National Academies Press: OpenBook

Drinking Water and Health,: Volume 3 (1980)

Chapter: IV Toxicity of Selected Drinking Water Contaminants

« Previous: III Problems of Risk Estimation
Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Suggested Citation:"IV Toxicity of Selected Drinking Water Contaminants." National Research Council. 1980. Drinking Water and Health,: Volume 3. Washington, DC: The National Academies Press. doi: 10.17226/324.
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Below is the uncorrected machine-read text of this chapter, intended to provide our own search engines and external engines with highly rich, chapter-representative searchable text of each book. Because it is UNCORRECTED material, please consider the following text as a useful but insufficient proxy for the authoritative book pages.

IV Toxicity of Selected Drinking Water Contaminants CHEMICALS EVALUATED the health effects of a large number of contaminants found in drinking water were evaluated in Drinking Water and Health (National Academy of Sciences, 19771. The compounds evaluated in this chapter were selected for the following reasons: 1. Sufficient new data have become available to justify further attention to several chemicals examined in the first study. 2. New contaminants have been identified in drinking water subse- quent to the first study. 3. Several compounds were judged to be of concern because of potential spill situations. 4. The chlorination of drinking water or the use of other disinfectants yields compounds that require toxicological evaluation. A list of such compounds was prepared by the Safe Drinking Water Subcommittee on Chemistry of Disinfectants and Products. They are evaluated in this chapter by the Subcommittee on Toxicology. The 1977 study (National Academy of Sciences, 1977) examined the radioactive, particulate, and chemical contaminants found in drinking 67

68 DRINKING WATER AND H"LTH water. Radioactive contaminants are not considered in this study. Asbestos was one of the particulates examined in the first study. A reevaluation of this contaminant will be justified when the several studies now underway are completed. The number of volatile organic com- pounds identified in drinking water supplies has increased from approximately 300 at the time of the first study to 700 at the present time and will continue to grow. Limitations of time, manpower, and scientific information have not permitted an in-depth evaluation of most of the compounds recently found in drinking water. It was the belief of this subcommittee that it could perform a more valuable service to the Environmental Protection Agency (EPA) in the future if it evaluated criteria documents that were prepared by the EPA or other groups contracted to conduct these tasks. It will be necessary for the EPA to develop a mechanism for a comprehensive search and review of the literature in order to make in-depth hazard assessments for these chemicals. It is the consensus of this subcommittee that this cannot be done appropriately by the National Academy of Sciences because time and staff requirements far exceed those available. Neither can it be expected that the scientists who donate their services on these subcom- mittees will have the resources or time to carry out the routine aspects of this task. ACUTE EXPOSURES In addition to providing information on chronic toxicity, the subcommit- tee has evaluated the potential acute toxicity insofar as justified by the available data. These data will provide a basis for making judgments of possible health erects resulting from accidental spills of chemicals into drinking water supplies. To this end the subcommittee has provided a suggested no-adverse-response level (SNARL) for acute exposures of 24 hr or 7 days. These values are calculated based on the assumption that 100% of the exposure to the chemical was supplied by drinking water during either the 24-hr or 7-day period. In those few cases where the chemical is a known or suspected carcinogen' the potential for carcinogenicity after an acute exposure has not been considered These acute SNARL's were calculated only when there was human exposure data or sublethal animal data. LD50's were not used as a basis for calculation. Some 7-day values were derived by dividing the 2=hr SNARL by 7, but only when the data were very good. The converse was

Toxicity of Selected Drinking Water Contaminants 69 not done nor were data obtained from studies of lifetime exposures used to establish acute SNARL's. In some cases in which data from inhalation exposures were used there was information on absorption/retention. The details concerning retention and absorption are given in the monographs for each chemical. It must be emphasized that these calculated acute SNARL's should not be used to estimate hazard from exposures exceeding 7 days. They are not a guarantee of absolute safety. Furthermore, SNARL's are based on exposure to a single agent and do not take into account possible interactions with other contaminants. In all cases the safety or uncertainty factor used in the calculations of the SNARL's reflect the degree of confidence regarding the data as well as the combined judgment of the subcommittee members. As in the previous report, the following assumptions were used when assigning an uncertainty factor to calculate either the acute or chronic SNARL's: · An uncertainty (safety) factor of 10 was used when good chronic or acute human exposure data were available and supported by chronic or acute data in other species. · A factor of 100 was used when good chronic or acute toxicity data were available for one or more species. · A factor of 1,000 was used when the acute or chronic toxicity data were limited or incomplete. CHRONIC EXPOSURES When the chemical of concern was not a known or suspected carcinogen, the subcommittee calculated a SNARL for chronic exposure. In most cases, whenever chronic SNARLS's were estimated, data were available from studies lasting a major portion of the lifetime of the experimental animal. For these SNARL's an arbitrary assumption was made that 20~o of the intake of the chemical of concern was from drinking water. Because of this assumption it would be inappropriate to use these values as though they were maximum contaminant intakes. A risk estimate rather than a chronic exposure SNARL was provided in those cases in which there was adequate evidence of carcinogenicity [see Drinking Water and Health (National Academy of Sciences, 1977) for details]. Table IV-1 summarizes the acute and chronic SNARL's as well as the carcinogenic risk estimates for the chemicals reviewed in this report.

70 DRINKING WATER AND H"LTH TABLE IV- ~ Summation of Acute and Chronic Exposure Levels and Carcinogenic Risk Estimates for Chemicals Reviewed Suggested No-Adverse-Response Upper 95% Level (SNARL), mg/liter, Confidence by Exposure Perioda Estimate of Lifetime Cancer Chemical24-hour 7-day Chronic Risk per,ug/literb Acrylonitrile1.3 x 10-6 Benzene 1 2.6 Benzene hexachloride3.5 0.5 Cadmium 0~08 0.005 Carbon tetrachloride14 2.0 Dichloro difluoromethane350 5.6 1,2-Dichloroethane7.0 x 10-7 Epichlorohydrin0.84 0.53 Ethylene dibromide9.1 x 10-6 Methylene chloride35 5.0 Polychlorinated biphenyl0.35 0.05 Tetrachloroethylene172 24.5 1.4 x 10-7 1, 1,1-Trichloroethane490 70 3.8 Trichloroethylene TrichloroQuoromethane Toluene Uranium Xylenes Bromide Catechol Chlorine dioxide Chlorite Chloroform Oi bromochloromet inane 2,4-Dichlorophenol Hexachlorobenzene Iodide Resorcinol 105 15 88 8 420 35 0.34 3.5 0.21 21 11.2 1,400 224 2.3 2.2 22 18 115.5 11.7 0.38 0.21 3.2 0.03 16.5 0.5 0.7 1.19 2.9 x 10-5 a See text for details on individual compounds. b See Drinking Water and Health (National Academy of Sciences, 1977) for details.

