). 一 390 一
14. Niefind , H. J. & Spaeth , G. Some aspects of the formation of dimethyl sulfide lhrough brewer's yeast and beer spoilage microorganisr~s. Proceedings
the Annual Meeting 01 the American Society Brewing Chemists , 33.54 (1 975). 15. Kunsman , J. E. & Riley , M. L. A comparison of other meats. Journal
01 01
hydrogen sulfide evolution from cooked lamb and
01
lood science , 40'506 (1 975).
16. Thomas , E. L. et a l. Determination of hydrogen sulphide in heated milks by gas liquid chromatographic head space analysis. Journal 1865 (1 975). 17. Butterworth , K. R. et a l. Short-term toxicity of dimethyl sulphide in the rat. Food and cosmetics
01
dairy science , 59'
to :x icology , 13 :1 5 (1 975). 18. Th ienes , C. H. & Haley , T. J. C /i nical to :x icology , 5th ed. Philadelphia , Lea & Febiger , 1972 , pp. 59-60. 19. Henkin , R. I. Effects of vapor phase pollutants on nervous system and sensory function. In , Finkel , A.
J. & Duel , W. C. , ed. Clinical implications
01
ai
,
pollution research. Acton , MA. Public Science Group Inc. , American Medical Association. , 1974 , pp. 193-212.
-391 一
7. 铁 7.1 概述
就重量而言，铁是地壳上第四位最丰富的元素。在水中 主要以二价和三价(亚铁和正铁〉的形式存在(1)。铸铁管和
钢管均用于给水系统，而且铁盐也用作生产自来水的絮凝 剂 (2)。
7.2
存在
在地面水中，铁通常以正铁 (Fem) 形式存在。在良好 曝气的水中很少会发现高浓度的铁，但是在还原的条件下，
如某些地下水、湖或蓄水池中以及没有硫化物和碳酸盐存在 时，就可能出现高浓度的可溶性亚铁离子 (3)。已有报道，在
地下水中铁浓度高于 1 毫克/升 ω。在未然水中的铁可能来 源于岩石和矿物的溶解、酸性矿的排水、填埋垃圾的渗漏、 污水或与铁有关的工业。
一般而言，在大气中存在着低浓度的快，这是由于钢铁 工业、热电厂和焚烧炉的废气排放所致，但是几乎没有关于
大气中铁浓度的资料。 食物中铁的含量很不相同。谷类(平均为 0.0295 毫
克/克〉和肉 (0.0262 毫克/克〉可能是该元素的主要膳食来 源、 E53 ，大多数其他天然食物中铁浓度低于 0.020 毫
克/克白， 7) 。而在补充铁的食品或用铁容器烹调的食品中浓度 -392 一
可能稍高 (8)。有证据表明，煮沸时食物中铁含量降低 (9)。
7.3 15........22 毫克白， 9 ， 10) 。
接触途径
在发达国家，从典型膳食中每日铁的摄入量估计为 饮水中铁的浓度一般低于 0.3 毫克/升，从食物中的摄
入量明显高于饮水。 从空气中吸入的铁与总摄入量相比，是微不足道的。
7.4
对健康的影响
铁是人的必需营养素 (7 ， 1 1)。存在于很多在生物学上非常
重要的蛋白质中，例如血红蛋白和细胞色素以及很多氧化 还原酶中。估计每日最低需要量波动在 7........14 毫克，这取决 于年龄和性别，任振妇女可能需要 15 毫克/日以上 (12)。平均
每日需要量为 10 毫克。 个体对铁的需要量(取决于年龄、性别和生理状态〉调 节着从膳食中铁的吸收量，该量波动在 1........20
%(1 2 , 13)。大多
数人约能吸收摄入量的 10% (7 ， 12) 。而从粪、尿和汗中的丢 失量为 1 毫克/日 (12)。在吸收的铁中 60........70% 用于合成血红
蛋白 J 5% 用于合成JUl红蛋白，其余部分主要贮存于肝、骨髓 和脾中。 大量摄入铁引起已知的血色素沉着症(正常调节机制未 能有效地起作用)，导致铁的蓄积而损伤组织。这种情况很少 由简单的膳食过量而引起吼叫 1 4)，但是却曾因长期使用铁制
炊具烹调酸性食品而发生过(1 4) 。在接受极高剂量铁的实验 动物中未看到血色素沉着症 E153. 而小孩摄入大量含铁片剂 后曾发生中毒 El63.
-393-
7.5
其他方面
从许多与健康无关的方面来看，在给水中存在铁是令人 厌恶的队 9， 17) 。在给水 pH 值的情况下，亚铁盐不稳定，并
以不溶性氢氧化铁作为铁锈泥沉下来。这种水常常昧道不佳，
而且使衣服和家中固定设施着色。沉积在配水系统的铁使水 流速度逐渐减慢，而且也促进了"铁细菌"的生长。这些细
菌能使亚铁氧化为正铁，在这个过程中，在营道上沉积了一 层粘滑的内壁。 当铁的浓度接近 0.3 毫克/升时，特别是在配水系统中，
通常出现上述问题，因此应尽可能使铁的浓度低于该水平。 参考文献
1. Quality criteria lor water. Washington , DC , US EnvironmentaI Protection Agency , 1976. 2. Cox , C. R. Operation and control (WHO Monograph Series No. 49).
01
water treatment
processes. Geneva. World Hea It h Organization , 1964 3. Hem , 1. D. Chemical lactors that inlluence the
auailability 01 iron and manganese in aqueous systems. Washington , DC , The GeologicaI Society of America Inc. , 1972 (SpeciaI Paper 140) p. 17. 4. Dart , F. J. The hazard
01
iron. Ottawa , Water and
Po lI ution Control Canada , 1974. 5. Méranger , J. C. & Smith , D. C. The heavy metaI content of a typical Canadian diet. Canadian journal
01 -394 一
public health , 63'53 (1 972).
6. Gormican , A. Inorganic elements in foods used in
hospital menus. J ournal
01
the American Dietetic
Associ ati on ，
目:
397 (1 976).
7. Watt , B. K. & Merr iIl, A. L. Composition
01
lood s-raw , processed , prepared. Washington , DC , Revised USDA Agriculture Handbook , Vo l. 8 (1 963). 8. Bowering , J. & MacPherson Sanchez , A. A conspectus 。f
research on iron requirements of man. J ournal (1 976).
01
nutrition , 106:985
9. Zoeteman , B. C. J. & Brinkmann , F. 1. 1. Intake of
minerals by man. In: Hardness
01
drinking water
and public hea !t h. Proceedings of the European Scientific Co Il oquium , Luxembourg , 1975. Oxford , Pergamon Press , 1976. 10. Kirkpatrick , D. C. & Coffin , D. E. The trace metal content of representative Canadian diets in 1970 and
197 1. Canadian Institute Technology journal , 7:56
01
Food Science and
(1 974).
1 1. Moore , C. V. Iron. In , Goodheart , R. S. & Shils , M.
E. , ed. , Modern nutrition in health and disease. Philadelphia , Lea & Febiger , 1973 , p. 297.
12. Dietary standard lor Canada. Ottawa , Department of National Health and Welfare , 1975. 13. Hopps , H. C. Ecology
01
disease in relation to
environmental trace elements-particularly iron. Washington , DC , The Geological Society of America Inc. , 1972 (Special Paper 140, p. 1). 14. Jacobs , A. Ir on overload-clinical and pathological aspects. Seminars in hematology , 14:89 (1 97 7). 15. Brown , E. B. et a I. Studies in iron transportation and metabolism. journal
01 Iaborator y
and c1 i ni cal -395-
medicine , 12'862 (1957). 16. Stein , M. et al. Acute iron poisoning in children.
Western journal 01 medicine. 125:289
(1 976).
17. Iron.ln: McKee , J. E. & Wolf , H. W. , ed. , Water
quality criteria. Sacramento , CA , California State Water Quality Control Board , 1971 (Publication 3-A ,
p. 202).
'‘
-396-
8. 锺 8.1 概述
空气中瞌的来源广泛。而排放在空气中的握主要与较小 尘粒结合，这一事实支持握的分布范围是很广的(1)。
在天然地面水中，辑是以海解的和悬浮的两种形式存在
的。而在厌氧的地下水中，常含有较高浓度的溶解性筐。
8.2
存在
在不同地区的谈水中可能含有 1 到数千微克/升的锤二 在加拿大、联邦德国、英国、美国和苏联的各种湖泊和河流 中，瞌的浓度范围为 1~600 微克/升左右(1月。在自由流动
的河水中有时发现高浓度的钮，这通常与工业污染有关。在 地下水以及某些湖泊和蓄水池中可能存在的还原条件，有利 于出现很高浓度的握。
在非工业区大气中，锚的平均浓度为 0.05 微克/立方米， 而在工业区浓度达 0.3 微克/立方米。 食物中锚的含量很不相同。已发现在乳制品 (O.O~ 1. 9 毫克/公斤〉、肉 (0.0~0.8 毫克/公斤〉和鱼 (O.O~
0.1 毫克/公斤〉中含有低浓度的髓，在谷类(1 .2~30.8 毫 克/公斤〉、坚果 (0.4~35 .1 毫克/公斤〉和蔬菜 (0.2--
12.7 毫克/公斤〉中含有高浓度的瞌(1)。在茶叶中含极高浓
度的毡，一杯茶可能含有1. 4~3.6 微克。 ,
-397-
8.3
接触途径 6 个月以内
辑是通过吸入以及从食物和饮水中摄入的。最大的来源 是食物，估计成人摄入量为 2.0--8.8 毫克/日 J 的婴儿为 2.5--25 微克/公斤体重/日(1)。在北美平均每日摄 入量为 3.0--4.1 毫克〈平均 3.6 毫克) (3侧530 为维持正常生 理功能，估计平均每日链的需要量为 3--5 毫克 (6)。
从饮水中锚的摄入量可能很不相同，正常情况下明显低
于从食物中的摄入量。现有资料表明，从饮水中的摄入量通 常小于 0.1 毫克/日，但是也可能高一个数量级 (7‘ 9) 。
、估计在非工业区居住的非职业接触人群每日从大气中吸 入的瞌为 2--10 微克 (9)。
每日摄入 3--7 毫克瞌对一个 70 公斤体重的人将增加身 体负荷 12--20 毫克 (11 ， 14)。
8.4
对健康的影响
瞌是动物和人的必需元素，但是仅能吸收摄入量的大约
3% (10)。在很多酶系统中需要辑作为辅因子，它在维持黄素 蛋白的良好功能和合成硫酸化粘多糖、胆固醇、血红蛋白以 及在很多其他重要的代谢过程中均起作用(1 1) 。目前已经发
现，饮水中钮的含量与心血管病死亡率呈负相关 (12)。吸收的
钮很快离开血流而蓄积在肝中，并与胆汁盐结合。锚的生物 半减期相当短 (10)。肝脏对握的贮存能力是有限的，约为
~1-- 1. 3 毫克/公斤〈湿重〉。在动物中，由实验引起的或天然 发生的钮缺乏均出现各种症状 E13 ，虽然还未见报道人因瞌缺 乏而引起特殊的症状，但是已表明懂的缺乏可能与某些紊乱 〈例如贫血、儿童的骨路改变(1 3) 和红斑狼疮)有关。 398 一
辑被吸收的主要途径是经呼吸道和胃肠道 E1530 由于键
在胃液中洛解度低，通过胃肠道仅能吸收经口给予剂量的
3.......4% (14) 0
锚的吸收与铁的吸收有密切关系 (16) 。贫血能
导致对铁和链的吸收量增加，贫血病人对瞌的吸收增加 2 倍
多。对锚的吸收与膳食中钙的含量有相反的关系 E53 ，而与佣 的含量直接相关 (17) 。
体内的组主要是通过排世而不是通过吸收与排泄两者来 进行调节的。某些远通过膜脏的分泌排泄 F 某些则直接通过 肠壁排泄，仅很小量(占 0.1.......2%) 经尿排出。磕被认为是 毒性最低的元素之一。给家兔、猪和家畜剂量为 1.......2 毫克/ 公斤体重的瞌进行慢性经口实验，结果除见到食欲改变和由 铁生成血红蛋白的代谢过程减慢之外，未见其他毒性作 用 (11) 。
除一次孤立的事故外，未见饮水造成辑中毒的报告。
1941 年在日本曾把一例脑炎样疾病的原因归咎于井水受到 污染，该井水中髓的浓度为 14 毫克/升，而其他金属(特别 是铮〉的浓度也很高，因此未能明确肯定该病的发生仅仅是 由于高浓度握引起的(1 8) 。在日本的另一个区域，饮水中远
的浓度为 0.75 毫克/升，未见对用户健康具有明显的有害作 用 (19) 。
经胃肠外慢性给家兔毡，引起精小管的明显变性，导致 不育 (20) 。但是与此相反，给小鼠少量钮可预防铺对辜丸引 起的坏死作用 El82，在另一项研究中见到铜与握具有协同作
用。 尚无证据证实钮是致癌的。而相反，数项研究表明，提 可能有抗癌作用。 Spivey FOX(2 1) sl 用的最初报告指出，芬
兰人的癌症发病率高与在地理上含有低水平易溶性握的土壤 -399 一
有关-已报道，撞对氨基偶氮染料在致癌过程中的代谢具有 抑制作用 (18) 。
在每日钮的摄入量如下表所示的情况下，未见对人体健 康的有害作用 (5) ， 均值(毫克)
范围(毫克〉
人厌恶的。瞌浓度超过 0.15 毫克/升使饮料味道不佳 j 并且 使家庭的固定设备和衣服着色 (22) 。当溶解的二价憧氧化时，
瞌就会沉淀下来形成结垢。甚至当浓度约为 0.02 毫克/升时，
撞将在管道上形成一层内壁，并可能作为黑色沉淀脱落下 来 (23) 。钮也能促使某些令人讨厌的微生物生长 (22 ， 24) ，这些
微生物能富积髓，在配水中引起味、臭和 j虫度等问题。 根据瞌引起着色的特性，推荐在饮水中的建议值为 0.1
毫克/升。显然，这是一种折衷的数值。为避免着色问题，水 中握的浓度应尽可能低。 参考文献 ι 1.