Toxicity of Selected Drinking Water Contaminants 71 CHEMICALS SELECTED BY EPA Acrylonitrile (CH2=CHCN) Acrylonitrile is an unsaturated synthetic organic compound that has a variety of applications. Primarily it is used in the production of acrylic and modacrylic fibers, nitrite rubber, and plastics. As a pot nt, highly effective fumigant it is used most often to protect grain, dried fruit, walnuts, and tobacco against insect pests. Annual production totals approximately 682 million kg (U.S. Environmental Protection Agency, 1 978a). Acrylonitrile, also known as 2-propenenitrile, vinyl cyanide, and cyanoethylene, is a colorless, highly flammable liquid with a mild, pungent odor resembling that of peach pits. It is manufactured by the reaction of propylene with ammonia in air. Its boiling point is 77.3°C. At 20°C its solubility in water is 7.35 g/100 ml, and the specific gravity of the liquid is 0.811 (Manufacturing Chemists Association, 19741. The principal exposure of humans to acrylonitrile is likely to occur through atmospheric contamination. Patterson et al. (1976) estimated that the total emission from manufacturing processes in 1974 was 14 million kg. There is relatively little information on the movement, fate, and persistence of acrylonitrile in water. Zabezhinskaya et al. (1962) reported that at an initial concentration of 10 mg/liter, only 46% remained after 24 hr. 19% after 48 hr. and 537O after 96 hr. Under some conditions, the relatively high vapor pressure of acrylonitrile would probably promote the escape of the compound to the atmosphere. METABOLISM The metabolism of acrylonitrile has not been studied systematically with radio-labeled material. Investigators studying the metabolism of acrylo- nitrile have concentrated on ascertaining the fate of the cyano group of the molecule. Brieger et al. (1952) reported that inhaled acrylonitrile is metabolized first to the free cyano group, and then to thiocyanate. They found high cyanate levels and high thiocyanate levels in the blood of rats, dogs, and monkeys that had been treated with acrylonitrile. In a study with both oral and intraperitoneal administration of acrylonitrile in rats, mice, and Chinese hamsters, Gut et al. (1975) observed that thiocyanate was eliminated in the urine and that acrylonitrile was bound strongly to components of the blood. There appeared to be an interspecific difference in the metabolic pattern: less cyanide formed in rats than in mice. Earlier studies (Dudley and Neal, 1942; Lawton et al.,

72 DRINKING WATER AND H"LTH 1943; Paulet et al., 1966) indicated the same lack of agreement on the principal metabolic product of acrylonitrile. Some of the studies reported cyanide, and others, thiocyanate, as the principal breakdown product. The fate of the remainder of the molecule has not been established. The possibility of conjugation has been raised by Hashimoto and Kanai (1972) who reported binding of acrylonitrile with cysteine and gluta- thione with a concomitant decrease in sulfhydryl (SH) groups. Earlier work by these investigators indicated that a large amount of injected acrylonitrile was unchanged after treatment of rabbits, guinea pigs, and rats. HEALTH ASPECTS Observations in Humans A recent report by the DuPont Corporation to government regulatory agencies stated that acrylonitrile may be a carcinogen (Anonymous, 1977~. It also stated that preliminary results of an epidemiological study of workers in a polymerization operation with potential for exposure to acrylonitrile indicated excess cancer incidence and cancer mortality as compared with company and national experi- ence. This study included about 470 males who began working in the polymerization area of the plant between 1950 and 1955 and who are still actively employed or have been retired from the company (Anonymous, 19771. The DuPont report emphasizes that the data are preliminary and that more exhaustive studies are under way. A number of incidents of illnesses and fatalities have been caused by the industrial and structural pest control uses of acrylonitrile. The compound is an acute poison and a severe skin and eye irritant. It may be toxic when inhaled, ingested, or absorbed through intact skin. Symptoms of exposure include nasal and respiratory oppression, vomiting, nausea, weakness, fatigue, headache, and diarrhea (Patterson en al., 19761. The symptoms of poisoning from acrylonitrile are very similar to those from cyanide. Such poisoning generally results from inhalation by workers of vapor in industrial settings where the concen- tration of acrylonitrile varies from 16 to 100 ppm. No fatalities have been reported under these conditions. In contrast to the safety record of acrylonitrile in the chemical industry, the use of the compound in structural pest control has resulted in a number of fatalities (Davis, 1967; Davis e! al., 1973; Patterson et al., 1976; Radimer et al., 1974; Sartorelli, 1966). The reported fatalities generally resulted from direct exposure to acrylonitrile or from too rapid a return to a building that had been fumigated with the compound.

Toxicity of Selected Drinking Water Contaminants 73 Death resulted following symptoms that are similar to those of cyanide poisoning. However, death has also occurred as the result of toxic epidermal necrolysis (Radimer et al., 1974~. Acrylonitnle, when used for fumigation, is generally combined with other materials such as meth- ylene chloride and carbon tetrachlonde. Consequently, it is somewhat difficult to disassociate the symptoms of one compound from another. Observations in Other Species Acute Elects Investigating acute poisoning after intravenous injec- tion of acrylonitrile in dogs and rabbits, Paulet et al. (1961) absented considerable differences in interspecific response to cyanide intoxication. Symptoms of nervous disorders dominated the picture. Electroencepha- lographic records show that the higher nervous centers were affected. The investigators also found hyperglycemia and a decrease in the concentration of plasma inorganic phosphate. Acute oral toxicity values (LDso's) for acrylonitrile range from 27 to 128 mg/kg for mice (Benes and Cerna, 1959; Zell!e,r et al., 1969) and from 78 to 93 mg/kg for rats (Benes and Cerna, 1959; Smyth and Carpenter, 1948~. Acrylonitrile has also been characterized as a serious hazard in inhalation studies conducted by the Union Carbide Corporation in rats (Union Carbide Corporation, 1970~. After breathing saturated air for 5 min. all exposed animals died. After breathing 100 ppm for 4 hr. all of six rats died; for 2 hr. one of six; and for 1 hr. none of six. After breathing 500 ppm for 4 hr. none of six died, and for 8 hr. one of six died (Union Carbide Corporation, 19701. Roudabush et al. (1965) reported that acute dermal LDso values from acrylonitrile were 0.28 mg/kg when applied to the abraded skin of rabbits, 0.46 mg/kg on the intact skin of guinea pigs, and 0.84 mg/kg on the abraded skin of guinea pigs. Subchronic and Chronic Effects There are surprisingly few long-term studies on the toxicity of acrylonitrile in laboratory animals. None of the few studies in the literature was designed to establish a no-adverse-e~ect or minimal-effect dosage level. Dudley et al. (1942) conducted a three-part study of the inhalation toxicity of acrylonitrile. In a preliminary series, they exposed four rhesus monkeys and two dogs for 4 hr/day, 5 days/week for 4 weeks to an average concentration of 0.12 mg/liter (56 ppm) of acrylonitrile in air. These experiments indicated that dogs are more susceptible to acryloni- trile than are monkeys and that repeated exposure to concentrations of 0.12 mg/liter produces no signs of cumulative action. In the second part of their study, they exposed 16 rats, 16 guinea pigs, 3 rabbits, and 4 cats

74 DRINKING WATER AND H"LTH in the same manner for 8 weeks to an average concentration of 0.22 mg/liter (100 ppm) of acrylonitrile in air. These experiments show that rats, guinea pigs, and rabbits tolerate repeated exposures to 0.22 mg/liter acrylonitrile in air over a period of 8 weeks; that cats are definitely more sensitive to acrylonitrile than are rodents; and that there is no evidence of cumulative action of acrylonitrile. In the third part of the study, they exposed 16 rats (8 adult, 8 young animals), 16 guinea pigs, 4 rabbits, 4 cats, and 2 rhesus monkeys in the same way to an average concentration of 0.33 mg/liter (153 ppm) acrylonitrile in air. These experiments showed that repeated exposures to 153 ppm were definitely toxic to guinea pigs, rats, and rabbits and were much more toxic to monkeys and cats. The exposures produced irritation of eyes and nose, loss of appetite, gastrointestinal disturbances, and incapacitating weakness of hind legs from which the animals recovered relatively rapidly. Even after exposure to such high concentrations no definite evidence of cumulative action was observed. In a study on the ejects of acrylonitrile on rats Barnes (1970) noted no adverse ejects. Six rats were given 15 successive oral doses of 30 mg/kg, followed by seven doses of 50 mg/kg, and then 13 doses of 75 mg/kg over a period of 7 weeks. The investigators supplied no details on the types of observations that were made to assess toxicity. Studies have also been conducted on the toxicity of acrylonitrile to adult rats following daily intraperitoneal administration of the com- pound (Knobloch et al., 19711. Daily injection of 50 mg/kg for 3 weeks produced a statistically significant loss of body weight; leucocytosis; signs of damage and functional disturbances of liver and kidneys; increase in the weight of liver, kidneys, and heart; and histological damage. Microscopic examination of the organs of these animals showed slight damage of neuronal cells of the cortex and brain stem and parenchymal degeneration of liver and kidneys. Unfortunately, no other dosage rates were included in this study. Following the reports of epidemiological evidence of the carcinogeni- city of acrylonitrile, Norris ( 1977) initiated a 2-year feeding and inhalation study in rats. In the ingestion study, acrylonitrile was incorporated into the drinking water of laboratory rats at concentrations of 0, 35, 100, and 300 mg/liter (corresponding to 0, 4, 10, and 30 mg/kg/day). In the inhalation study, male and female rats were exposed to 0, 20, and 80 ppm acrylonitrile for 6 hr/day' 5 days/week. In April 1977, interim results of the 2-year studies were reported (National Institute for Occupational Safety and Health, 1977; Norris, 1977~. Rats ingesting 35 mg/liter acrylonitrile exhibited mild signs of toxicity while