2. National Water Qualify Data Bank. Ottawa , lnland Waters Dircιora， te ，
-
400 一
食水空
从与健康无关的很多方面来看，在给水中存在远也是令
物气
3.000 0.005 0.002
2.0-7.0 0.0- 1. 0 0.0-0.029
8.5
其他方面
M anganese , Gencva , World Hcalth Organization , 1981 (Environmental Hcalth Critcria , No. 17) Water Quality Branch.
Environment , Canada , 1976.
3. Kirkpatrick , D. C.
& Coffin , D. E. The trace metal
content of representative Canadian diets in 1970 and
197 1. Canadian Institute of Food Science and Technology journal , 7:56 (1 974). 4. Méranger , J. C. & Smith , D. C. The heavy metal content of a typical Canadian diet. Canadian journal
of public health , 63:53 (1 972). 5. Schroeder , H. A. et a l. Essential trace metals in man , manganese. A study in homeostasis. J ournal of chronic
diseases , 19:545 (1 966). 6. Kay , H. O. Micro-nutrient elements-a recapitulation.
Journal of food technology , 2'99 (1 967). 7. US Environmental Protection Agency. Scientific and
technical assessment re port on manganese. Wash ington , DC , Office of Research and Development 1975 (Report No. 8. Craun , G. F. EPA-600/6一 75-002).
& McCabe ,
L. J. Problems associated
with metals in drinking water. Journal of the
American
Wate俨 Works
Association , 67:593 (1 975).
9. Guide 1i nes for Canadian drinking-water quality. Quebec , Supply and Services , 1980 (supporting documentation) . 10. WHO Technical Report Series , No. 647 , 1980 品
(Recommended health-based limits on occupational exposure to heavy metals: report of a WHO Study Croup) . 11. National Research Council (Committee on Medical and Bi ological Effects of Environmental Pollutants). M anganese. Washington , DC , National Academy of Sciences, 1973. 一 401 一
12. Masironi , R. International studies on trace elements in the etiology of cardiovascular diseases. N utrition
reports international , 7 1 51 (1 973). 13. Pier , S. M. The role of heavy metals in human health. Texas reports on biology and medicine , 33' 85 (1 975). 14. Mena , J. The role of manganese in human disease.
Annals of clinical and laborator y science , 4' 487 (1 974) .
15. Rodier , J. Manganese poisoning in Moroccan miners.
British journal
0/ industrial medicine , 12 '21 (1 955).
16. Mena , J. et a l. Chronic manganese poisoning.
N eurology. 19.1000 (1 969). 17. Underwood , E. J. Trace elements in human and animal
nutrition , 3rd ed. New York , Academic Press , 197 1. 18. Enuironmental hea1t h criteria programme for manganese and its compound s Oapanes"e Report). Geneva , World Health Organization. 1974. 19. Suzuki , Y. (Environmental contamination hy manganese.) J apanese journal 12'529~533
0/ industrial health ,
(1970).
20. Chandra , S.
& Tandon , S.
K. Enhanced manganese
toxicity in iron-deficient rats. Enuirontnental
physiology and biochemistry , 3.230 (1 973). 2 1. Spivey Fox , M. R. In. Lee , D. H. K. , ed. M etallic
contaminants and human health. New York , Academic Press , 1972. 22. Griffin , A. E. Significance and removal of manganese in water supplies. Journal of the American Water Works Association , 52' 1326 (1 960). -402-
23. Bean , E. L. Potable water-quality goals. J ournal 01
the American Water W orks Association , BB: 221 (1 974).
24. Wolfe , R. S. Microbial concentration of iron and manganese in water with low concentrations of these elements. Journal 01 the American Water Works
Association , 5211335 (1 960).
-403 一
9. 溶解氧 9.1 概述
水中溶解氧主要是对包括铁、钮、铜以及含氮和硫的化 合物等物质的氧化-还原反应产生作用。在一定的配水系统
中，溶解氧的浓度可能随时间的延长有降低的趋势。尽管这 种改变通常是表明腐蚀过程，但是也可能是由于管道内有机 质(特别是沉积物〉中的细菌呼吸作用所致。所以当水中铁
浓度无明显增加的情况下溶解氧可能降低。 30 相反，由于腐 蚀作用可能使水中含有高浓度的铁，然而几乎未引起溶解氧
的耗竭。
9.2
与其他水质指标的关系
在铁的腐蚀过程中，只需要相当少量的氧，因此
2Fe = 2Fe H + 4e 4e + 2H 2 0 + O 2 = 40H1 毫克/升的氧将产生 3.5 毫克/升的亚铁，因此发生大 量铁腐蚀时，可能几乎未察觉到溶解氧的改变。 当水中溶解氧耗尽时，就会发生厌氧腐蚀过程，包括硫
酸盐-还原菌的活动，因而将疏酸盐还原为硫化物。
8H+ + 8e + SOr 一 404 .，....，
=S2- + 4H 20
当溶解氧消耗到低于饱和溶解氧的 80% 时，往往导致更 多的用户抱怨，特别是对昧、臭和水的颜色的不满(1)。
饮水中氧的消耗经常伴随着其他问题。在厌氧条件下，
细菌将硝酸盐还原为亚硝酸盐 EIU 将硫酸盐还原为硫化物， 往往引起臭的问题。 缺氧的水可能腐蚀性不是特别强，但是因为腐蚀产物常 常对管壁的粘着性较小，所以更可能引起用户抱怨水有色。 在厌氧条件下，整个配水系统的溶解性亚铁浓度可能增加。
例如，在贮存罐和水塔中与空气接触后，水中氧浓度增加将 造成不溶性三价铁沉淀，从而导致距发生问题很远地方的水 中也出现颜色 ω。
含溶解氧低的配水有很多缺点。建议在配水系统中的水 应该始终含有适当浓度的溶解氧，但是难以推荐一个建议值， 因为这一可接受的水平还受水中的其他成分影响。
参考文献 1. Ridgway , J. et a l. Water quality changes-chemical and microbiological studies. In: Water distribution
systems: maintenallce 0/ water quality and pipeline integrity. Medmenham , England , Water Research Centre , 1979. 2. Hall , E. S. & Smith , J. G. RU5ty water cured by oxygen injection. Water services , 78: 941 (1 974) •
• ← 405 一-
10. 10.1
pH
值
概述
溶液的 pH 值以氢离子活度 (aH+) 常用对数的负值表 刁飞 z
pH = -loglo
(aH 吵
在稀释液中，氢离子活度约等于氢离子浓度。 一个水样的 pH 值通常使用玻璃电极进行测量。，2) 。而 温度对 pH 的测定具有明显的影响口， 2) 。
10.2
影响 pH 的因素
水的 pH 是各种可洛性化合物所达到的酸-碱平衡程度 的指标，在大多数天然水中，它是通过二氧化碳-碳酸氢盐 碳酸盐平衡系统来维持的 (3)。该系统包括各种组成成分的平
衡，所有成分均受温度的影响。在纯水中，温度每升高 25"(; ， pH 值约降低 0.45(4) 。在用碳酸氢盐、碳酸盐和氧氧离子构 成具有缓冲能力的水中，就会改变这种温度的影响(4)。 大多数原水水源的 pH 值在 6.5~8.5 的范围内 (5) 。
水处理时可能使氢离子浓度发生明显改变。水的氯化可 能使 pH 值降低，而使用过量石灰/苏打灰进行水的软化时， 能使 pH 值增加.
-406-
10.3
与腐蚀、结垢和其他 水质指标的关系
总水管和水处理厂的腐蚀问题可能是一项巨大的经济负
担阳。除腐蚀外，由于引起碳酸钙沉淀，造成配水容积减少 及水泵费用的相应增加 (7) 。
用于配水系统的金属，例如铁、钢和铜，因其热力学的
不稳定性，在与水接触时易被腐蚀。而通常用于配水系统的 混蛙土、石棉-水泥和内衬水泥的铸铁管也会发生变质。天 然水含有气体、肢质以及各种电解质和非电解质，与 pH 值 综合在一起就决定了对一个系统的腐蚀程度 E83 ，并且也决定
了水的"侵蚀"性。水中存在的阴离子与金属形成可溶性化
合物时，就会增加水对那种金属的腐蚀性，而阴离子与金属 形成不溶性化合物时则可能增加金属的钝性。 Drane 总结了 pH 在用于配水系统的金属腐蚀中的作 用 (8) 。
碳酸钙的沉积可能抑制腐蚀。影响这一过程的因素是温
度、 pH 、溶解性总固体、硬度、二氧化碳和碱度。然而在 实际情况下，控制碳酸钙/碳酸氢钙的平衡极为困难，甚至 是不可能的。因此，需使用易于测定的指标，为此已经提出 了许多半经验和经验方程式。最广泛使用的是由 Langelier 提出的方程式 (4 ，的。
改变 pH 和碱度的防腐蚀效果取决于碳酸盐/碳酸氢盐 平衡系统的精确平衡。确定达到平衡的水，即碳酸钙处于稳 定状态。由于不能使碳酸钙沉淀，所以通常这种水对铁和钢
均育腐蚀性。另一方面，除非经过适宜的处理，否则过饱和 的水将会形成大量的结垢。这种水垢能否抑制腐蚀取决于其
-
407 一
多孔性和对金属的粘着力的。
过硬的水通常不会引起严重的腐蚀问题，但是却易于造 成大量结垢。投加石灰/苏打灰进行软化处理的硬水 pH 可达 10.9 ，因此易于形成结垢(1 0) 。用再碳酸化的方法，加二氧化
碳使 pH 到 9.t7~10 ，或每升水加 0.25-0.5 毫克多磷酸纳， 可以达到稳定状态 E102 ，已经建议再进行破酸化，使 pH 达
8.6 ，以便使水稳定， 酸钙 E1130
从而防止在配水系统中产生过量的碳
在配水管表面的生物粘着物能防止氧化产物脱落和氧透 过管壁，因此抑制腐蚀作用。另一方面，生物粘着物的过量 生长能产生二氧化碳，从而使管道的局部表面产生低 pH 区， 造成局部腐蚀，而大量的水仍具有良好的稳定性和侵蚀指 标 (12) 。铁细菌的生长与 pH 关系极大，当 pH 为 5.5-8.2 时能生长，最适 pH 约为 6.5 (1 3) 。红水的出现往往由于铁
细菌突然大量的生长，产生代谢终产物氢氧化铁所致。在适 宜条件下，铁细菌生长极为迅速，以致于能在数周内使水营 严重堵塞。
通过数种不同的途径，
pH 几乎与每项水质指标有关，
因为水的化学平衡总是包括氢〈和氢氧〉离子。 硫化氢气体的形成，可能是硫污染所致，能使水产生
"臭蛋"睐，这种情况在热力学上常发生在 pH 小于 7.0 时 C14) 。在水的氧化过程中，当 pH 小于 7.0 时易于形成大 量的三氯化氮，产生令人厌恶的刺激气味肘， 16) 。水中 pH 过高时会产生苦味C! 7) 。 增加 pH 值可使水的色度增加 E183 ，这种现象一般称为
吁旨示效应"。因此，为控制水质而测寇色度时，建议应在 标准 pHS.3 的条件下进行 C19) 。 -408 一
混凝过程的效果与 pH 有密切关系，因此在水处理时必 须调节 pH ，以形成最佳的絮状物 (20 ， 2 1)。在某些情况下， pH 对过滤效果也有很大影响 E2230 大多数细菌通常能耐受一般水源的 pH 范围 (13 ， 23 ， 24) 。
水的细菌学质量与 pH 值有关， pH 能影响消毒的效果。在 pH 值高的情况下，水中氧的杀菌效果较低， 这是由于随着 pH 的升高而过氯酸浓度降低 (25-27) 。在饮水中所见到的 pH
范围内，作为另外的消毒剂臭氧和二氧化氯的效果不变(1 6) 。
在给水系统中的腐蚀作用是饮水中金属污染的一种主要 来源 (28) 。铅和铺是其中的两种最麻烦的金属。在纯水中 pH
6 以上时，铅能耐腐蚀。当碳酸盐和破酸氢盐存在时，
pH
为 4 ......， 12 之间，铅是钝性的。但是 pH 大于 12 则发生腐蚀 作用 (29) 。向使用铅设备的居民供应含低碱度和低 pH 的水 会导致出现高浓度的铅 E3030 在纯水中， pH 为 9......， 13.5 时铺
是钝性的，根据实验资料，仅当 pH 低于 6 时才出现明显腐 蚀 E2930
10.4
对健康的影响
不太可能确定人体健康和饮水 pH 之间的直接关系，因
为 pH 与水质的其他方面密切相关。 Taylor 及其同事 (3D对给水进行流行病学调查时，把
pH 作为考虑的指标之一，
他们未能确定病毒性肝炎 A 的发
生率与出厂7]< pH 之间具有任何明显的相关，但是，必须给
予正确估价，因为该项研究主要是针对 pH 对消毒的作用。
就 pH 对水处理(用于去除病毒、细菌和其他有害微生 物〉中各个过程的影响而言，可以认为 pH 对健康有间接作 用. -409 一
对 pH 的建议值为 6.5----8.5 ，然而认识到当 pH 低于 7 时，在配水系统中可能出现某些问题。 参考文献 1. pH value. Ina Standard methods /or the
examination
0/
water and waste water , 14th ed.
Washington , DC , APHA , AWWA , WPCF , 1976 , p. 460. 2. Standard method of test for pH of water and waste water.1na Annual book Materials , 1976 , p. 178. 3. Goldman , J. C. et a l. Water research , 61637 (1 972). 4. Langelier , W. F. Effect of temperature on the pH of natural waters. Journa/ 5. Webber ,
0/
ASTM standards , Part
31. Philadelphia , American Society for Testing and
0/
the American Water
Works Association , 38:179 (1 946). W. 上， Jr. & Stumm , W. Mechanìsm of hydrogen ion buffering in natural waters. Journal
0/
the American Water Works Association , 55: 1553 (1 963) .
6. Hudson , H. E. , Jr. & Gilcreas , F. W. Health and economic aspects of water hardness and corrosiveness.
JQurnal
0/
the American Water Works (1 976).