Toxicity of Selected Drinking Water Contaminants 75 those ingesting lOO and 300 mg/liter showed marked signs of toxicity. Norris (1977) reported that both male and female rats that ingested lOO or 300 mg/liter acrylonitrile for 12 months developed stomach papillo- mas (1 of 20 rats at lOO mg/liter and 12 of 20 at 300 mg/liter); central nervous system tumors (2 of 20 at 35 mg/liter, 2 of 20 at 100 mg/liter, and 3 of 20 at 300 mg/liter); and Zymbal gland carcinoma (2 of 20 at lOO mg/liter, and 2 of 20 at 300 mg/liter). No such tumors were seen in control animals. In the inhalation study, after 1 year of exposure to 80 ppm acrylonitrile. 26 rats died and three developed central nervous system tumors that were comparable to those reported in the ingestion study. Gross examination of other rats in this study, who were also exposed to 80 ppm acrylonitrile by inhalation, revealed an increased incidence of ear canal tumors and mammary region masses. In animals exposed to 20 ppm, there was an apparent increase in subcutaneous masses of the mammary region, although no ear canal or central nervous _ 1 ~.1 ~1 system tumors were observed. (other than neoplasms, signs of toXlClty were limited to decreased water and food consumption and decreased body weight gain. Mutagenicity The mutagenicity of acrylonitrile has been demon- strated in the Salmonella typhimurium test (Milvy and Wolff, 1977) and in E. cold WP2 strains (Venitt and Bushell, 1977~. In the Ames (Salmonella) assay, acrylonitrile was active in the presence of a mouse liver homogenate, producing mutations in three tester strains. Bacteria were exposed by spotting the acrylonitrile on a lawn of Salmonella; by shaking a reaction mixture consisting of bacteria, liver homogenate, and acrylonitrile; and by exposing the homogenate and bacteria to an atmosphere containing acrylonitrile. By the latter method mutagenicity was observed at exposures as low as 57 mg/liter. Acrylonitrile was also mutagenic in various DNA-repair strains of E. cold WP-2. The effects were weak in plate incorporation tests, but assays using a simplified fluctuation test showed acrylonitrile to be significantly mutagenic at doses that were 20 to 40 times lower than those giving significant results in the plate test. Use of the different DNA-repair strains indicated that acrylonitrile causes DNA damage of the type that is exemplified by methyl methanesulfonate. Carcinogenicity Acrylonitrile has given positive results in a rat feeding study. Epidemiological evidence also contributes strong evidence to implicate acrylonitrile as a carcinogen. These studies are discussed above.

76 DRINKING WATER AND H"LTH Reproduction The subcommittee found no studies on reproductive effects of acrylonitrile. Teralogenicity The subcommittee noted no studies reporting the teratogenicity of acrylonitrile. Carcinogenic Risk Estimate The interim results of a 2-year ingestion study with acrylonitrile in the drinking water of rats give evidence of what appears to be an increase in cancer at several sites (Norris, 1977~. Dose-response data (Norris, 1977) were used to estimate both the lifetime risk and an upper 957 confidence bound on the lifetime risk at the low dose level. These are estimates of lifetime human risks which have been corrected for species conversion on a dose/surface area basis. The risk estimates are expressed as a probability of cancer after a lifetime consumption of 1 liter/day of water containing 1 ~g/liter of the compound under study. For example, a risk of 1 x 10-6 implies a lifetime probability of 2 x 10-5 of cancer if 2 liters/day were consumed and the concentration of the carcinogen was 10 ,ug/liter. This means that at a concentration of 10 ~g/liter during a lifetime exposure this compound would be expected to produce one excess case of cancer for every 50,000 persons exposed. If the population of the United States is taken to be 220 million people this translates into 4,400 excess lifetime deaths from cancer or 62.8 per year. For acrylonitrile at a concentration of I ~g/liter, the estimated lifetime risk for humans is 6.7 x 10-7. The upper 95% confidence estimate is 1.3 x 10-6. Both of these estimates are the averaged risks calculated from the male and female rats. They are based on preliminary data of Norris (1977) and are subject to change when the study is completed. CONCLUSIONS AND RECOMMENDATIONS Based on toxicological investigations, the Food and Agriculture Organi- zation/World Health Organization (FAD/WHO, 1965) concluded that an acceptable daily intake of acrylonitrile for humans could not be determined. Both epidemiological and controlled feeding and inhalation studies since 1965 indicate that acrylonitrile is a carcinogen. Therefore, it would not seem possible to establish a long-te~ acceptable level for this compound in drinking water. Unfortunately, since LD50 studies indicate only the level of short-term toxicity, short-te~ exposure limits cannot be calculated.

Toxicity of Selected Drinking Water Contaminants 77 Antimony (Sb) Antimony is a metal that is chiefly a by-product of base metal and silver ores. It has oxidation states of + 3 (trivalent) and + 5 (pentavalent) and forms compounds with halides, oxygen, sulfur, and organic anions such as tartrate, thioglycollate, and thioglycollamide. It is used industrially in flameproofing textiles, in vulcanizing of rubber, and in the manufacture of paint pigments, electronic semiconductors, thermoelectric devices, and fireworks (National Institute for Occupational Safety and Health, 1978; Robert and Boston, 1974~. Antimony compounds also have medical application (Gross, 1974; Harvey, 1975~. Organic antimonial compounds are used as parasiticides to treat different forms of schistosomiasis, bilharziasis, and leishmaniasis. Most exposure of humans to antimony compounds occurs in industrial settings. Schroeder (1970) has estimated the human daily intake from all sources to be about 100 ,ug. METABOLISM Trivalent and pentavalent antimony are differently distributed and excreted. Trivalent compounds have a great affinity for erythrocytes and, therefore, give low plasma concentrations. Pentavalent compounds tend to remain in the plasma. The trivalent form is excreted at a much slower rate in the urine than pentavalent antimony probably because it collects at much lower levels in plasma (National Institute for Occupational Safety and Health, 1978; Robert and Boston, 1974~. After administration of a single therapeutic dose of trivalent antimony only logo was recovered in 24 hr. whereas 5097O of the pentavalent form was recovered in 24 hr (Harvey, 19751. Most trivalent antimony (tartar emetic) is excreted in the feces, whereas the pentavalent forms are excreted mainly in the urine. The distribution of antimony to the tissues has not been thoroughly studied. However, in guinea pigs the trivalent compounds are found in high concentration in the thyroid and liver while the pentavalent forms are found in the liver and spleen after oral dosing. Abdallah and Saif (1962), in their studies of humans, showed that the highest concentrations of antimony occur in the liver, followed by the thyroid and heart. They administered sodium antimony dimercaptosuccinate (~24Sb) intra- venously. The liver, heart, and thyroid retained antimony for 20 days. When three 100-mg doses of antimony were administered intramuscular- ly over 9 days there was still considerable antimony in these tissues 53 days later.