Associ ation , 68: 201
'/. McClanahan , M. A. & Mancy , K. H. Effect of pH on quality of calcium carbonate film deposited from moderately hard and hard water. Journal
0/
the (1 974).
American Water Works Association , 66:49
8. Drane , C. W. N a.tural w-8. ters-..Inl Shreir , L. L. , ed.
Corrcsion , 2nd ed. London , Newnes-Butterworths , 410 一
Chapter 2. 9. Langelier , W. F. Chemical equilibria in water treatment. J ournal 0/ the American Water W orks
Associ ati on , 38: Hi9 (1 946). 10. Dye , J. F. & Tuepker , J. L. Chemistry of the limesoda process. In. Water qual it y and treatment , 3rd ed. Toronto , McGraw-Hill , 1971 , p. 313. 1 1. Sawyer , C. N. & McCarty , P. L. Residual chlorine and chlorine demand. Ir.. C hemistr y /or sanitar y
engineers , 2nd. ed. Toronto. McGraw-Hill , 1967 , p. 363. 12. O'Connor ,
J.
T. et a l. Deterioration of water quality
in distribution systems. J ournal 0/ the American
Water Works Association , 67:113 (1 975). 13. Shair , S. Iron bacteria and red water. lndustrial
water ellgineering , March-April. 16 (1 975). 14. Pourbaix', M.Atlas
0/
electrochemical equilibria in
aqueous solutions , 2nd ed. Houston , National Association o{' Corrosion Engineers , 1974 , p. , 545. 15. American Water Works AS50ciation Research Foundation. Handbook 0/ taste and odor control
experiences in the US and Callada. Denver , CO , AWWA 1976. 16. Morris , J. C. 'Chlorination and disi n.fection-state of the árt. Journal 0/ the American Water Works
Assóci atjon , 63: 769 (1 971). 17. Statement 0/ basis and purpose /or the national sècondary drinking water regulations. Washington , DC , US Environmental Protection Agency , 1977. 18. Black , A. P.
& Chri&tman , R. F. Characteristics of
coloured surface waters. JouTnal 01 thll American
Water Works Association , 55:753 (963). 19. Singley , J. E. et a 1. Correction of color rneasurernents to standard conditions. J ournal 01 the Ameri can
Water W orks Associ ation , 58 :4 55 (1 966). 20. Sawyer , C. N. & McCarty , P. L. Chernical coagulation of water. Inl Chemistry lor sanitary engineers , 2nd ed. Toronto , McGraw-Hi Il, 1967 , p. 34 1. 2 1. Maudling , 1. S. & Harris , R. H. Effect of ionic environrnent and ternperature on the coagulation of color-causing organic compounds with ferric sulfate.
Journal 01 the American Water Works Association , 60:460 (1 968). 22. Cornrnittee Report. Coagulation-fi It ration practice as related to research. Journal 01 the American Water
Works Association , 66=502 (1 974). 23. Davis , B. D. et a l. Microbiology , 2nd ed. New York , Harper and Row , 1973 , pp. 92-93. 24. Rudolfs , W. et a l. Literature review on the 。 ccurrence
and survival of enteric, pathogenic , and
relative organisms in so i1, water , sewage , and sludges , and on vegetation. Sewage and industrial wastes , 22 =1261 (1 950). 25. ButterHeld , C. T. et a 1. Influence of "p H. and temperature on the survival of coliforrns and enteric pathogens when exposed to free chlorine. Public 26. Srnith ，明七百T
health reports , 58=1837 (1 943). & Bodkin , R. E. The in fI uence of
hydrogen ion concentration on the bactericidal action of ozone and chlorine. J ournal 01
bacteriology , 47: (A17) 445 (1 944). 27. Scarpino , P. V. et a l. A comparative study of the inactivation of viruses in water hy chlorine. Water
research , 6:959 (1972). 28. Craun , G. E.
& McCabe ,
L. 1. Problems associated
with metals in drinking water. J ournal 29. Pourbaix , M. Atlas
0/ the
American Water Works Association , 67:593 (1 975).
0/ electrochemical equilibria in aqueous solvtions , 2nd ed. Houston , National
ÁS8ociation of Corrosion Engineers , 1974 , pp. 488-49 1. 30. McFarren , E. F. et a l. Water quality deterioration
in the distribution system. Kansas City , MO , Water Quality Technology Conference , 1977. 31. Taylor , F. B. et a l. The case for water-borne infectious hepatitis. American journal
0/
public
heallh , 56 :2093 (1 966).
-413-
11. 纳 1 1. 1 概述
近几十年来地面水和地下水污染的增多已在世界上不同 地区引起了饮水中纳含量的明显增加。水厂的处理过程以及 家庭饮水的软化也造成了饮水中纳含量的增高。 硫酸纳用于生产色素和染料，纸浆和造纸工业以及很多 其他的现代化生产活动，其废水中含有高浓度的制 E130
在很多国家为融化公路上的冰雪，一次使用大量的氯化
锅，其用量 E 在逐步增加。氯化制也用于生产苛性锅、氯和 很多工业化学物。还大量用于食品加工、屠宰和肉类包装， 乳、鱼、谷类及酿造工业 (2) 。
1 1. 2
存在
纳是碱金属中最丰富的元素，其化合物广泛分布在自然 界，在地壳中的含量为 26 克/公斤。
土壤中含铀量为 1-10 克/公斤，主要存在于硅酸盐矿 石，例如闪石和长石中。
一些地下水含有高浓度的纳，在某些情况下，这能造成 河流和小溪中盐度增加。 在高速公路加盐融化冰、雪也能增加土壤中铀的负荷。 径流的水足以能污染公共给水，估计公路上所用盐量的 25~ 50% 渗入地下水中 (3) 。 -414 一
污水、工业废水、沿岸区域海水的流入以及使用铀化合
物控制腐蚀和水软化过程均可使水中的纳浓度升高，因为制 盐和矿物质的溶解度高。水中铀的浓度很不相同，这取决于 地区和局部的水文和地质情况、季节和使用盐的类型。地下 水中铀的含量差异很大，但正常范围为 6.......130 毫克/升 E42，
浓度高可能与盐碱土壤有关.在地面水中，铀浓度可能小于 1 毫克/升或大于 300 毫克/升，这取决于地理位置 (5- 7)。
据报道，在公共给水中铀的浓度小于 1 毫克/升至大于 1000 毫克/升(7)。水的软化处理能明显增高给水中铀的浓 度 (8) 。
铺天然存在于所有食物中，但不同的食物含量很不相同，
食品加工对锅含量有明显影响。鲜豌豆约含 9 毫克制/公斤。 而在干的罐头豌豆申却为 2.3 克/公斤，在冰冻豌豆中为 1 克/公斤 (9) 。 新鲜水果和蔬菜含制为小于 0.01 克/公斤至约 1 克/公斤，而谷类食物和奶酶可能含者 10......20 克/公斤 (10) 。
11. 3
接触途径
人体对铀的每日摄入量差别很大，这是由于食物中饷含 量不同以及个人在使用盐作为食品调味剂方面的差异。已发
现在 3833 人的→个样本中， 45% 的男性和 30% 的女性常规 向自己的食物中加盐 (11 ， 12'0 20......64 岁的加拿大男性铀的每 日平均摄入量估计为 3600 毫克 (10) 。与从食物中的摄入量相
比，从饮水中盐的每日摄入量仅占很小的比例。
1 1. 4 对健康的影响 a 锅是最丰富的细胞外阳离子，制及与其相结合的阴离子 ·关于饮水中纳含量对健康影响的详细讨论自曹见第三章第 16 节。
-415 一
一起主要构成了细胞外浓的渗透活性。 水和电解质的平衡是通过从食物和水的膳食摄入与从 尿、粪、汗和呼出气的排泄来维持的。一个 70 公斤体重的 正常人约含 69 克具有代谢活性的制和 45 升水。 水和纳平衡的调节是通过包括神经系统和内分泌系统在
内的一系列相互联系的复杂过程实现的。该平衡是由铀的排
泄而不是控制从肠道的吸收来维持的。调节制排泄的最重要 因素是盐皮质激素，即醒固酣。 由于身体能很有效地调节铀的排泄，所以纳不是一种具
者急性毒性的金属。据报道，给 14 名婴儿误用含制浓度为 21140 毫克/升的水后引起 6 名死亡 (13) 。锅中毒的症状一般
累及中枢神经系统，表现为敏感性增加。有一些研究表明， 铀的摄入量与人的高血压呈正相关 E142 ，而其他研究未证实 这一点 (15 自 17) 。
一般认为，铀的每日摄入量在 1600~9600 毫克范围内 对正常人的健康无有害作用 (18)。为保证膳食总摄入量为 500
毫克/天，需要将饮水中制含量限定到约 20 毫克/升，从而为 了采用现有的处理技术必须花费额外的水处理费 (19)。
为确保大多数用户感觉水无昧，水中的盐含量应接近于 唾液。唾液内铀的平均含量为 300 毫克/升，但是也许超过 该值 2 倍。
铺在水中的昧阔取决于与其相结合的阴离子和温度。碳
酸纳味阔最低，而碳酸氢锅则最高。在室温下，铀的各种味 阔值如下: NazCO.. 为 20 毫克/升， NaCl 为 150 毫克/升， NaNO 坷为 190 毫克/升， Na 2 SO. 为 220 毫克/升，
NaHC0 3
为 420 毫克/升。 对铀的建议值为 200 毫克/升，这是根据上述的味阐值， aa· < nv 吨'且'
而不是根据对健康的影响。 参考文献 1. K ilI in , A. F. Sodium sulphate. In. Canadian rninerals
yearbook. OUawa , Department of Energy , Mines and Resources , 1974. 2. Killin , A. F. Salt. In , Canadian minerals yearbook. Ottawa , Department of Energy , Mines and Resources , 1974. <1. McConne lI, H. H. & Lewis , J. Add salt to taste.
Environrnent , 14:38 (1 972). 4. Bond , R. G. & Straub , C. P. Genetic types of subterranean waters in relation to their salinity. Inl
Handbook
0/
environrnental contro/. Vol. 3. Water Co 叶 1973 ，
çupply and treatrnent , 1st ed. Cleveland , OH , Chemical Rubber p. 85. Proceedings of the 5. Weiler , R. R. & Chawla , V. K. Dissolved rnineral
quality
0/ Great Lakes waters.
12th Conference on Great Lakes Research , Ann Arbor , MI , 1969 , p. 80 1. 6. Dobson , H. H. Principal ions and dissolved
oxygen in Lake
Onta 俨 io.
Proceedings of the 10th
Conference on Great Lakes Research , 1967 , p. 337.
7. Sodiurn , chlorides and conductivity in drinking ω ater.
Report on a WHO working group. Copenhagen ,
WHO Regional Office for Europe , 1979 (EURO Reports and Studies , No. 2). 8. E lI iou , G. B. & Alexander , E. A. Sodium from drinking water as an unsuspected cause of cardiac ，他
decompensation. Circulation , 23:562 (1 961). 9. Meneely , G. R. A review of sources and the toxic effects of excess sodium chloride and the protective effect of extra potassium in the diet. Qualitas
plantarum. Plant /oods /or human nutrition , 23:3 (1 973) .
10. Gormican , A. Inorganic elements in foods used in hospital menus. J ournal
0/ the Ameri can Dietetic
Association , 561397 11. Statement
(1 970).
0/ basis and purpose /or the national
interim primary drinking water regulation , Washington , DC , US Environmental Protection Agency , 1975. 12. National Heart and Lung Institute. The public and
high blood pressure. Washington , DC , US Department of Health , Education and Welfare , 1973 (Publica tion N o. (NIH) 74-356). 13. Finberg , L. et a l. Mass accidental poisoning in infancy. Journal
0/
the American Medical (1 963).
Association , 184:187
14. Dahl , L. K. Salt and hypertension. American journal 15.
0/ clinical nutrition , 251231 (1 972). Kerkendall , W. M. The effects of dietary J. & Koiw ，且，
sodium on
the blood pressure of normotensive man. Inl Genest , ed. H y pertension , Heidelberg , Springer-Verlag , 1972 , p. 360. 16. Evans. J. G. & Rose , G. Hypertension. British medical bulletin. 27 137 (1 97 1). 17. Dauber , T. R. et a l. In. Stamber , J. et a l.,
Environmental factors in hypertension. New York , auz
Grune and Stratton Inc. , 1967. 18. Dahl , L. K. Possible role of salt intake in the development of essential hypertension. In , Cottier , P.
& Bock , K. D. , ed. Essential hypertension , an i nternati onal symposi um. Heidel berg , Spr ingerVerlag , 1960 , p. 53. 19. National Academy of Sciences and National Academy of Engineering. Water qua/ity criteria , 1972. Washington , DC , US Government Printing Office , 1974.
-419 一
12. 硫酸盐 12.1 概述
除铅、顿和摆的硫酸盐外，大多数硫酸盐溶于水(1)。溶
解性硫酸盐被认为是水中的一种永久性溶质，但是，它可被 还原为硫化物，以硫化氢形式挥发到空气中，以不溶性盐的
形式沉淀或掺入活的生物体内 ω 。 硫酸盐可随很多不同工业的废物排入水环境中(1， 3) 。大
气中的二氧化疏 (S02) 是在矿物燃料燃烧中形成的，并且
可在金属培烧过程中排放，这些二氧化硫也可能是地面水中 硫酸盐的来源。三氧化昧 (S03) ，是二氧化硫在光解或催化 氧化巾产生的，与水蒸气结合形成疏酸，以"酸雨"或雪沉 降下来 (3) 。
12.2
存在
在大多数谈水中硫酸盐的浓度很低，而在美国东部、加 拿大和欧洲的大部分地区普遍的浓度为 20~50 毫克/升 (4 ， 5)。
据报道，在 23 个欧洲共同体的大城市公共给水中，硫 酸盐平均浓度为 64 毫克/升(范围为 9~125 毫克/升) (6)。根
据在 5 年(1 969~1973) 多的时间内，在英格兰、苏格兰和
威尔士大约供应 60% 人口的约 600 个水源中所获得的资料，
在英国给水中硫酸盐的浓度范围为 4~303 毫克/升飞硫酸 • Powell 引证的数值。
-4哩。一
铝广泛用作水处理的混凝剂，可使出厂水的硫酸盐含量增加 20-50 毫克/升。常规水处理不能从水中去除疏酸盐。在 欧洲共同体市场上的瓶装矿泉水中硫酸盐浓度均值为 223 毫 克l升(范围 0-1182 毫克/升)(的。
12.3
接触途径
尚缺乏关于硫酸盐的每日膳食摄入量资料。在美国疏酸 盐用作食品添加剂，估计美国人从这一来源的每日平均摄入 量为 453 毫克 (7 ， 8) 。
每日从饮水中摄入的硫酸盐含量变动极大，特别是在饮 用瓶装矿泉水的情况下。
12.4
对健康的影响
硫酸盐不易经人的肠道吸收 E82 ，它缓慢地透过哺乳动物 细胞膜，并且迅速地经肾被排出 E930
据报道，硫酸镜对哺乳动物的最小致死量是 200 毫克/ 公斤体重 E1030 硫酸盐剂量1. 0-2.0 克对人有导泻作用，引
起胃肠道腹泻。飞婴儿摄入相当于 21 毫克/公斤/天的硫酸
盐也会导致这种效应。疏酸镜浓度高于 1000 毫克/升对正常 人有腹泻作用，显然，低于此浓度在生理学上是无害的 (1 ， 10)。
敏感的人在硫酸镜浓度低至 400 毫克/升时就有反应。在浓度
高于 700 毫克/升时，新用户或偶尔饮用的人受到影响。但 是人能在-定时间内适应饮水中较高浓度的硫酸盐 (11 )。
最常见硫酸盐的味阔浓度是 z 硫酸纳 200-500 毫克/ 升，硫酸钙 250 ......， 900 毫克/升 F 升 0 ， 1])0
硫酸镇 400"""'600 毫克/
主要基于上述数值，以及硫酸盐的导泻作用，对饮水中 -421 一
硫酸盐的建议值为 400 毫克/升。
12.5 蚀，特别是低碱度的水。
其他方面
饮水中高浓度的硫酸盐可能引起配水系统中金属的腐
参考文献 1. McKee ,
J.
E.
& Wolf ,
H. W. Water quality
criteria , 2nd ed. Sacramento , CA. California State Water Quality Control Board , 1963 , pp. 136 , 213 , 247 , 270 , 275-277. 2. National Research Council , Drinking t且later and
health , part [. Washington , DC , National Academy of Sciences, 1977. 3. Delisle , C. E. & Schmidt , J. W. The effects of sulphur on water and aquatic life in Canada. In:
Sulphur and its inorganic derivatives in the Canadian environment. Ottawa , National Research Council of Canada , 1977 , pp. 227-284. 4. Hitchcock , D. R. Biogenic contributions to atmospheric sulfate levels. Proceedings
0/
the Second National
Con/erence on Complete Water Re-use. Chicago , American Institute of Chemical Engineers , 1975. pp. 291-310. 5. Katz M. The Canadian sulphur problem. In
,
,
Sul phur
and its inorganic derivatives in the Canadian envirollment. Ottawa , National Research Council of Canada , 1977 , pp. 21-67. 6. Amavis , R 叫 et a l., ed. H ard ness 422 一
0/
d rinking water
and public health. Oxford , Pergamon Press , 1976 , pp. 176-199. 7. Subcommittee on Research of GRAS (Genera Il y Recognized as Safe) Li st (Phase 11). Food
ingredients , Washington , DC , National Academy of Sciences, 1972 (DHEW No. FDA 70-22). 8. Novikov , Yu. V. & Erisman , F ,. F. The 'potable water' standard (GOST 2874-73). A new stage in development of water hygiene. Vestnik Akademii
Meditsinskikh Nauk SSR , No. 3, 59 translation No. 3, 76 (1 975).
(1 975). English
9. Senning , A. Sulfur in organic and inorganic
chemistry. Vo I. 2. New York , Marcel Dekker Inc. , 1972 , p. 160. 10. Arthur D. Li ttJe , Inc. Inorganic chemical pollution
0/ freshwater. Washington , DC , US Environmental Protection Agency , 1971 (Water Po Il ution Control Research Series No. D PV 18010). 1 1. Zoeteman , B. C. J. Sensory assessment
0/ water
quality. Oxford , Pergamon Press , 1980.
一 423 一
13. 昧和臭 13.1 13.1.1 概述
味
就"昧"这个字的严格定义而言，水中的味是唾液和水
中溶解的物质之间相互作用，由味蕾感受器所接受的感觉。 口腔内约有 3000--10000 个昧营，大部分在舌的上表面、舌
尖、边缘和后表面。 当尝水"昧"时，味觉和嗅觉均活跃起来，极难加以辨 别。因此，昧和嗅的联合效应常归为"昧"。 味觉远不如嗅觉敏感口， 2) 。但是当把明显无嗅的水放入
口中时可能产生令人厌恶的"昧"。这可能是由于口腔温度 高，水中洛解的有机物逸入鼻腔所致。可见，在这种浓集的
形式下，嗅觉才察觉到水中存在的溶质，因此，尝水的味道， 与仅仅闻水的气味相比，往往是评价水质的更敏感的感 觉 E330
在评价饮水水质时，昧和嗅是相辅相成的。一般而言， 味觉对检查水中无机成分更有用，而嗅觉对有机成分更有用。 昧的测定往往受到明显的批评 α-6) 。在味阔测定中，使
用蒸馆水作为标准"无味"水和漱口水会对结果带来偏差。 由于味觉对含盐浓度高于或低于唾液的水均可感觉到与唾液 不同，所以保持最低浓度的离子〈如纳、氯化物、钙和碳酸 424 一
氢钙〉对于制备无味水是必要的。昧的测定涉及到一大批实 验室的评定者，所以费时，并且在小水厂难以进行 ω，因此 建议使用一组用户作为评定者 (3)。在水厂进行昧的测定与输
送至用户时相比可能低估了水昧，因为在水厂时，余氯可能 掩盖令人厌恶的昧，而在配水系统中因余氯降低，这种掩盖 作用就会减少 (4) 。 在评价水昧前，水的化学脱氯会增加水
昧 (4) 。用户评定小组在家庭中评价水味可以克服这些问
题 ω 。另外的困难是来自于测定方法的不同，例如稀释水的 质量、评定小组成员的数目和动机以及对测定资料的统计处 理和解释的差异。 测定水味强度的适宜方法类似于嗅的测定.
13. 1. 2 存在 给水中臭和味的问题能引起大量用户的抱怨，这对任何
类型的水以及一年中的任何时间均可能发生。某些问题是由 天然因素造成，而有些则是人类的工业活动所致。问题的产 生主要与原水、水处理方法、配水系统或三者的综合作用有
关。 加拿大、荷兰和美国已经发表了对水处理厂所进行的臭 和味的调查结果 (3 ， 7) 。一般来说，地下水源昧的问题最小。
多数地面水臭和昧的问题有季节变异，表明该问题可能是由 生物原因引起的。
高色度和出度的水常与非特异性味(和臭〉的问题有 关 (8) 。
据报道，昧的强度取决于温度 (9 ， 10) ，而温度对昧的影响
程度与特定的产昧物质有关 (10) 。较高温度能促进微生物的
生长速度，而其中某些微生物能产生具有令人厌恶味道的代 一 425 一
谢物，而增高温度也能增加令人厌恶味道产物的形成速度。 由于 pH 能控制溶液中物质以中性和离子型的平衡浓 度，所以 pH 能明显影响味〈和臭)。 虽然几年来已就余氯对感宫的影响进行了数项研究，但 是在关键问题尚未解决之前仍需进一步研究。在理想的情况
节，用户水龙头水中游离余氯的含量应高到足以确保水在细 菌学方面的安全性，但又要足以低到能避免产生令人厌恶的
昧和臭。余氯的味阔和嗅阔是很重要的，最新的研究发现， 当 pH 从 5. '0增至 9..0 时，游离余氯的平均味阔浓度从 0.075 毫克/升增至 0.450 毫克/升 (11) 。当 pH7.(} 时平均昧阔浓度
为 0.156 毫克/升，范围为 0.02~0.29 毫克/升。 尽管不需要作更多的工作，但是大多数水处理厂的经理
都知道如何恰当地权衡余氯浓度和用户的抱怨问题。对矿泉 水 (12)和咖啡中 (3) 氯的昧阔测定结果表明，水中其他的产昧
成分能影响氯的闹浓度。因此，原水的性质是影响检测余氯
味阔浓度的一个重要因素。已知在某些国家，为确保用户用 本的安全性，要求水中存在轻微的氯味。 水中存在许多无机化合物，当其浓度远低于急性毒性作
用浓度时就会使水产生不佳味道。因此依据毒性对这些物质 所规定的限值浓度会使用卢感到不快。下面将对这些物质作
简要讨论。 已报道，饮水中的主要阳离子〈例如钙、锐、制和锦} 在蒸馆水中的味阔浓度约分别为 100 、 .30 、 100 和 300 毫克/ 升 (3 ， 14) 。这些数值可能发生变动，这在很大程度上取决于与
阳离子结合的阴离子对昧的影响。 对铁〔如 Fe(ll)) 的昧阔测定表明，评定组中 5% 的
最敏感成员能察觉在蒸馆水中的浓度为 0.04 毫克/升，但是 -426 一
在矿泉水中港解性总固体含量为 500 毫克/升时，铁的昧阔 浓度为 0.12 毫克/升 (15) 。在蒸馆水中铮浓度为 4.3 毫克/升 时可以被察觉，而在矿泉水中却为 6.8 毫克/升 E1530
关于水中硫化物昧闹和嗅阔的可靠资料很少，这是由于
pH 对硫化物-二硫化物-硫化氢平衡是有影响的，所以情况 有些复杂。
13. 1. 3
对健康的影响
在公共给水中存在令人不快的味道就可能迫使用户去寻
找另外的水源，但该水源的细菌学质量可能不同于原来的结 水。加利福尼亚州公共卫生局的调查就是一个例子，他们发
现有意见的用户大量购买瓶装饮水。在荷兰，由于水气味不 佳，造成了自来水消费量的减少 (3) 。很遗憾，无味道的水并
不能保证其中无病原菌或无有毒的无机化合物。但是幸运的
是无机物昧阔中值通常远低于引起有害健康作用的浓度。 在公共给水中，正常水昧的短期改变可能是原水水源水 质改变的信号、水处理不充分、或在配水系统中发生化学腐 蚀和生物生长的结果。 给水的目标是向大多数 (90%) 用户提供没有不良味道
的水。实现这个目标的最直接途径是定期向一个选定的用户 组征求意见。实验室评定组可用分级的方法〈即，好一未能察 觉一弱一令人厌恶一坏)评价水昧或用强迫选择法(见 13.2.1 节〉评价味值。在后一种情况下，建议把饮水的味值保持在
1 以下。用户评定组最适于评价龙头水的水昧，而实验室的 评定者对于评价水处理过程的水味更有价值。 在特殊的地区，水中可能出现不可避免的味。在这种情
况下，地区卫生当局应该首先进行消毒，以确保控制致病污 一 ~27 →
染物，如致病菌。 参考文献 1. Rosen , A. A.
& Booth , R. L. Taste and odour contro l. In , Water quality and treatment , 3rd ed Toronto , McGraw-Hill , 1971 , p. 225.
2. Suffett , J. H. & Sega lJ, S. Detecting taste and odour in drinking water. Journal
0/
the American
Wate γ
Works Association , 63:605 (1 97 1). .3. Zoeteman , B. C. J. Sensory assessment
0/
water
quality. Oxford , Pergamon Press , 1980. 4. Baker , R. A. Dechlorination and sensory contro I. ,
Journal
0/
the American Water Works Association.
56: 1578 (1 964). 5. Bruvold , W. H. Human perception and evaluation of water quality. CRC critical reviews in
envÎ ronmental control , 5: 153 (1 975). 6. Swets , J. A. Is there a sensory thresholdy Science , 134: 168 (1 91)1).
7. Handbook 0/ taste and odour control experiences in the US and Canada. Denver , CO , American W a ter W or ks Associa tion , 1976. 8. Riddick , T. M. Zeta potential and polymers. JOllrnal
0/ the American Water Works Association , 58:719 (1 966) .
9. Standard methods/or the examination 0/ water and Z且laste water , 14th ed. Washington , DC , American Public Health Association , 1976 , p. 12 1. 10. Pangborn , R. M. & Bertolero , L. L. Influence of 一 428 一
temperatl.lf e :on taste intensity and degree of -1i lking of drinkillg water. J ournal 01 the Ameri can Water
Works Association , 64:511 (1972). 1 1. Bryan , P. E. et al. Taste thresholds of h Jl og~ns
in
water. Journal 01 the American Water Works
Associ ation , 65: 363 (1 973). 12. Pangborn , R. M. et a l. Sensory examination of mineralized , chlorinated waters. J ournal
0/ the
American Water Works Association , 62:572 (1 970). 13. Campbell , C. L. et a l. Effects of certain chemicals in water on thé flavour of brewed coffee. Food
research , 23: 575 (1 958). 14. National Academy of Sciences. Water quality
criteria 1972. Washington , DC , US Government Printing Office , 1973 (EP A-R-73-033). 15. Cohen , J. M. et a l. Taste threshold concentrations of metals in drinking water. J ournal
0/ the American
Water Works Association , 52:660 (1 950).