78 DRINKING WATER AND H"LTH Antimonial compounds are absorbed very slowly from the gastrointes- tinal tract. Therefore, when used medicinally, they are administered parenterally (Harvey, 1975~. HEALTH ASPECTS Antimony is not essential to the life and health of humans or animals. It resembles arsenic both chemically and biologically, and symptoms of acute and chronic toxicity from antimony closely resemble those induced by arsenic. Generally, the trivalent compounds are more toxic than the pentavalent. Reports of toxic reactions to antimony are few. When they are observed, they are usually a result of industrial exposure or antimonial treatment of parasitic diseases (National Institute for Occupational Safety and Health, 1978~. Observations in Humans The major toxic symptoms that are associated with antimonial compounds in humans involve the gastrointestinal tract, heart, respiratory tract, skin, and liver. The most serious effects of these compounds, which are exerted on the heart, have been observed during antimonial therapy and during industrial exposure (National Institute for Occupational Safety and Health, 1978~. Such cardiac alterations include changes in the electrocardiogram, primarily in the T-wave (suppressed amplitude? inversion of the T-wave, and prolongation of the Q-T interval); bradycardia; and fluctuations in blood pressure (Brieger et al., 1954; National Institute for Occupational Safety and Health, 19781. Respiratory changes include irritation of the mucous membranes upper respiratory tract irritation, and, more seriously, pneumonoco- niosis. Pneumonia has also been cited as a side effect to the therapeutic use of antimony. Gastrointestinal symptoms include cramps, nausea, pain, anorexia, diarrhea, and vomiting. There is little information concerning dose-response relationships. Many toxicities have occurred as the result of industrial exposure to inhaled antimony and are complicated by the presence of other agents (National Institute for Occupational Safety and Health, 1978; Robert and Boston, 1974) However, most toxic syndromes have been observed in patients receiving antimonial therapy, and many have been repro- duced in animal studies. Observations in Other Species Acute Effects The LD50 values for different antimony compounds vary considerably. Bradley and Frederick (1941) administered antimony

Toxicity of Selected Drinking Water Contaminants 79 compounds intraperitoneally to rats and found the following LD50 values: antimony potassium tartrate, 1 1 mg/kg; antimony trioxide, 3.25 g/kg; and antimony metal, 100 mg/kg. The oral LD50 for antimony potassium tartrate was 300 mg/kg. After the oral administration of antimony chloride to rats, Arzamastsev (1964) reported that the LD50 of trivalent antimony chloride (SbCl3) was 675 mg/kg, and that of the pentavalent compound (SbCl5), 1.1 15 g/kg. These compounds were somewhat more toxic in guinea pigs, i.e., the LD50 for SbCl3 was 574 mg/kg, and for SbCl5, 900 mg/kg. Chronic Elects Effects of chronic feeding studies vary depending upon the antimonal compound under study. Arzamastsev (1964) gave trivalent antimony chloride per os to rats for 10 days at 135 mg/kg. Toxic symptoms included myocardial degeneration from which two rats died during the treatment. Studies were also conducted using guinea pigs. Ten-day studies at 12 and 20 mg/kg doses of trivalent antimony chloride produced blood changes (decreased hemoglobin, increased reticulocyte count). A longer term study was performed using guinea pigs given trivalent antimony chloride orally in doses ranging from 0.0025 to 2.5 mg/kg for 6 months. No toxic symptoms were observed with the lowest dose. Robert and Boston (1974) reported work done in 1927 by Flury in which antimony potassium tartrate was administered to rats for 85 days in oral doses up to 200 mg/rat at which time the rats died. Feeding antimony trioxide in doses less than 2 g/rat/day did not give rise to observable toxic symptoms. Mutagenicity Paton and Allison ( 1972) investigated the effects of antimony sodium tartrate (2.3 nM) on human leukocytes in vitro. They examined 100 metaphases and observed that 12~ of the cells displayed chromatic breaks. The authors did not draw any conclusions. Carcinogenicity There is no direct evidence that antimony induces carcinogenicity. Some reports (National Institute for Occupational Safety and Health, 1978) suggest that deaths from lung cancer were higher in factory workers in Great Britain who had been exposed to antimony than in the general population. Because there were many complicating factors in this situation, no firm conclusions could be drawn. Kanisawa and Schroeder (1969) reported that 76 mice given 5 ,ug of antimony potassium tartrate per milliliter of drinking water through- out their lifetimes showed no increase in incidence of tumors.

80 DRINKING WATER AND H"LTH Teratogenicity A report by Belyayeva (1967) is the only recent investigation concerning the ejects of antimony on reproduction. This investigator compared reproductive indices between women working in an antimony metallurgical plant (located in the USSR) with a similar group that had not been exposed to antimony. The antimony workers had a greater incidence of spontaneous abortion, premature births, and gynecological problems (menstrual cycle disorders and inflammatory diseases) than did the control group. In animal studies, Belyayeva (1967) administered metallic antimony (< 5-,um diem) by inhalation at a dose of 50 mg/kg to 30 female rats. Only 15 rats became pregnant, some only after multiple matings. These rats produced fewer offspring (5.4 versus 7.8) than controls. Gross inspection of the placentas and newborns showed no morphological changes. CONCLUSIONS AND RECOMMENDATIONS Suggested Ho-Adverse-Response Level (SNARLJ The data from which to calculate either 24-hr or 7-day SNARL's are inadequate. If a chronic SNARL is calculated using the lowest no-observed-adverse-effect level of 0.0025 mg/kg, which was reported by Arzamastsev (1964), the resulting value is less than the estimated daily intake of antimony. However, maximal environmental levels have been set by others as follows: I. Trivalent antimony 0.05 mg/liter of potable water (Dawson, 1974) (based on USSR standard). 2. Antimony 0.05 mg/liter maximum permissible concentration, USSR (Stofen, 1973~. 3. National Institute for Occupational Safety and Health (NIOSH) occupational standard 0.5 mg/m3 for a 10-hr work shift in a Bohr workweek. The taste threshold for either trivalent or pentavalent antimony is 0.6 mg/liter (Arzamastsev, 19641. Benzene (C6H6) This compound was evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 688~. The following material, which

Toxicity of Selected Drinking Water Contaminants 81 became available after that 1977 publication, updates and, in some instances, reevaluates the information in the earlier report. Also included are some references that were not assessed in the original report. Benzene is produced by petroleum refining, coal tar distillation, coal processing, and coal coking. In 1976 the U.S. production of benzene exceeded 0.5 million kg. It is used primarily as a chemical intermediate in the manufacture of styrene, cyclohexane, detergents, and pesticides. Bowden (1972) reported that gasoline that is used in motor vehicles in the United States usually contains 0.8% by volume benzene, but the range extends to 2%. A survey of inhalation exposures for 12 personnel at service stations that dispensed gasoline containing 1.6% benzene by volume showed a maximum of 0.39 ,ug/liter (0.123 ppm) (Hartle and Young, 1976~. Benzene is somewhat soluble in water (0.82 g/liter). In the National Organic Monitoring Survey (NOMS), which was conducted from March 1976 through January 1977, benzene analyses were performed on three samplings from community water supplies, which were representative of various types of sources and treatment processes. The numbers of positive benzene analyses per numbers of cities sampled were 0/111, 7/113 (0.4 ~g/liter), and 4/16 (0.95 ~g/liter) (U.S. Environ- mental Protection Agency, 1978c). METABOLISM Whether administered by inhalation, orally, or by another route, benzene is eliminated rapidly by expiration and excretion in the urine. Parke and Williams (1953) administered i4C-labeled benzene orally to rabbits (0.34 0.5 g/kg) and collected samples over 3 days Benzene expired in air accounted for 4397O of the dose; 34.5~O was excreted in the urine as glucuronide or ethereal sulfate conjugates of the metabolic oxidation products phenol, quinol, catechol, and hydroxyquinol together with small amounts of other products; and 5% to 10570 remained in the tissues. Benzene is metabolized similarly in humans (Laskin and Goldstein, 19771. It is readily absorbed through the lungs and has a retention rate of about 50% (Srbova et al., 1950~. Inhalation at 52 to 62 ppm for 4 hr resulted in 30% retention after 3 hr (Nomiyama and Nomiyama, 19741. Hunter and Blair ( 1972) reported that humans retained 230 mg after exposure to 80 to 100 ~g/liter for 6 hr. Retained benzene is distributed in tissues according to their fat content. Bone marrow, which, in toto, is an organ about two-thirds of the size of the liver, has a high tissue blood partition coefficient for benzene. Metabo- lites are believed to be important in the development of hematotoxicity, partly because of the effect of altered liver metabolism upon leukopenia