13.2 13.2.1 概述
臭
"饮水的臭"定义可认为是由于存在具有一定熏气庄的 物质刺激了在鼻腔和鼻窦中的感觉器官而引起的感觉。一般 而言，引起嗅觉产生反应的浓度〈数微克/升或更低 1 比味觉 (数毫克/升或更高〉要低得多。 水的嗅强度通常以嗅阔值 (TON) 表示，即用无嗅水 稀释水样，当一组评定者在严格控制的条件下刚能嗅出气味 一 429 一
时，水样稀释倍数的几何均数叫"嗅阔值" (1)。正象昧阔测定
的情况一样，实验室评定组对水质的定级或平均阔值仅仅是
对全部用水人口的估计值。飞由于水的性质不同，也许嗅阔
值为 2 的水比嗅阔值为 4 的水会招来更多用户的抱怨 ω 。 另一种测定嗅阔值评价嗅强度的方法是强迫选择法，这
种方法具有某些优点 ω。该方法是给每个评定组成员一系列 成对的样品，其中一个是呈一定稀释度的受试样品，另一个
是无嗅水对照。每个成员必须判定每对样品中哪一个样品具 有较强的气味，即使未察觉到差别，也必须做出选择。对每
一个稀释度计算正确回答的百分数，并以 50% 机率进行校 正，因为随机判定时获得正确应答的机率为 50% 。嗅值表示
为校正后的正确应答率为 50% 时的稀释度，它是以稀释度和
与其对应的校 E 的应答率作图进行计算得到的。 Zoete man(3) 巳详细地论述了这一方法。使用费时较少的"间断 分度法(i nterval
scaling
method)" 也获得了良好结
果，这种方法最适用于测定嗅强度很低的饮水。
嗅强度的测定通常是非特异性的。但是，对水中特定物 质的强度测定一般采用嗅阔浓度表示 E42 ，即 50% 的评定组成 员能嗅出其气昧的物质浓度。使用嗅阔浓度测定法所得的结
果清楚表明，个体间对嗅的检测阔差异很大。在一个大的人 群样本中， 5% 最敏感的人能可靠地嗅出气昧的浓度为平均嗅 阔浓度的 1% (5) 。因此，为获得可靠的资料，应该由大量的
评定组人员对水质进行嗅评价。为了更方便起见，也可能使
南较少的评定组人员进行评价，但是必须认识到这样就降低 了测定的正确性和可靠性。
注明测定嗅强度时的温度是很重要的，因为嗅强度与任 何产臭物质的蒸气压有关，因此直接与水温有关。 430 一
13.2.2 存在 水的气味，主要是由水中的有机物引起的。已报道，在 水中存在很多产臭化合物白， 7) 。
饮水中令人厌恶的气味可能来源于坐物或工业，而某些 天然的气味可能间接来源于人的活动，例如污水直接排入水 环境能加快生物的生长，而形成产臭产物。
天然的气味，一方面常常被描述为土昧、霉味或酸味， 另一方面为鱼昧、草味或黄瓜样气昧，包括象 geosmin 和 哭醒类化合物的气味 (8叫 l 飞来自工业的气味象石油或杂酌
油或药味。这方面的典型例子是案以及氯苯和氯盼 E530 地下 水的气昧问题较小，尽管气昧问题既不局限于任何单一类型
的水，也不局限于任何特殊的季节。在配水系统流速慢的部 分或原水和出厂水贮水池中，当水流停滞的情况下也可能产
生气昧问题。在水净化过程中也可能使弱臭物质(如股和盼〉 转化为强臭物质〈如氯股和氯盼〉 E1230 令人厌恶的生物，如 铁细菌和硫细菌，在配水系统中的增长，也可能是气味的来
源。 非特异性的鱼昧，草昧和霉味一般与生物生长有关，最 常发生在一年中暖和季节的温暖地面水中 (10 ， 1 1)。
加拿大和美国关于水昧和臭的调查已经鉴别出 50 种引起 水臭的所谓"令人厌恶"的生物。由放线菌类生物产生的强
烈霉昧物质可能是公共给水中气味污染的主要来源。已经建 议监测原水中的放线菌类 (1 ， 10) 。且然在某些地区通过这种监
测发现放线菌类与气味的问题呈正相关，但是在布些水处理 厂则友现相反的情况。
-
431 一
13.2.3
对健康的影响
自来水出现气味几乎毫无例外地表明了某些形式的水源 污染或水处理以及配水过程出现了故障。生物来源的气味表 明生物活性增加，可能包括在给水系统中危险病原体的负荷
增加。工业来源的气昧与工业废物污染水源水有关，其中一 些物质可能有毒。卫生调查应包括调查潜在的和现有的产生
气味的来源，并应努力鉴别引起气昧的现有来源。 因毒性而令人担心的某些化学污染物也能引起气味问 题。例如，据报道水中氧氢酸的嗅阔为 0.001 毫克/升 (6) 。
基于该资料饮水中氟化物的含量应是本"准则"建议值的百
分之一〈见第一卷，第 55 页)。在这个和其他例子中，嗅觉 比现有最好的分析仪器更敏感。可能除氯丹外，水中农药的
气味太弱，以至于在建议值水平或以下是不能被检出的。 理想的情况下，任何用户不应感觉饮水具有明显的气味。
但是，由于人群中各人的嗅阔差异很大，所以比较现实的目
标是为大多数人(即 90%) 提供无令人厌恶气昧的水。达到 这一目标的最直接途径是与供水区域中由大量用户组成的评
定组〈即 100 个用户〉合作，要求他们在家中定期评价水的 臭和味。参加评定的成员应该指明水的级别〈即，好一未察 觉臭和味一中度不佳一坏〉。实验室的评定者在评价水的臭和
味中起到更关键的作用。但是，由经过训练的 10........20 人组 成的实验室评定组也能指出是否一种水在感官上可为大多数 用户所接受 (3) 。如果采用强迫选择法，在室温下由实验室评
定组人员测定嗅值时，除非局部情况要求进行淌毒并需感觉
到存在余氯之外，建议嗅的目标值小于 10
-
432 一
参考文献 1. Standard methods /or the examination AWWA , WPCF , 1976 , p. 75. 2. Baker , R. A. Dechlorination and sensory control.
0/ water and wastewater , 14th ed. Washington , DC , APHA ,
Journal
0/
the American Water Works
Association , 56: 1578 (1 964). 3. Zoeteman , B. C. J. Sensor y assessments
0/
water
quality. Oxford , Pergamon Press , 1980. 4. Baker , R. A. Threshold odors of organic chemicals.
Journal 0/ the American Water Works Association , 55:913 (1 963). 5. Zoeteman , B. C. J. & Piet , G. J. Cause and identification of taste and odour compounds in water. Science (1 974) .
0/
the total environment , 3:103
6. Van Gemert , L. J. & Nettenbreijer , A. H. , ed.
Compilation 0/ odour threshold values in air alld water. Voorburg , National Institute for Water Supply; Zeist , Netherlands , Central Institute for Nutrition and Food Research , TNO , 1977. 7. Stahl , W. H. , ed. Compilation
0/
odor and taste
threshold values data. Philadelphia , American Society for Testing and Materials , 1973 (ASTM Data Series Publication No. DS 48).
8. Handbook 0/ taste and odour control experiences in the US and Canada. Denver , CO , American Water Works Association , 1976 , p. XIV-l. aaτ
su 。
9. Zoeteman , B. C.
J. & Piet , G. J. On the nature of
odours in drinking water resources of the Netherlands. Science (1 972/73) .
0/ the total environment ,
1'399
10. Morris , R. L. et a 1. Chemical aspects of Actinomycetes metabo 1i tes as contributors of taste and odour. Journal 11. McKee ,
0/ the American Water Works
Association , 55: 1380 (1 963).
J. E. & Wo 1f, H. W. , ed. Water quality CA~
'crite r'i a.2nd ed. Sacramento , 3- A).
California State
Water Quality Control Board , 1963 (Publication No. 12. Burttschell , R. H. et a 1. Chlorine derivatives of phenol causing taste and odour. Journal
0/ the
American Water Works Association , 51.205 (1 959).
-434-
14 14.1
温度 概述
一般而言，化学反应的速度随着温度的降低而降低。在
化学平衡中反应物和生成物的相对浓度也随着温度而改变。 因此，温度可能对于饮用水的处理和运送的各个方面都有影 响。
14.2
物理学方面
冷的饮用水比温的更可取。在室温下水的味强度最大， 通过致冷或加热的方法均可明显降低水的味强度。增加温度 也能增加水中微量的挥发性化合物的蒸气压，从而引起嗅的 增强。 由于混凝效果明显地随温度而变化，所以浊度和色度也
间接地与温度者关系。当温度增高时，对混凝作用的最适 pH 降低(1)。因此，为了最经济地使用混凝剂，小型试验应该在 处理水的温度下进行，而不是在室温下进行 (2) 。 温度降低时，水的粘滞性增加，沉淀和过滤的速度降低。
因此在冬季的温度条件下混凝、沉淀和过滤对去除颜色和浊 度的效果要比夏季差。在较低温度的条件下，通过过洁、去除
烛度的效果降低，这可能是由于絮凝力降低或平均粒子减小 所致 E330 温度对话性碳的过捷、作用也有影响 p 当温度降低时， 活性碳的吸附作用增强 (4) 。 ← 435 一
14.3
微生物学方面
饮水的微生物学特性与温度有关，因为温度影响水处理
过程〈特别是消毒)以及微生物的生长和存活。 一般而言，增加温度有助于消毒。 Butterfield 及其 同事发现，在温度 20.......25 0C 时氯对埃希民大肠杆菌的消毒效 果比 2.......5 'C时提高 5 倍 (5) 0 Ames 和 Smith 在美国军队
的研究表明，温度从 8 'C至 40 0C 消毒效果提高 9 倍 (6)oCha mbers 报道，在 pH 7.......8.5 之间温度对氯消毒效果的影响
不明显 E 而当温度为 4 'C至 22 C 时在较高 pH 的条件下，消 毒效果提高 4.......8
0
1'1} (7) 。对于病毒已获得类似的结果 (8) 。据
报道，臭氧对分枝忏菌 (fortuitum) 的灭活效果随温度 的增加而增加 (9) 。
水的混凝和沉淀能降低悬浮微生物的数量，正如前面所 讨论的那样，温度能影响这些过程。 在一定的 pH 条件下，提高温度能增加次氯酸的离解， 但是，这种作用对加氯消毒杀菌效果的影响程度小于提高温
度对杀菌效果的增强作用。
尚不能就已发表的水温对细菌存活影响的资料得出肯定 结论(1 0) 。虽然观察到原水水源中大肠杆菌数随季节变动 E113 ，
但是，温度只是造成这种变动的许多因素之一。 在低温下，病毒的存活时间比细菌要长得多，据报道， 在低温下小儿麻痹病毒能在龙头水中存活 6 个月 (12) 。但是，
在美国 13 个城市对病毒性甲型肝炎的流行病学研究表明，感 染率和原水温度之间没有关系 (13) 。
温度增高能缩短寄生蠕虫包囊相卵在水中的存活时间。
例如，血吸虫卵在 29--32 'C时 9 天内死亡， 15--24 "c时 3 436 一
周内死亡，在 7 0C时 3 个月死亡 04 飞
温暖的水能使令人厌恶的生物生长速度加快，从而导致
产生不佳的嗅和味。
14.4
化学方面
在加氯的饮水中三卤甲皖的形成速度随着温度的增加而 增加 E152，也许这是影响三卤甲皖浓度随季节变动的最主要因 素(1 6) 。
研究温度对水处理系统的腐蚀作用证实，腐蚀作用的增 加与温度呈函数关系 (17) 。用氢氧化纳调节 pH ，在相同温度
范围内能使腐蚀作用的增加减半。但是温度低于 10"C 时，含 氢氧化铀的水比未处理的水有较高的腐蚀率。腐蚀率也是水
中溶解氧浓度的函数。但是与上述腐蚀率的明显变化相比， 榕解氧随温度的变化是小的，所以溶解氧在温度对腐蚀的影 响中不起重要的作用。 碳酸钙的溶解产物随温度的改变而降低，但是在 碱度 的情况下〈碳酸钙为 50 毫克/升) ，实际上碳酸钙的溶解度 随着温度的升高和 pH 降低而增加。对饱和指数的这种影响 有助于减轻碳酸钙的结垢作用，同时增加了水的腐蚀性，从 而导致对热水系统的腐蚀性增加 (3) 。
参考文献 1. Maudling , J. S. & Harris , R. H. .Effect of ionic environment and temperature on the coagulation of color-causing organic compounds with ferric sulfate.
Journal 01 the American Water WOI.ks Association , 60:460 (1 968). -437-
2. Camp , T. R. et a l. Effects of temperature on rate of floc formation Journal 01 the American Water
Works Association , 32:1913 (1940). 3. American Water Works Association. Water quality σ nd
treatment , 3rd ed. Toronto , McGraw-Hill , 1971 ,
pp. 89 , 305. 4. Weber , W.
J. &
Morris ,
J
C. Equilibria and capacities
f or adsorption on carbon. J ournal 01 t he S anitar y
Engineering Division , Proceedings 01 the American Society 01 Civil (1 964) . Enginee 俨 s ，
90: (5A3) 79
5. Butterfield , C. T. et a l. Influence of pH and temperature on the survival of coliforms and enteric pathogens when exposed to free chlorine. Public
health
repo 俨 ts
58=1837 (1943).
6. Ames , M. 7.
& Whitney-Smith , W. Journal 01 bacteriology , 47:445 (1 944). Chambers , C. W. An overview of the problems
of in
disinfection. S ymposium on wastewater
t γ eatment
cold climates. Saskatoon , Canada , University of Saskatchewan , 1974 , p. 423 (EPS 3-WP-74-3). 8. White , G. C. Disinfection: The last line of defense for potable water. Journal 01 the American Water
Works Association , 67=410 (1 975). 9. Farooq , S. et a l. Influence of temperature and UV light on disinfectÎon with ozone. Water research 11 =737 (1 977). 10. Rudolfs , W. et a l. Literature review on the occurrence and survival of enteric , pathogenic , and relative organisms in soil , wateI , sewage , and sludges , and on 438 一
vegetation. Sewage and ind ustrial wates , 22: 1261 (1950). 1 1. Rao. S. S.
& Henderson , J.
Summar y report 0/
microbiological baseline data on Lake Superior 1973. Ottàwa , Environment Canada , lnland Waters Directorate , 1974 , (Scientific Series No. 45). 12. Health and Welfare Canada. M icrobiological qualit y
0/
drinking water. Ottawa , Health and Welfare
Canada , 1977, (77-EHD-2). 13. Taylor , f.. B. et a l. The cuse for water-borne infectious hepatitis. American Journal a/ public
health , 56.2093 (1 966). 14. Temperature. ln , Water quality criteria , 2nd ed. Sacramento , CA , California State Water Quality Control Bpard , 1963, p. 283. 15. Stevens. A. A. et a l. Chlorination of organics in drinking water. Jρ urnal
0/ the American Water
W orks Association , 68'615 (1 976). 16. Sm ilI ie , R. D. et al. Organics in Ontario drinking
water , Part II. Toronto , Ontario Ministry of the Environment , 1977. 17. Mullen. E. b. & Ritter , J. A. Potable wíiter corrosion contro l. Journal
0/
the Americèm Water
Works Association , 66'473 (1974).
-439 一
15. 溶解性总固体 15.1 概述
水中溶解性总固体 (TDS) 包括无机盐和少量有机物。 构成 TDS 的主要离子是碳酸盐、碳酸氢盐、氯化物、硫酸
盐、硝酸盐、制、何、钙和墟。飞 TDS 对于水质的其他方 面，例如昧、硬度、腐蚀性和结垢等性质均有影响。
15.2 业废物的排放。
存在
水中 TDS 可能来自于天然来源，污水、城市径流或工 与花岗岩、硅砂、沥滤良好的土壤或其他相对不溶性物 质接触的水中溶解性总固体的含量低于 30 毫克/升 (2) 。前寒 武纪地质区的水中含量通常低于 65 毫克/升 (3) 。古生代和中
生代沉积岩区域的水中 TDS 含量较高，其范围从 195 至 1100 毫克/升 (3)，碳酸盐、氯化物、钙、握和硫酸盐是存在 的主要离子。， 4)。除天然沥滤过程外， 污水和工业废物能引
起 TDS 的进一步增加二在干旱的情况下，小提流中 TDS 浓度能增至 15 克/升 ω 。据记载，在其他地方的海水中 TDS 含量超过 35 克/升 (3) 。
在某些国家，冬季时用盐融化公路上的冰雪而造成地面
水和地下水的污染，致使水中 TDS 的含量明显增加。 1969----1970 年冬季， -440 一
丹麦在公路网上使用了 203000 吨盐，
1974 年加拿大使用了 250 万吨盐 J 1970 年美国将 900 万吨 盐用于路面除冰，所有这些全部的盐均构成了溪流、航道和 地下水给水中的 TDS 含量 (5) 。
15.3 对健康的影响 尚无证据表明， TDS 含量高于 1000 毫克/升的饮水引 起饮用者出现有害的生理反应 α， 4 ， 6) 。而某些流行病学研究
的结果却显示出饮水中的 TDS 可能对健康具有有益的作 用。 一般认为，通常的溶解性矿物盐类影响水昧 (7-11)。关
于很多具有味道的矿物质已在有关昧的一节〈见 424 页〉中 进行了单独的论述。根据水中 TDS 的含量， 等 (8)将饮水味道良好的程度进行了分级 s 极好s 好a
Bruvold
低于 300 毫克/升 300-600 毫克/升 600-900 毫克/升 900-1200 毫克/升
尚好， 不好 s
不可接受.
高于 1200 毫克/升
但是， TDS 含量极低的水也可能是不被接受的，因为这 种水淡而无味。
15.4
其他方面
TDS 中的某些成分，例如氯化物、硫酸盐、钮、钙和 碳酸盐影响配水系统的腐蚀或结垢 ω 。
常规水处理厂一般不能去除溶解性总固体。
虽然据报道水中 TDS 含量高于 1000 毫克/升时未见引 起有害的在理学作用，但是通常认为超过该值是不可接受
-
441 一
的，故建议值为 1000 毫克/卉。 参考文献 1. Quality criteria for water. Washirrgton , DC , US Environmental Protection Agency , 1976 (EPA-4401 9-76~023)
.
2. Rainwater , F. H. & Thatcher , L. L. Methods for
collection and analysis of water samples. Geological Survey Water-Supply Paper , Washington , DC , US Government Printing Office , 196Q. 3. Garrison Investigative Board. Water qualit y 俨 eport
(Appendix A). Garrison Diversion Study , Report to the International Joint Commission: US-Canada , Windsor , Ontario , 1977. 4. Durfor , C. J. & Becker , E. Constituents and properties of water. In: Pettyjohn , W. 丸，
ed. Water
qualit y in a stressed environment. Minnesota , Burgess Pub Ii shing Company , 1972. 5. Sodium , chlorides and conductivity in drinkingwater supplies. Copenhagen , WHO Regional Office for Europe , 1979 (EURO Reports and Studies , No. 2). 6. Ongerth , H. J. et a l. The taste of water. Public health
reports. 791351 (1 964). 7. Bruvold , W. H. & Pangborn , R. M. Rated acceptability of mineral taste in water. Journal 01 applied
psychology , 50:22 (1 966). 8. Bruvold , W. H. et al Consumer attitudes toward mineral t 且 ste in domestic water. Journal 01 the
American Water Works Association. 59:547 (1 967). -442 一
9. Bruvold , W. H. Scales for rating the taste of water.
lournal 01 applied psychology , 52 1 245 (1 968). 10. Bruvold , W. H. Mineral taste and the potability of domestic water. Water research , 4:331 (1 970). 1 1. Bruvold , W. H. & Ongerth , H. J. Taste quality of mineralized water. lournal 01 the Americα n Water
Works Associafion. 61: 170 (1 969).
,,
一 443 一
16. 浊度 16.1 概述
水中油度是由悬浮物质(例如泥土、淤泥、肢体有机颗 粒、浮游生物和其它的微小生物〉所形成的。浊度是表示水 样特定的散射光和吸收光的特性。在很大程度上，该指标的
意义取决于测定技术。?昆 j虫水散射光线的总强度和角度分布 代表了粒子内和粒子间相互作用的总效应，它是通过复杂的 方式由诸如外来颗粒的数量、大小、形状和折射指数等因素 以及入射光线的波长所决定的。虽然这些因素是复杂的，但 是仍可以得出很多一般性的规律口， 2) 。
有五种方法可用于测定水的浊度，然而只有其中的两种 方法，即散射出度法 (nephelometry )和比 1虫法 (tu rbidimetry) 是目前标准法的基础 (3-6) 。
在历史上，测量废水和饮水的浊度是使用 Jackson 的 烛光浊度计进行的 (7) ，即在特制容量管中，透过样品垂直观
察燃烧着的标准蜡烛，根据刚好能鉴别蜡烛影象时样品的深 度凭经验定量。 Jackson 浊度单位 (JTU) 以样品深度表 示， 21.5 厘米的深度相当于 100JTUC5 ， 6) 。 由于 Jackson
烛光出度计仅适用于测定浊度高于 25JTU 的水，因此不宜 用于监测饮用水。经过改进的仪器，例如 Patterson 浊度 计 (3) ，使用电灯光源和镜片，能测定较低的 j虫度。作为取代
Jackson 一 444 一
j，虫度单位的一种方法，是用能产生一起 j虫度(容量
概念)的悬浮固体浓度(毫克/升〉对温度计进行校正。通常 使用硅薄土制备标准混悬液。这类表示法(有时称为漂白土 分度法〉是人为的，并且因所用特殊泥土的类型和颗位大小 不同而异 E830
散射 j虫度法是目前用于测定浊度的方法 EM-1130 它是测
定以 90 .向入射光通路上散射的光线强度。对这种 j虫度计设
计的差异将会造成所测温度的不同。 为了尽量减少差异，对光源设备、检测器的几何学要求 以及校正方法均进行了特殊的限制 (6) 。水厂几乎均采用 fo
rmazin 聚合物混悬液作为温度参比标准 (7) 。在特定的条件 下，将硫酸胖 (50 毫克/升〉与六甲股四胶 (500 毫克/升〉
反应形成的 formazin 7.昆悬液规定为 40 散射烛度单位
(NTU)
(6 ， 10) ，也称为 formazin 浊度单位 (FTU) 。当
使用烛光温度计测定时，该标准混悬液的浊度约为
40 JTU(6) 正如上述方法所定义的，浊度是非特异性的测定悬浮固 体浓度。目前已有垃子电功计数器，能正确计数和记录悬浮 颗粒。一般而言，出度(范围为 0.2.........1 散射浊度单位〉与颗 粒数有关，但是两种方法之间并不存在点与点间的良好一 致 (11 )。
16.2
存在 泥土、植物
形成水中浊度的颗粒大小的范围从胶态(约 10 毫微米〉 至直径为 0.1 毫米。这些颗粒一般可分为三类 z 和动物碎片分解而形成的有机颗粒和纤维颗粒，即石棉矿物 质 (12) 。通常泥土颗粒大小的上限约为 0.002 毫米直径。
在太多数天然水中的大部分悬浮物质是由土壤表面的受
- 445 •
侵蚀形成的土壤颗粒构成的。粗砂和 1M 泥是全部或部分地被 有机物所包裹。叶硅酸盐 (phy llosilica te) 泥土颗控以及 非泥土物质(例如铁和铝的氧化物和氢氧化物、石英、无定 型硅胶，碳酸盐和长石等〉均构成泥土成分(1 2) 。泥土和有 机颗粒也常一起构成"泥土-有机物"复合物(1 2) 。腐殖质比 无机泥土具有更高的离子交换能力 (13) 。在很多情况下，以腐
殖质成分的作用占优势。 由于大量微生物蓄积所造成的有机浊度甚至可以使水混
浊而看不清。例如夏季时在地面水中蓝-绿藻的生长，藻类 碎片以及在配水系统中铁细菌的碎片〈造成红水现象〉就是 如此(14)。
原水浊度的变动范围从低于 lNTU 至高于 1000NTU 0
用简单过滤的方法或更有效地综合使用混凝、沉淀和过惊、相
结合的方法均可去除浊度。 通过砂滤床或其它单一介质滤器进行过捷、可以使浊度达 到 1 或低于 lNTU 。在水处理的全部过程中连续监测出度有
助于达到这一水平。
16.3
与其他水质指标的关系
水中 i虫度与饮水水质的许多其他指标有关或影响这些指 标。颗粒物质也可能是营养物的来源或能保护某些微生物。
有证据表明，水中很大一部分颜色来源于胶体颗粒，其 中 50% 的颜色是由于腐殖质的"肢体成分"所产生的 (15) 。 因此，真色系指从水中去除 j虫度后的颜色(l的。
长期以来就认识到，无论在原水和过滤、的水中，高油度
均与水昧和颜色有犬，而且给水中的颗粒物会佳用户感到厌 恶、 (17) 。
-446-
浊度对饮水的微生物学质量有明显的影响，能使对于细 菌和病毒的检测复杂化。在水中颗粒物的表面，以及天然产
生的怯散的或在混凝过程中形成的絮凝物内部，微生物能大
量地生长〈见第一章"微生物学指标")。由于营养物吸附在 表面上，能促进细菌的生长，所以贴附在表面的细菌比在混 悬液中游离的细菌生长的更迅速(1 8 ， 19) 。
同样，已经证明，河流的淤泥也容易吸附病毒 (20)。在
水处理的混凝过程中，细菌和病毒被吸附在絮凝物上，并与 i虫度一起被除去 (21 ， 22) 。但是即使在出厂水中浊度仍低于
0.5
JTU 时，也会随着絮凝物透过谑床而引起病毒透过的
增加 (23) 。
无论有机物、无机物或大量微生物所形成的颗粒物均能 保护细菌和病毒免受消毒剂的作用。 Sanderson 和 Kelly 报道，对1虫度为 3.8~84 NTU 的水进行加氯消毒时，既使游
离余氯达 0.1~0.5 毫克/升，最低接触时间为 30 分钟后，仍 然含有大肠菌· 0 Neefe 及其同事的工作表明， 对人为用粪 便污染的饮水进行加氯消毒是不足以预防病毒性甲型肝炎 的 E243，因此只有在加氯前进行混凝和过媳才能提供安全的饮 水。实验室试验表明，水中的各种泥土和腐殖酸能降低紫外 绪对产气克雷自民菌的消毒效果 (25) 。
饮用高温度、加氯消毒的饮用水可能对健康带来危 害 (26 N 29) 。
某些悬浮颗粒具有吸附能力，能吸附水中有害的无机化
合物和有机化合物，因此 1虫度与水质对健康的影响有间接关 系。在这方面起最重要作用的是构成浊度的有机成分或腐殖 质成分 E30-333. ·对 Clark. N. A. 等文章的评边，参考文献 21.
-44 '7-
在1虫度成分中，某些金属-腐殖酸盐络合物的结合力可
使天然水中微量金属的分析测定更为复杂，从而可能造成对 金属含量的低估 (34) 。
有机分子也能被天然有机物所吸附。除草剂，例如 2 ， 4-D 、百草枯 (paraquat) 和敌草快 (Diqua 的，就能
吸附在泥土-腐殖酸颗粒上，而这种吸附受到腐殖质中金属 阳离子的明显影响 (35) 。因此，浊度也可能干扰对水样中杀 菌剂的检测。 迄今，浊度被用于衡量在水净化过程中去除颗粒物效果
的一种指标，出厂水中的浊度低表明混凝、沉淀和过掠过手里
是有效的。
16.4
对健康的影响
当 i虫度超过建议值 5NTU 时，用户往往会感到厌 恶 (36) 。如果在用户龙头水中的温度高于进入配水系统时的
j虫度，可能表明存在着水处理后的污染、腐蚀或其他的配水 问题。由于过高的浊度能影响对微生物的消毒作用、剌激水 中的细菌生长以及消超大量的氯，所以用氯消毒时应保持低
浊度，最好小于 1 NTU ，这对于提供安全的饮用水是至关 重要的。
参考文献
1. Black , A. P. & Hannah , S. A. Measurement of low tur bidities. J ournal 01 t he Ameri can Water W orks
Associatiotl. 57:901
(1 965).
2. Mcclulley , W. R. Radiometry of water turbidity measurements. J ournal 01 the Water Pollution 448 一
Control Federation. 47:252 (1 975). 3. Eden , G. E. The measurement of turbidity in water.