82 DRINKING WATER AND H"LTH and other hematopoietic responses. But little is known of the metabolic fate of benzene in the bone marrow (Snyder and Kocsis, 1975~. HEALTH ASPECTS Observations in Humans The toxicity to the hematopoietic system of chronic exposure of humans to benzene is well documented. Reported effects include myelocytic anemia, thrombocytopenia, or leukopenia (occurring either separately or in cases of pancytopenia) and leukemia, particularly acute myelogenous and monocytic leukemia. In many of these studies benzene was exposed with additional solvents at relatively high concentrations. Data on the level and duration of exposure are inadequate for deriving dose-response relationships of chronic benzene toxicity (Vigliani, 1976~. Infante et al. (1977) reported a retrospective cohort study of two populations of workers who were involved in production of rubber sheeting (Pliofilm). Benzene was the only material in their work environment that was known to be associated with blood disorders. In both plants during 194~1949, the occupational exposure of 561 workers to benzene was apparently well within the maximum allowable concentration of 100 ppm that was usually recommended. Vital status to 1975, which was obtained for 7597O of the workers, showed a significant excess of leukemia in those exposed to benzene, indicating a 10-fold increase in risk of death from myeloid and monocytic leukemia. The onset of leukemia is usually preceded by many observable effects on the hematopoietic system (Snyder and Kocsis, 1975~. It is not known whether benzene causes leukemia as one aspect of its hematotoxic effects, whether the leukemia is a consequence of benzene-induced damage to immunological components of the bone marrow, or whether the leukemic effects are unrelated to the other hematopoietic manifesta- tions (Laskin and Goldstein, 1977~. The data suggest that benzene is a leukemogen in humans. The current U.S. standard sets maximum occupational exposure to benzene at 10 ppm as a time-weighted average for up to a 10-hr day. This standard is based on an accumulation of data indicating possible effects on the hematopoietic system from prolonged inhalation of benzene in the range of 20 to 60 ppm (National Institute for Occupational Safety and Health, 1974~. However, in consideration of the evidence for benzene as a leukemogen, OSHA in February 1978 proposed the adoption of 1 ppm (3.2 ~g/liter) benzene as a "lowest feasible level" in air as a time-weighted average for an 8-fur day with a ceiling of 5 ppm for any 15-min period (U.S. Department of Labor, 1978~.

Toxicity of Selected Drinking Water Contaminants 83 Effects of benzene on proliferation of bone marrow in humans include increased incidence of chromosomal aberrations with aneuploidy and breakage. In cases of frank hematopoietic toxicity, such changes might be due to the disease, but Forni et al. (1971) noted significant chromosome aberrations in cultured leukocytes that were obtained from workers who were exposed to benzene but showed no symptoms of benzene poisoning. Among workers who had recovered from other signs of benzene hematotoxicity, aneuploidy in leukocytes persisted for 10 years, but karyotypes appeared normal by 12 years (Pollini and Biscaldi, 1977). Observations in Other Species Acute Effects Six rabbits that were exposed by inhalation to 35,000 to 45,000 ppm underwent anesthesia after 3.7 min arid showed other effects in the central nervous system until death, which ensued after 22.5 to 71 min (Carpenter et al., 1944~. Kimura et al. (1971) studied acute oral toxicity in Sprague-Dawley rats. They used 6 to 12 rats of both sexes per group in testing newborn and 14-day-old rats, and 6 male rats per group for the other ages. Single dose LDso values (9597O confidence level) for newborn, 14-day-old, young adult, and old adult rats were <0.9, 3.0 (1.8-5.0), 3.3 (2.6~.2), and 4.9 (3.5-6.2) g/kg body weight, respectively. Chronic Effects Leukopenia is the most commonly observed effect of chronic benzene intoxication in laboratory animals. Deichmann et al. (1963) reported that repeated inhalation of as little as 61 ppm benzene by rats at 5 hr/day, 4 days/week for 6 or more weeks causes a significant decrease in circulating leukocytes. However, at lower exposures, a fall in leukocyte concentration typically leads to increased cell proliferation which causes cyclical fluctuations. Moreover, there is normally wide variation among cell counts during diurnal cycles and among individual animals. Laskin and Goldstein (1977) concluded that nearly all reports of leukopenia in animals exposed to less than 1,000 ppm benzene are unlikely to stand up to statistical evaluation. Benzene mixed with equal parts of olive oil was administered to rats by subcutaneous injection (Gerarde. 1956; Latta and Davies, 1941~. Weight loss and leukopenia resulted from doses of 880 mg benzene/kg body weight which were given daily for 14 days (Gerarde, 1956), and from doses of 1.32 g benzene/kg/body weight which were given daily for 3 to 60 days (Latta and Davies 19411. In Latta's group ~ rat that died after 10 days had hyperplastic bone marrow. and one that died at 21 days had acute leukopenia and hypoplastic bone marrow. Oral adminis

84 DRINKING WATER AND H"LTH tration of benzene to rats in daily doses of 1, 10, 50, and 100 mg/kg body weight during 132 days over 6 months resulted in leukopenia and erythrocytopenia at the lowest minimal effect level of 50 mg/kg (Wolf et al., 1956~. Mutagenicity Toxic effects on bone marrow cells of rats and other laboratory animals include changes in chromosome number and chro- mosome breakage that resemble those in humans. There was no clear evidence for dose-dependent response (Laskin and Goldstein, 1977~. Lyon (1975) used the Ames assay, with Salmonella typhimurium strains TA-98 and TA-100, to test benzene for mutagenicity in doses ranging from 0.1 to 1.0 p1/plate, both without and with microsomal fractions at concentrations from 1 to 50,ul/plate. Postmitochondrial supernatant suspensions of microsomes were prepared from liver homogenates from normal rats and from rats that had been treated with phenobarbital and 3-methylcholanthrene (MCA), and from the bone marrow of normal and MCA-treated rats. Benzene was uniformly negative in all these assays and was also inactive in the dominant lethal assay in rats. Carcinogenicity No benzene-induced carcinogenesis has been ob- served in an animal system. An extensive study with mice (Ward et al., 1975) indicated some increase in granulocytic leukemia. After reviewing the data from that study, the National Academy of Sciences Safe Drinking Water Committee concluded that the increase is not statistical- ly significant, even when time to response is incorporated into the analysis (National Academy of Sciences, 1977, p. 690~. Teratogenicity The subcommittee noted only one study of teratogen- icity. In this study, Watanabe and Yoshida (1970) gave subcutaneous injections of acute toxic doses of benzene (3 ml/kg body weight) to pregnant mice on days 11 through 15 of gestation. Malformations were most prevalent in the group that was treated on the 13th day. Four of 15 litters, involving 10 of 127 fetuses, had cleft palate, agnathia, or micrognathia (Watanabe and Yoshida, 1970~. Decreased white cell count and weight gain in the benzene-treated mice were the same whether the litters were normal or included malformed fetuses. As an indication of the toxicity of the dose used in this experiment, five male mice that received benzene at 3 ml/kg body weight survived, while four of five male mice that had been injected with 4 ml/kg died within 3 days. Carcinogenic Risk Estimate As noted above, there are no data from animal models for use in extrapolation. Occupational studies on human