Procedings
01
the Society
01 Water
Treatment and
Examination , 14:27 (1 965). 4. AWW A Task Group. Progress toward a filtrability index test. Journal
01
the American Water Works
Association , 51 :1539 (1 959). 5. Standard methods of test for turbidity of waler , D1889-71 (1 977). Jn: 1980 Annual book for Testing and Materials , 1980 , p. 260. 6. American Public Health Association , Stand ard
01
AST M
standards , Part 3 1. Philadelphia , American Society
methods lor the examination 01 water and wastewater , 14th ed. Washington , DC , American Water Works Association , 1976 , p. 13 1. 7. Hach , C. C. Understanding tur bidity measurement.
lndustrial water engineering , 9:18 (1 972). 8. Packham , R. F. The preparation of turbidity standard.
Proccedings
01
the Society
01
Water Treatment
and Examination , 11:64 (1 962). 9. National interim Primary drinking ω ater regulations. Washington , DC , US Environmental Protection Agency. 1976 , p. 12 (EP A-570/9-76-003). 10. Turbidity. In , Methods lor chemical analysis
01
water and wastes. Washington , DC , US Environmental Protection Agency. 1976 , p. 295 (EP A-625-6-7 4-003a). 1 1. Beard , J. D. & Tanaka , T. S. A comparison of particle counting and nephelometry. Journal Ame 门ican
Water Works
01 the Association , 59:533
(1 97 7).
12. NationaI Research Counci I. Drinking water and -449-
hea/t!>. Wa&hington , DC , National Academy of Sciences , 1977 , Chapter IV.
13. Narkis , N. & Rebhun , M. The mechanism of flocculation processes in the presence of humic substances. J ournal
0/ the American (1 975).
Wate 俨 Works
Association , 67 :1 01
14. MacKenthun , K. M. & Keup , L. E. Biological problems encountered in water supplies. Journal Ame 俨 ican
0/ the (1 970).
Water Works Association , 62 1 520
15. Pemmanen , V. Humus fractions and their distribution in some lakes in Finland. Ina Povoledo , D. & Golterman , H. L 叫 ed. ， Humic substances , their
structure and /unction in the biosphere. Wageningen , The Netherlands , Pudo 口，
1975 , p. 207.
16. American Pub 1ic Health Association. Stand ard
methods /or the examination ω astewater ，
0/
water σnd
14th ed. Washington , DC , American Water
Works Association , 1976 , p. 64.
17. Atkins , P. F. & Tomlinson , H. D. Evaluation of daily carbon chloroform extracts with CAM. Water sewage
works , 110:281
(1 963).
18. Brock , T. D. Principles
0/
microbial ecology. Now
Jersey , Prentice-Hall Inc. , 1966. pp. 72-74.
19. Stotzky. G. Influence of clay minerals on microorganisms. 111. Effect of particle size , cation exchange capacity , and surf ace area on bacteria.
Canadian jourllal 0/ microbiology , 12'1235 (1 966). 20. Berg , G. Removal of viruses from sewage , effluents , and waters. 2. Present and future trends. Bulletin
0/
the World Health Organization , 49 :4 61 (1973). 一 450 一
21. Clarke , N. A. et a l. Human enteric viruses in water. source , survival and removabiHty. Jn. Proceed ings
01 the
Inte俨 national
Conlerence in
Wate俨
Po l/ ution Research , London , 1962. Aduances in water po l/ ution research , 2:523 (1 964).
22. Foliguet , J. M. & Doncoeur , F. Elimination des enterovirus au ('ours du traitement des eaux d' alimentation par coagulation-floculation-filtration.
Water research , 9: 953
(1 975).
23. Robeck , G. G. et al. Effectiveness of water treatment processes in virus remova l. Journal 01 the American
Water Works Association , 54:1275
(1 962).
24. Neefe , J. R. et al. Inactivation of the virus of infectious hepatitis in drinking water. American
journal 01 pub !i c health , 37: 365 (1 947). 25. Bitton , G. et a l. Effect of several clay minerals and humic acid on the survival of Klebsiella aerogenes exposed to ultraviolet irradiation. Applied
microbiology , 23:870
(1 972).
26. Dennis , J. M. 1955-56 Infectious hepatitis epidemic in Delhi , India. Journal 01 the American Water
Works Association , 51: 1288 (1 959). . ‘ 27. Symons , J. M. & Hoff , J. C. Rationale for turbidity maximum contaminant leve l. Presented at 3rd Water
Quality Technology Conlerence , Atl anta. Washington , DC , American Water Works Associ 现 tion ，
1975.
28. Hudson , H. E. High-quality water production and viral disease. Journal 01 the American Water Works
Asçociation , 54 =1 265 (1 962). 29. Taylor , F. B. et al. The case for water-borne 一 451 一
infectious hepatitis. American journal of public
health , 56:2093 (1966). 30. Schnitzer , M. pp. 204-25 1. 31. Chau , Y. K.
& Kahn , S.
U. Humic substances in the
environmen t. New York , Marcel Dekker Inc. , 1972 ,
& Lum-Shue-Chan , K. Measurement of
complexing capacity of lake waters. In. Povoledo , D. & Golterman , H. L. , ed. , Humic substances , their
structure and function in the biosphere. Wageningen , The Netherlands , Pudoc , 1975 , p. 11. 32. Oliver , B. G. Heavy metal levels of Ottawa and Rideau River sediments. Envi ronmental science
and technology , 7:135 (1 973). 33. Ramamoorthy , S. & Rust , R. R. Mercury sorption and desorption characteristics of some Ottawa River sediments. Canadian journal of earth sciences , 13: 530 (1 976). 34. Gardiner , J. The chemistry of cadmium in natural water- 1. A study of cadmium complex formation using the cadmium specific-ion electrode. Water
research , 8:23 (1 974). 35. Kahn , S. U. Adsorption of 2 , 4-D from aqueous solution hy Fulvic acid-clay complex. Environmental
science and technol ogy , 4: 236 (1 974). 36. Public health service drinking water standards , Rockville , MD , US Department of Health , Education and Welfare , 1962 , p. 21.
一 452 一
17. 钵 17.1 概述
样是一种含量丰富的元素，在地壳中的含量约为 0.04 克/公斤 C I)。最常见的铮矿是闪悻矿 (ZnS) ，它常与其他金 属元素〈如铅、铜、铺和铁〉的硫化物结合 E230 估计土壤巾 铮的天然含量为 1--300 毫克/公斤。飞
大气中铮的浓度变动很大，这取决于与发生源的距离等 因素。在农村，辞的一般浓度为 10--100 毫微克/立方米; 而在城市，最常见的浓度为 100--500 毫微克/立方米 C4)。
铸的碳酸盐、氧化物和硫化物微溶于水，而高度海解的 氯化物和硫酸盐容易水解生成氢氧化辞和破酸铮。因此天然 水中停的浓度通常很低。又由于这些铮被吸附在底泥上从而 更进一步地降低了可溶性铮的浓度 E530 可、
17.2
存
在
龙头水中铮的浓度可能明显高于地面水，这是由于铮从 镀铮管、黄铜管以及含铮设备中游离出来的缘故。在龙头水 巾铮的浓度通常波动在 0.01--1 毫克/升 (6) 。
17.3
接触途径
悻是一种重要的营养素，在美国已经有食物中铮含量的 综合表 E7， 82. 肉和乳制品含有丰富的铮，而谷类和坚果也是 一 453 一
铸的重要来源 (8) 。在某些重要的食物中铮含量如下 ω': 牛肉、猪肉和羊肉 乳 鱼和海产品 豆类和小麦
20~60 毫克/公斤 3~5 毫克/公斤 高于 15 毫克/公斤
15~50 毫克/公斤
叶菜和水果
低于 2 毫克/公斤(鲜重〉
迄今，食物是人类铸的最主要来源。据报道，"正常" 人的每日摄入量为 12 毫克 (10) ，从饮用水中的平均每日摄入
量可能不超过 400 微克，而从空气中的来源是微不足道的。
17.4
对健康的影响
铮是动物和人的一种必需元素，对于维持各种酶系统的
功能是必要的，这些酶系统包括碱性磷酸酶、碳酸配酶和乙 醇脱氢酶 (11)。已知含金属铸的酶达 70 多种(1 2) 。
根据年龄和性别，建议铮的膳食摄入量为 4........15 毫克/ 日。对于孕妇和产妇要求达到 16 毫克/日 (13) 。
已经报道了在埃及和伊朗伊斯兰共和国发生的一种地方 性铮缺乏综合征(在年轻人中) (14， 15) 。其特点是生长迟滞和
发育不全的其他体缸"包括贫血，可能由于肠吸收辞较少的 缘故。每日口服大剂量硫酸悻可得到满意的治疗(1 6) 。 对于人和动物，铮的吸收受很多因素(例如，摄入蛋白 质、维生素和金属〉的影响 (17)。摄入的悻少以及体重轻都能 增加对铸的吸收，而口服大剂量铮、钙和肌醇六磷酸盐却能
减少对铸的吸收 F 至于摄入的悻能被吸收多少是难以确定 的，因为铮也可能被排入肠道(1 8， 19) 。但是，铮不蓄积在组
织中，并且认:为吸收的比例与摄入量呈反相关 E2230
在人的血清和血浆中，辞浓度约为 1 毫克/升，而在全血 ← 454 一
中其浓度约增高 5 倍，这是由于在红细胞中浓度高 (10 毫克/ 升〉所造成的 (20) 。在前列腺中铮的含量最高(1 00 毫克/公 斤温重) ，在骨、肌肉、肝和膜中含量也高 (2 1)。
可以认为辞是无毒的。由于铮的毒性低以及在人体内存 在着有效的控制体内平衡的机制，因此从饮用水和膳食中来 源的铮引起慢性中毒是不可能的。人的辞中毒症状包括呕吐、 脱水、电解质失去平衡、腹痛、恶心、嗜眠、眩晕和肌肉缺 乏协调 (17) 。已有关于氯化铮引起急性肾衰竭的报道 (23)。
每日给予 150 毫克铮能干扰铜和铁的代谢，因为铮是这 两种金属的代谢拮抗物。但是，只要从膳食中摄入足量的铜 和铁，既使给予大剂量的铮也几乎没有问题。铮也是铺的代 谢拮抗物，因此可以预期，摄入大剂量铮能对机体有一些保 护作用，以对抗环境中榻的毒性作用 (20 。
昧阔测定表明，在人群中有 5% 的人能辨别出不含铮的 水和含铮浓度为 4.3 毫克/升〈硫酸铮)的7]< (25) 。用其他铮
盐测定的结果要偏高一些。
17.5
其他方面
铮使水产生不良的涩昧，此外，当水中浓度超过 5.0 毫 克/升时呈现乳白色，并在煮沸时出现一层滑腻的薄膜。因
此建议值为 5 毫克/升。但是为了避免产生上述任何问题， 应将水中辞浓度保持在该值以下。 (秦任慧译〉 参考文献 1. Browing , E. Toxicity
0/ industrial metals.
2nd ed.
London , Butterworths , 1969 , p. 348.
2. Quality criteria for water. Washington , DC , US 一 455 一
Environmental Protection Agency , 1976 , p. 48 1. 3. Levinson , A. A. /ntroduction to exploration
geochernistry. Calgary , Applied Publishing Co., 1974 , p.44. 4. Nriag 口， J. 0. , ed. Zinc in the environment. Part /1
ecological cycling. New York , John Wiley & Sons , 1980. 5. Hem , J , D. Zinc. in: Study and interpretation
0/
the chemical characteristics (Water-Supply Paper 1473).
0/
natural water.
Washington , DC , US Geological Survey , 1970 , p. 125 6. Zoeteman , B. C. J. & Brinkman F. J. J. Human intake of minerals from drinkingwater in the European communities. Jn Hardness
0/
drinking water and
pub !i c health. Proceedings of the European Scientific Colloquium , Luxembourg , 1975. Oxford , Pergamon Press , 1976 , p. 173. 7. Murphy , E. W. et al. Provisional tahles on the zinc content of foods. Journal
0/
the American Dietetic
Association , 66:345 (1 975). 8. Freeland , J. H. & Cousins , R. J. Zinc content of selected f oods. J ournal
0/
the American Dietetic
Association , 68:526 (1 976). 9. WHO Technical Report Series , No. 532 , 1973 (Trace
elements i" hurnan nutrition: report of a WHO Expert Committee) . 10. Warren , H. V. Some trace element concentrations in various environments. Jn: Howe , G. M. & Loraine , J. A. , ed. Environmental medicine. London , William Heinemann Medical Books Ltd.. 1973 , p. 9. -456 一
11. Parisic , A. F. & Vallee , B. L. Zinc metalloenzymes: characteristics and significance in biology and medicine.
American journal 01 cli nical nutrition , 22: 1222 (1 969) .
12. Symposium on trace elements. M ed ical clinics 01
North America , 60.4
(1 976).
13. Food and Nutrition Board. Recommended dietary
allowances. 8th revised version. Washington , DC , Nationa! Acadeiny of Sciences , 1974 , pp. 99-10 1. 14. Prasad , A. S. et a! , Syndrome of iron , anemia , hepatosplenomegaly , hypogonadism , dwarfism and geophagia. American journal 01 medicine , 31 '532 (1 96 1) .
15. Halsted , J. A. et a I. Zinc deficiency in man. The Shiraz experiment. American journal 01 medicine , 53' 277 (1 972). 16. Michaelson , G. Zinc therapy in acrodermatitis enteropathica. Acta dermatologica , 54:377 (1 974 人
17. Prasad , A. S. & Oberleas , D. , ed. Trace elements in
human health and disease. Vo l. 1: zinc and copper. New York , Academic Press , 1976 , p. 470. 18. Becker , W. M. & Hoekstra , W. G. The intestinal absorption of zinc. In: Skorylla , S. C. & ValdronEdwards , D. , ed. Intestinal absorption Pergamon Press , p. 229. 19. Honstead , J. F. & Brady , D. N. The uptake and retention of 32P and 65Zn from the consumption of Columbia River fish. Health physics , 13:455 (1 967). 20. Sunderman , F. W. In , Goyer R. A. & Mehlman , M. - - 457-
0/ metal
ions. trace elements and radionuclides. New York ,
A. , ed. Advances in modern toxicology. Washington , DC , Hemisphere Publishing Corporation , 1976. 2 1. Halsted. (1 974) .