Toxicity of Selected Drinking Water Contaminants 85 exposure (Aksoy e! al., 1972, 1974a,b, 1976; Ishimaru e! al., 1971; Thorpe, 1974) do not contain adequate information on degree of exposure or size of population at risk. In addition, the workers in benzene-related occupations typically were exposed to other chemicals, as in the study reported by Ott et al. (1978~. Consequently, extrapolation of benzene-induced cancer risk from such data as these would be tenuous. In a study by Infante et al. (1977) workers were exposed to benzene as the sole chemical suspected of affecting the hematopoietic system. In these cases, benzene concentrations apparently were high during the first years of exposure and were lower thereafter. There are no data indicating how often short exposures at elevated levels may have occurred. Estimates of actual exposure are inadequate for extrapolation for risk of benzene-induced leukemia. CONCLUSIONS AND RECOMMENDATIONS The acute ejects of benzene cover a wide range of signs and symptoms. Thee effects are transitory but may lead to more lasting chronic effects such as anemia. If exposure is continuous and great enough, leukemia is a strong possibility for susceptible members of the population. There are no dose-response data on animals, and the data on humans are inadequate to calculate a risk estimate for benzene with mathematical models. Suggested No-AdYerse-Response Level (SNARL) Twenty-Four-Hour Exposure There are no adequate data from which this calculation can be made. Seven-Day Exposure Wolf et al. (1956) administered benzene orally to rats 5 days/week for 6 months. They reported leukopenia and erythrocytopenia at the daily dose level of 50 mg/kg. Quantitative effects were not reported but this could be considered a minimal effect. Adjusting to a 7-day exposure, 50 mg/kg/day x 5/7 days = 36 mg/kg/day, applying a safety factor of 100, and assuming total exposure to be from 2 liters of drinking water per day during this period: 36 mg/kg x 70 kg 100 x 2 liters = 12.6 mg/liter. This SNARL ignores the established mutagenicity and suspected carcinogenicity of benzene.

86 DR!NKING WATER AND H"LTH Chronic Exposure No chronic SNARL can be calculated because benzene is a suspected human carcinogen. In summary, there is no adequate source of data (animal or human) on which to base a statistical extrapolation from high to low exposure. More studies on the role of metabolism and mode of action of benzene metabolites in the bone marrow are needed, particularly on their relationship to chromosome aberrations and mutation and to immuno- logical systems. More data are also needed on the teratogenicity and cocarcinogenicity of benzene. If there are data on industrial exposure to benzene, then systematic monitoring should be started with a view to following the population groups at risk. Before chronic limits for benzene in drinking water can be established, more extensive toxicological data must be gathered and evaluated. In the absence of such data, it is useful to consider what portion of the average dose of benzene is contributed by its presence in drinking water. Benzene is a natural constituent of fruit, fish, vegetables, dairy products, nuts, eggs, and rum, and has been found in cooked meat. In these forms, it provides humans with an estimated daily intake of 250 ,ug/day (Kraybill, 1977~. The benzene content of ambient air at automobile service stations was recorded at 0.39 ,ug/liter (Hartle and Young, 1976~. The general urban atmosphere contains much less: an unreferenced value was 0.05 ,ug/liter. At this concentration, assuming a daily inhalation of 24 me and a retention of 50%, the dose from inhalation would be 600 ,ug/day. Based on the same assumptions, the current OSHA recommendation for a 1~ hr time-weighted average of 3.2 ,ug/liter (1 ppm) as a lowest feasible level in the workplace would allow a daily dose of 16 ma. From food and air the average urban dweller receives approximately 850 ,ug of benzene daily. In 1978 the level of benzene in U.S. drinking water was reported to be 1 ,ug/liter, or a 2-,ug daily dose, assuming a 2 liter/day intake (U.S. Environmental Protection Agency, 1978c). Although such exposure is less than 0.3% of an average daily intake, major spills of benzene into drinking water could result in much higher exposures, since the solubility of benzene in water is 820 ,ug/liter. Benzene Hexachloride (BHC) and Linda ne (C6H6CI6) These compounds were evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 583~. The following material, which became available after that 1977 publication, updates and, in some instances, reevaluates the information in the earlier report. Also included are some references that were not assessed in the original report.

Toxicity of Selected Drinking Water Contaminants 87 HEALTH ASPECTS Observations in Humans In a study of the racial stratification of organochlorine insecticide residues, Kutz et al. (1977) found lindane residues in samples from Negroes more than twice as often as in samples from Caucasians. There was little racial difference noted for residues of ,0-benzene hexachloride (BHC). Observations in Other Species Acute Effects Kashyap et al. (1976) treated Wistar rats with technical BHC (15% gamma isomer) by intubation of a single dose as a 5% solution in olive oil. Ten rats per group were treated with 100, 200, 300, 400, and 500 mg/kg of body weight. The animals were observed for 48 hr and then sacrificed. The investigators observed mortality, physical symptoms, serum glutamic pyruvic transaminase (SGPI), prothrombin time, coagulation time, cholinesterase activity, and histology of the liver and other organs at autopsy. Results indicated a significant increase in SGPT at all treatment levels, which indicates liver damage. Prothrombin time and coagulation time had also increased, even in animals that had received the lowest doses. Mutagenicity Ishidate and Odashima (1977) obtained negative results for lindane when screening for chromosomal aberrations in Chinese hamster cells in vitro. Published reports indicate that lindane does not have significant mutagenic potential (National Academy of Sciences, 1977~. Carcinogenicity Thorpe and Walker (1973) fed groups of male and female CFI mice diets containing ,B-BHC at 200 mg/kg or y-BHC at 400 mg/kg. Liver tumors were observed in 24% of the male and 23% of the female controls, in 73~o of the males and 43% of the females that had received 200-ppm ,l?-BHC, and in 9397O of the males and 60% of the females that had received 400-ppm y-BHC. Lung metastases were found in some males that had received ,8- and y-BHC and in some females that had received y-BHC. The incidence of other tumors was not increased by exposure to either isomer. However, the dose of 400 ppm of y-BHC may be higher than the maximum tolerated dose. Therefore, nonspecific toxicity may have biased the observed effects. Only 3% of the females and 17~o of the males that were fed lindane survived for the duration of the experiment at this dose, which is approximately 70% of the acute oral LDso.

88 DRINKING WATER AND H"LTH Weisse and Herbst (1977) treated groups of 100 mice (Chbi:NMRI [SPF]) of both sexes with 12.5, 25, and 50 ppm lindane. There were 200 controls. The concentrations represent 2.4, 4.1, and 8.2 mg/kg body weight for males and 2.0, 3.9, and 7.8 mg/kg body weight for females. The study lasted 80 weeks. No lindane-related production of tumors was evident at these dosages. Electron microscopic examination of the livers produced no evidence of lindane-induced fine-structural hepatocellular alterations. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level(SNARL) Twenty-Four-Hour Exposure The data of Kashyap et al. (1976) provide a minimal-adverse-response level of 100 mg/kg for technical BHC (15% lindane) in rats. The concentration suggested for a 24hr SNARL assumes that lOO~o of the exposure to the compound is provided by drinking water during that period and that a 70-kg human consumes 2 liters/day of water. Applying an uncertainty factor of 1,000, the following calculation may be made: 100 mg/kg x 70 kg 1,000 x 2 liters = 3.5 mg/liter. Seven-Day Exposure Since there are no data on subchronic oral administration of BHC to animals, it is possible to calculate a 7-day SNARL by dividing the 24-hr value by 7. This result is 0.5 mg/liter. Chronic Exposure This value cannot be calculated because technical BHC is a carcinogen in animals. Bis(2-Chloroethylkther (ClCH2CH2=H2CH2Cl) This compound was evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 710~. The following material, which became available after that 1977 publication, updates and, in some instances, reevaluates the information in the earlier report. Also included are some references that were not assessed in the original report.

Toxicity of Selected Drinking Water Contaminants 89 HEALTH ASPECTS Observations in Humans No new data. Observations in Other Species Mutagenicity After giving three different doses to mice by Savage in the heritable translocation test, Jorgenson et al. (1977) determined that the mutagenic potential for this compound was negative. It was weakly mutagenic in a Salmonella assay (TA-100 without S-9 mix) when incorporated into agar and very mutagenic when assayed in desiccators or in suspension (Simmon et al., 1978~. Carcinogenicity Theiss et al. (1977) tested for pulmonary tumor induction in strain A/St male mice by giving them intraperitoneal injections of three dose levels 3 times a week for a total of 24 injections. Upon sacrifice, 24 weeks after the first injection, results were negative. Survival at 20 mg/kg, which was given 24 times (total dose of 480 mg/kg), was lOO~o, whereas at 40 mg/kg, which was given 4 times (total dose of 160 mg/kg), it was 75%. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARLJ Acute Exposure Oral LD50 values in three laboratory species average approximately 125 mg/kg. Using that LDso as the basis of extrapolation to humans, and assuming a 70-kg body weight and 2-liter consumption of water per day, the LD50 in humans would be 4.3 mg/ml. Because the solubility of the compound in water is approximately 10 mg/ml, an acutely toxic dose in humans is theoretically possible. Since only acute LD50 animal data (no sublethal acute dose levels) were available, a 24hr or 7-day SNARL cannot be calculated. Chronic Exposure Since the compound is a "suspected animal carcinogen" (Drinking Water and Health, Table VI-60, p. 794, National Academy of Sciences, 1977), calculation of a chronic SNARL would not be appropriate. Further studies should be conducted on this compound if it is found to be present in a large number of water supplies.

90 DRINKING WATER AND H"LTH Bis(2-Chloropropylkther (CH3CHCICH2OCH2CHCICH3) Bis(2-chloropropyl~ether is a by-product of the manufacture of propyl- ene glycol. It is commonly dumped into streams where it is stable. METABOLISM Smith et al. (1977a) conducted studies in rats and monkeys on blood levels, blood level kinetics, tissue distribution, and excretion. Half-life determinations were made for 2 days in the rat, and 5 hr (a phase) and 2 days ~ phase) in the monkey. There were significant accumulations of bis(2-chloropropyl~ether in the livers of monkeys. Excretion was essen- tially complete in both species at 24 hr. Its metabolites, 1-chlor - 9- propanol, propylene oxide, and 2~1-methyl-2-chloroethoxy~propionic acid, were all found in the urine. HEALTH ASPECTS Observation in Humans No available data. Observations in Other Species Acute Elects Smyth et al. (1951) reported the acute oral LD50 in rats to be 240 mg/kg. In the guinea pig, Spector (1956) found it to be 450 mg/kg. In inhalation studies in rats, 700 ppm killed all five animals after 6 hr. two of five died after a 6-fur exposure to 350 ppm, and one of four died after an 8-fur exposure to 175 ppm (Hake and Rowe, 1963~. Chronic Effects Twenty-two gastric intubations in rats for 31 days at 10 mg/kg and 200 mg/kg resulted in decreased growth. The 200 mg/kg exposure also resulted in increases in the weights of the liver, kidney, and spleen (Hake and Rowe, 1963~. Mutagenicity A test for mutagenic potential by the heritable translo- cation test (Jorgenson et al., 1977) was negative in mice that received three different oral doses by Savage. The compound was weakly mutagenic in the Ames Salmonella assay using the agar-incorporatioI1 system and very mutagenic in desiccator or suspension assay systems. The mutagenic activity was enhanced by the addition of an S-9 activation mix that was prepared from human liver (Simmon et al., 1 978~.

Toxicity of Selected Drinking Water Contaminants 91 Carcinogenicity No available data. Teratogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARL) Acute Exposure Using the only available data, which is an LDso of 240 mg/kg in the rat, and extrapolating directly to humans, assuming 70- kg body weight and 2-liter consumption of water per day, the LD50 for humans would be 16.8 g. This amounts to 8.4 mg/ml. Since the solubility of this compound in water is 1.7 mg/ml, it is unlikely that an acutely lethal dose in humans would be possible. Data are inadequate for calculations of a 24-hr or 7-day SNARL. Chronic Exposure There are insufficient data for estimates of chronic toxicity. Long-term oral experiments should be conducted in at least two species of animals so that chronic risk can be calculated. In view of the probable frequent contamination of water supplies by this compound and its structural similarity to other carcinogenic haloethers, additional short- and long-term toxicity data are needed. Cadmium (Cd) Cadmium is a silvery-white metal that is found in the same periodic group (JIB) as mercury and zinc. In industry, it is used principally in electroplating, in the manufacture of pigments, as a plasticizer, in batteries, and in electrical conductors (Friberg et al., 1974; National Institute for Occupational Safety and Health, 1 976a). Although not abundant, cadmium is distributed widely through the crust of the earth and is found wherever zinc is found in nature. Cadmium in the environment originates primarily in industrial sources (Webb, 1975~. Concentrations of cadmium in unpolluted fresh water vary around 1 g/liter. The drinking water standard is 10 ~g/liter. The major source of cadmium for the general population is diet. Average daily intake values may vary between 50 and 150 ,ug/day. Cigarette smoke contains approximately 1 ppm cadmium. 1 hose who smoke one pack of cigarettes daily are exposed to 2 to 4 fig of cadmium per day from that source (Friberg et al., 1974~. Occupational exposures to cadmium are consider- able, averaging approximately 0.02 to 0.05 mg/m3.

92 DRINKING WATER AND H"LTH METABOLISM After cadmium is ingested or inhaled, it is distributed to most tissues of the body, but is found in highest concentrations in the liver and kidney. Within the various organs cadmium is bound primarily to a unique low molecular weight protein, termed metallothionein (Friberg et al., 1974; Webb, 1975~. Following oral administration, approximately 1% to 2% of the dose is absorbed in laboratory animals, while estimates of absorption in humans range up to 5% (World Health Organization, 1977~. Variations may be induced by many factors such as age, dietary calcium, and dietary protein. Excretion of cadmium occurs primarily via the kidney at a very slow rate. Consequently, the accumulation of cadmium in humans is considerable. In fact, cadmium levels in the newborn are virtually negligible but may reach body burdens of 30 mg by age 30 (Schroeder, 1965~. The biological half-life of cadmium is quite long and is estimated to be on the order of decades in humans (World Health Organization, 1977~. HEALTH ASPECTS Cadmium is a very toxic element. It is not a trace element that is essential to nutrition. Toxicity from this element can result from acute or chronic exposure, and the toxic syndromes may differ. Certain toxic ejects may be prevented by the administration of such trace elements as zinc and selenium. The health aspects of cadmium have been reviewed extensively (Fleischer et al., 1974; Nationa! Institute for Occupational Safety and Health, 1976a; Friberg et al., 1974; Webb, 1975; World Health Organization, 19771. The following synopsis was derived from these sources unless otherwise noted. Observations in Humans Acute exposures to cadmium in humans produce different effects depending upon the route of exposure. Among the general population, ingestion of food and fluids that have been contaminated with cadmium resulted in acute gastrointestinal distur- bances such as nausea, vomiting, abdominal pain, diarrhea, and tenesmus. This contamination has occurred when cadmium-plated vessels have been used to prepare lemonade or other acidic beverages. The acute effects of industrial exposure to cadmium are generally manifested as lung damage, which may result in death. Typically, the symptoms include chest pain and pulmonary edema. Pulmonary fibrosis has been observed in survivors, but there are few data pertaining to the