J. A. et a I. Journal 01 nutrition , 104:345
22. Heth , D. A. et a I. Effect of calcium , phosphorus and zinc on zinc-65 absorption and turnover in rats fed semipurified diets. J ournal 01 nut rition , 88: 331 (1 966). 23. Csata , S. et a l. Akute Niereninsuffizienz als folge einer. Zinkchloridvergiftung. Zeitschri/t /ür
U rologi e , 61: 327
(1 968).
24. Underwood , E. J. Trace element& in human and
animal nutrition , 4th ed. New York , Academic Press , 1977 , p. 545. 25. Cohen , J. M.'et a l.、 Taste
threshold concentrations of
metals in drinking water. J ournal
0/ the American (1 960).
Water Works Association , 52:660
-
458--
第六章放射性物质
去一口
1. 前
在 1970 和 1971 年世界卫生组织 (1 ，'2)发表的标准中建议
了饮水中放射性的水平，它是基于国际放射防护委员会
<I CRP)
1959"'1966 年的资料。但是，此后又有了新的资
料 (3-5) ，在准备本"准则"时已将其考虑在内。
按成人饮水量每日 2 升计，建议饮水中总 α 放射性的参 考水平为 O.lBq/升，总声放射性为 1Bq/升。并建议当放射
性超过该值时，应向主管当局报告，以便采取相应措施。有 关总 α 和总自放射性以及个别放射性核素的测定程序将在其 他文章中论述 (6) 。
放射性物质可从很多来源〈天然的和人为的)进入环境。 天然的来源，包括由宇宙射线产生的放射性物质(可随雨水和
径流进入水中〉以及在岩石和土壤中存在的放射性物质，例 如 238 铀及其衰变产物川孀和 222 氨 F 人为来源的放射性核
素，来自核试验的落下灰、核电站以及在医学和其他方面应
用的放射性物质。 个人接受的天然辐射的剂量取决于若干因素，诸如居住 地的海拔高度、土壤中放射性核素的数量和种类以及从空气、
食物和水中摄入到体内的放射性核素的量。随着核应用的增 长，放射性物质更广泛地释放到环境中，从而使地面水和地 下水的放射性物质增加，直接影响到公共给水水源中放射性 的水平。
一 460 一
2. 基本考虑 在评价'电离辐射照射时，国际原子能机构(l AEA)~7) 和国际放射防护委员会(l CRP)C4 ， 5) 做如下建议。 ICRP
现已引用"危害 (detriment 沪的概念，以便对全部有害
作用进行鉴别和定量〈如果可能的话)
0
"危害"可以定义为
辐射照射所致损害的数学"期望"值，不仅要考虑每一种有 害效应的发生机率，而且也要考虑效应的严重程度。 ICRP
认为，在未加说明的情况]了，有关吸收剂量本身尚不足以预 测照射所致对健康方面的有害效应的严重程度和发生机率。
ICRP 建议用于放射防护的剂量当量限值不适用于或不包 括天然辐射照射的水平，仅涉及由于人为活动增加了的某些 成分。但是 ICRP 清楚地认识到 2 在"正常"的天然辐射
水平与由于人类活动或环境的选捧而提高了的辐射水平之间 没有明显的界线。 ICRP 已经建立了剂量限制体系，具有下列要求 z
"( a )除非这种实践带来绝对的利益二否则不应采纳， (b) 所有照射均应保持在可以合理做到的最低水平，
应考虑到经济和社会因素，
(c
)个人所受的剂量当量，不得超过委员会对相应的
情况所建议的限值"。
2.1
剂量-反应关系
辐射照射对健康的有害效应，→是驱体效应，即辐射效
应显现在受照射者本人身上 F 一是遗传效应，即影响到受照 射者的后代。
-46).--
甚至在放射防护剂量水平时，某些躯体效应(例如致癌 作用)以及遗传效应发生的机率〈而不是严重程度〉与剂量
呈函数关系，这种效应无阔值，称为随机效应。其他躯体效 应的严重程度却随剂量而变化，并且只有高于阔剂量时才出 现反应，称为非随机效应。
放射防护的目的在于防止有害的非随机效应，并限制随 机效应的发生率使之达到被认为可以接受的水平。后→个目 的可通过在第 2.2 节中论述的剂量限制体系来实现。而防止
非随机效应则通过制订剂量当量限值来实现，将限值限定在 足够低的水平，即使终生接受照射也不会达到阔剂量。对饮
水中放射性物质的水平，既要求其随机效应的发生率降至可 接受的水平，显然也将排除发生非随机效应的可能性。 剂量当量是对剂量进行定量的一种方法，能使辐射照射 与引起的有害效应〈特别是晚发的随机效应〉之间呈现更好
的相关。 ICRP 的基本假设是在所关心的辐射照射范围内，随机
效应的发生机率，与所接受的剂量成比例。但是体内不同组 织对放射的敏感性不同，因此 ICRP 引入剂量当量加权系 数 (dose-equivalent
- weighting
factors) 以便在相
等危险度的基础上进行危害的量度。每种组织的加权剂量当 量系数的总和代表了受照射的总危险度，称为有效剂量当 量 (effective dose-equivalent) 。
长寿命放射性核素在体内存留的时闸相当长，能使机体
接受照射的时间持续很多年。因此将有效剂量当量 (H E ， 50) 定义为摄入放射性核素后 50 年内有效剂量当量的总和。
在这一讨论中，可以简单地使用"剂量"来衡量照射的 水平。 -162 一
2.2
剂量当量限值
委员会建议的剂量限值用于两类照射一一职业性照射和 一般照射。剂量当量限值是指外照射所产生的年剂量当量和 在一年实践中摄入体内的放射性产生的待积剂量当量二者的 总和。对职业性照射确定的剂量当量限值应当视为上限值。
而对公众中个人的剂量当量限值是更偏于理论的一个模念， 其用意是要保证个体不可能接受超过规定值的剂量当量。往 往通过抽样和统计学计算以及控制产生照射的辐射源来检验 这种限制的效果。 根据 ICRP 的建议，对放射防护的基本安全标准确定 为，公众个体成员的有效剂量当量限值是 5mSv/ 年，该值 用于"关键人群组"
(cri tical
grou P ).
作为平均剂量当
量。但是当同一个体在持续多年内接受接近于该限值的剂量 照射时，应谨慎地采取措施将终生剂量当量限制在相当于年
平均有效剂量当量为 lmSv 。
2.3
对饮水水质的要求
为将上述原则应用于饮水，需要测知来源于饮水的可能 照射。虽然根据经核准的从核装置中排放入环境的人工放射 性物质的资料以及天然放射性的地质资料等可以做出一些估
计，但是仍需直接测定放射性的水平。如果饮水中的放射性 已低到可以接受的程度，则需根据特定的环境以及权威当局
的建议来决定进一步测定的频度。 • ICRP 对"关键人群组"的定义为: i主组人群具有一些特性，而使得他们 从某-实践中受到的照射水平高于受照人群中的其他成员。这些人群组可作为 某项实践所产生的个体剂量上限恒的一种量庭。
-463-
对于新的给水，需要具有原水中放射活性的资料，并需 检查处理后的出厂水中放射性水平。
权威当局最好把饮水中的放射性水平与饮用该水的人群 受到其他各种辐射源照射的总量联系起来考虑。
3. 辐射照射的来源 根据 ICRpC4 ， 5) 建议的剂量限制体系， 建立了估算个
体受照射水平的基本准则。照射可来源于天然放射性核素产 生的"正常水平"或因人类活动使正常照射有所增高，例如 磷肥的使用或矿山废水的排放，以及来源于逸入环境的人工 放射性核素，诸如核试验的落下灰、从核能设施逸出的排出 物和在医学、工业和研究中使用放射性核素排放的废物 (8)。
事实上，联合国原子辐射效应科学委员会 (UNSCEAR)
尽可能对天然和人工放射性的照射水平进行了全球性的定期 评价，最近的报告发表于 1977 年 E830 报告的资料表明，饮
水中接触的剂量仅占总照射量的很小部分。
4. 对总 α 和总卢放射 性的建议值 自然环境中存在的以及人类活动产生的放射性核素主要
是发射 α 和声射线的放射源，其中一些还有衰变产物。天然 存在的发射 α 射线的放射性核素以 M 锚为代表，发射声射 线的人工放射性核素以.。媳为代表。
但是，当放射性浓度很低时，往往不一起需要鉴别特定 -464 一
核素。在这种情况下，测定总 α 和总 R 放射活性就能确定可 以接受的放射毒性水平。因此测定总 α 和总 R 放射性对常规
监测是特别重要的。 以总放射活性作为筛选程序，要求总 α 和总自放射活性 的参考水平满足两项准则
( a )无论各核素在总活性中的贡献如何，其给出的总照 射量应足以低到不需要进一步详细分析和考虑，
( b) 所制定的参考水平要足够高，以使绝大多数给水水 源符合该值，从而避免做进一步的详细分析。 1979 年建议总 α 和总自放射活性的参考水平分别为 0.1 和 0.8B q (3)。该值能满足第二条准则，但是随着科学知识的
发展，还需继续满足第一条准则。为此必须分别考虑有关的 核素。表 5 列出了有关的核素，并且注明了当每天饮用 2 升
水，
α 和 R 放射源的污染水平分别为 O.IBqj 升和 lBqj 升 目前还没有公认的方法能用来精确计算一个人从含氧饮
时，对各种核素的可能照射水平。 水中接受的真实剂量，虽然 UNSCEAR 和 ICRP 正在
审议这一主题。但是，已知饮水的途径具有较小的意义。 在测定总 α 和总 P 放射活性时，不包括氧和佩(一种低
放射毒性的核素〉是可以理解的。如果怀疑两者中的任何一 种核素出现区域性的局部增高，都应请权威当局进行咨询。
已经接受的一种看法是，某些超铀同位素比表中所列核 素毒性更大，但是在饮水中达到可察觉浓度的情况是罕见
的。当怀疑存在这种同位素时，应请权威当局进行指导以采 取特殊的预防措施。 在表中列入的发射 α 射线的核素中， ZSZ 址不可能是总 α
活性的主要构成者，因此保守的作法是将 α 放射活性完全归 -465 一
表 5
每天饮用 2 升水，在 1 年内对各种 α 和月放射源的可能照射量 HE , 50jBq(5v). HE , 50 →年内的摇 λ
核素
(m5v) α 放射i原
73Bq( =0.1Bqj 升) 4.36 X 10- 7 7.63 x 10- 1 3.05X 1O- 7 7.4X 1O- 7 7.07 x 10- 1 6.32 x 10- 1 0.032 0.006 0.022 0.054 0.005 。 .005
210pO 224Ra 21ð Ra 232Th 2UU 13 SU
'放射源 aoc 。
730Bq(= lBqj 升) 6.97 x 10- 8 2.17 X 10- 8 3.6x10- S 7.4xl0- 1 1. 4x lO- S 1. 98 x 10- 1 1. 36x lO- S 1. 36 x 10- s
0.005 0.002 0.026 0.054 0.010 0.014 0.010 0.993 0.214
a"Sr 80 5 r 120 1 13 1J I !4 CS
137 CS 21 0Ph 22SRa
3.3xl0- 7
资料来源于 ICRP 第 30号出版物的增刊 [5 )。
于 ZZð 错。
就自放射源而言，可以采用与超铀同位素相同的理由而 将 12U腆排除。 228 孀和更特殊的 210 铅是表中最毒的核素，但是，
正常情况下这些核素仅占总 9 放射活性的很小部分，而在7](. 巾氧浓度很高时例外。因此，在氧浓度不高时，一种保守的 假设是将自放射活性完全以 eù 媳来考虑。 -:66 一
由此可见，将每升 0.1Bq 的 α 放射性以 M 锚计，以及
1Bq 的 R 放射性以，。银计， 的接触剂量为 0.048mSv 。
当每天饮水量为 2 升时，每年
在上述浓度下的实际接触量将不超过 0.048mSvj 年， 事实上甚至几乎不可能达到这一剂量。 根据 ICRP 的资料，该剂量相当于总危险度为每年 10 → .-...10 町'，这一数值比任何一个公众成员"可能接受的水
平"低一个数量级，它完全符合饮水仅占总照射量一部分的 事实。 因此，建议参考水平为 z 总 α 放射活性 1Bqj升和总自 放射活性 0.1Bqj 升。
5. 氨 几个国家的资料表明，在某些居民区给水的地下水源 (深井〉中氧的活性达 10 3 Bqj 升。由于在处理过程中氧易
于从水中丢失，故难以评价摄入量。虽然许多人试图精确计 算饮用含氧饮水者所接受的真实剂量，但是都不可能做到这 一点 (3)。
值得注意的是，就氧而言，在使用自来水的地区，人们
从室内空气中吸入氧所带来的对健康危害要大于饮水的危 害 E83.
〈秦钮慧译，张明校〉
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Geneva , World Health Organization , 1970.
2. International standards for drinking-water , 3rd 467 一
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Organizatio刀，
197 1.
3. Radiological examination 0/ drinking- u, ater. Report 0/ a W HO Working Group. Copenhagen , WHO Regional Office for Europe , 1979 (EURO Reports and Studies No. 17). 4. Jnternational Commission on Radiological Protection. Recommendations of the International Commission on Radiological Protection. Annals
0/ the ICRP , 1(8)'
1-53 (1 977) (ICRP Publication 26).
5. International Commission on Radiological Protection. Limits for intakes of radionuclides by workers.
Annals
0/ the ICRP , 2-8 ,
(1 979-1982) (ICRP
Publication 30 and supplements).
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examinεtion.
In. Suess ,
J. , ed. Examination 0/ water /or pollution
control , vol. 2 , Oxford , Pergamon Press , 1982 , cha pter 5. 7. International Atomic Energy. AGENCY. Basic sa/ety
standards /or radiation protection. Vienna , IAEA , 1982 (Safey Series No. 9) . 8. United Nations Scientific Committee on the Effects 。f
Atomic Radiation. Sources and e//ects
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radiation. New York , United Nations , 1977.
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