Toxicity of Selected Drinking Water Contaminants 93 acute dose-response relationship for cadmium toxicity in humans. Estimates of lethal concentrations vary from 2,500 to 2,900 mg/m3. Chronic exposure to cadmium may also cause a wide variety of effects involving many organs and systems. The severity of the manifestations depend upon the magnitude and duration of exposure. The two major effects of chronic cadmium toxicity in persons that have been occupationally exposed to cadmium are obstructive lung disease and renal dysfunction. The lung disorders are primarily sugges- tive of pulmonary emphysema. Dose-response relationships are uncer- tain because time-weighted averages are either not available or available only for short periods. The most common abnormality from chronic cadmium exposure involves renal toxicity characterized by proteinuria. Other disturbances of renal tubular function include glycosuria, amino aciduria, decreased urine concentrating ability, and abnormalities in renal processing of uric acid, calcium, and phosphorus. Proteinuria can occur along with these other changes. Therefore, it may be the earliest sign of renal dysfunction in cadmium toxicity. After prolonged expo- sures, the kidney is generally regarded as the critical organ. Autopsy data, as well as animal data, indicate that the critical level of cadmium that produces damage in the kidney is approximately 150 to 250 Agog wet tissue. Other manifestations of chronic occupational cadmium exposure include olfactory changes such as anosmia, anemia, and osteomalacia. Excessive exposure to cadmium has also occurred in the general population through the ingestion of food, e.g., rice, and of drinking water containing high concentrations of cadmium. Such exposure has been implicated in itai-itai disease, which is characterized by osteomalacia and renal tubular dysfunction. Although cadmium has been widely implicat- ed in this disorder, there may be other disposing factors involved as well. The role of cadmium as a predisposing factor in hypertension is controversial. While epidemiological investigations have found a positive correlation between cardiovascular disease (including both hypertension and atherosclerotic diseases) and ambient cadmium levels, no direct cause-effect relationships have been established. Moreover, workers that have been exposed to cadmium do not exhibit a higher prevalence of hypertension than do other groups. Observations in Other Species A cute Effects After acute exposure to relatively high doses of cadmium, toxic effects are seen primarily in the gonads (testicular

94 DRINKING WATER AND H"LTH necrosis), liver (structural and functional changes), kidney (tubular damage), and central nervous system (hemorrhagic lesions of sensory ganglia). Sarcomas at the injection site of cadmium have also been noted (see section on carcinogenicity). The LD50 value in male rats after oral administration of cadmium varies from 175 to 225 mg/kg (Kotsonis and Klassen, 1977; Schnell et al., 19781. Carcinogenicity There has been considerable interest in the possibili- ty that cadmium exposure may cause cancer. Studies by Gunn et al. (1963, 1967) have shown that injection of 0.03 mmol/kg cadmium chloride could induce subcutaneous sarcomas at the injection site as well as testicular tumors (interstitial cell tumors) in rats and mice. Concomi- tant administration of zinc prevented development of such tumors. No carcinogenic elects were found by Schroeder et al. (1965) who fed cadmium to mice in concentrations of 5 ppm in drinking water over their lifetime. Several epidemiological surveys have suggested that there is an increased incidence of prostate cancer in cadmium workers when compared to the general population (Kipling and Waterhouse, 1967; Lemen et al., 1976; Potts, 1965~. However, because these studies had small samples, age groups varying from 60 to 75 years, and concurrent exposure to other environmental contaminants, firm conclusions con- cerning the carcinogenicity of cadmium in humans cannot be drawn. Mutagenicity There is little information concerning the mutagenic effects of cadmium. Epstein et al. (1972) reported that cadmium does not produce an increase in dominant lethal mutations in mice. In a study by Gilliavod and Leonard (1975), male mice (BALB/c) received 1.75 mg/kg cadmium chloride intraperitoneally and then were mated. There was no increased incidence of dominant lethal mutations in the offspring. In the treated males, dividing spermatocytes did not have chromosomal rearrangements such as reciprocal translocations. Conflicting results have been reported for humans. Paton and Allison (1972) did not find that cadmium induced chromosomal damage in vitro in cultured lymphocytes, while Shiraishi et al. (1972) did report such damage. Bui et al. (1975) performed chromosomal analyses on Swedish workers who had been exposed to cadmium and Japanese itai-itai patients. They found no increased frequency of chromosome aberrations. Teratogenicity When injected parenterally, cadmium induces terato- genic effects in laboratory rats, mice, and hamsters. Cherno~ (1973) administered cadmium chloride (~12 mg/kg) on days 13 to 16 of gestation in CD strain rats and found a dose-related increase in fetal

Toxicity of Selected Drinking Water Contaminants 95 deaths, a decrease in fetal weight, and an increase in the rate of anomalies such as micrognathia, cleft palate, club foot, and small lungs. Ishizu et al. (1973) gave subcutaneous injections of cadmium chloride (0.33-0.35 mg/kg) to mice on day 7 of gestation. The most common anomaly was exencephaly in fetuses that were removed on day 18. Other anomalies included spine bifida, absence of tail, and malformations of ribs, skull, and vertebrae. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARLJ Twenty-Four-Hour Exposure There are no adequate data from which to calculate a 24-hr SNARL. Seven-Day Exposure This 7-day SNARL is based on data of Loeser and Lorke (1977) who fed cadmium chloride to rats in their diets in concentrations of 1 to 30 ppm for 3 months without effect. Assuming that the rats consumed 20 g/day of food and that their average weight was 250 g, their exposure is calculated: 30 mg/kg/day x 0.02 kg/day 0.25 kg = 2.4 mg/kg. For a 70-kg human consuming 2 liters/day, using a safety factor of 1,000 and assuming that 100% of exposure is from water during this period, the 7-day SNARL is calculated: 2.4 mg/kg x 70 kg 1,000 x 2 liters = 0.08 mg/liter. Chronic Exposure The SNARL for chronic exposures is based on data of Decker et al. (1958), who gave cadmium to rats in their drinking water at a concentration of 10 ppm for 1 year without effect. They did observe effects such as anemia after exposing rats in the same manner to 50 ppm of cadmium for 3 months. In this study the rats consumed an average of 30 ml/day. Assuming that their average weight was 400 g, their daily exposure is calculated: 10 mg/liter x 0.03 liter/day 0.4 kg = 0.75 mg/kg. Using a safety factor of 1,000 for a 70-kg human consuming 2

96 DRINKING WATER AND H"LTH liters/day and assuming that 20~o of exposure is from water, the chronic SNARL is calculated: 0.75 mg/kg x 70 kg x 0.2 1,000 x 2 liters = 0.005 mg/liter. Carbon Tetrachloride (CCI4) This compound was evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 7031. The following material, which became available after that 1977 publication, updates and, in some instances, reevaluates existing information in the earlier report. Also included are some references that were not assessed in the original report. HEALTH ASPECTS Observations ir' Humans Van Oettingen (1964) reported that accidental ingestion of 14 to 20 ml has repeatedly caused fatal poisonings within 3 to 5 days; however, larger doses have occasionally not been fatal. Gosselin et al. (1976) reported that the mean lethal dose of carbon tetrachloride for humans lies between 5 and 10 ml (oral) although as little as 2 ml has caused death. Symptoms and/or death following ingestion occur after a latent period of 24 to 36 hr (Gosselin et al., 1976~. Observations in Other Species Acute Effects A single oral Savage of 0.25 mg/kg in adult male rats produces increases in liver tyrosine a-ketoglutarate transaminase (200~o3 and plasma alanine cY-ketoglutarate transaminase (120~o3 within 5 hr (Murphy and Malley, 1969~. Oral doses of 2.5 ml/kg result in decreases in liver and lung microsomal cytochrome P-450 and in decreases in the rates of metabolism of several substrates for the mixed function oxidase system (Chen et al., 1977~. These effects have been observed as early as a few hours after administration. Carbon tetrachloride itself is not thought to be the active agent; rather, it is probably the 'CCl3 free radical, which is produced via the cytochrome P-450 system. Carbon tetrachloride induces lipid peroxidation, which in turn is responsible for the mo~pho- logical changes (Smuckler, 1976~. The toxicity of carbon tetrachloride can be modulated by agents that alter levels of liver cytochrome P-450. Phenobarbital increases liver P-450 and enhances carbon tetrachloride

Toxicity of Selected Drinking Water Contaminants 97 toxicity. On the other hand, 3-methylcholanthrene, which induces the formation of P-448, does not enhance carbon tetrachloride toxicity. In fact, there is evidence that it may even lower it (Suarez and Bhonsle, 1976~. Cobaltous chloride, which inhibits P-450 selectively, decreases carbon tetrachloride hepatotoxicity (Suarez and Bhonsle, 1976~. Doses of carbon tetrachloride that induce biochemical and morpho- logical changes in rat liver also induce an increase in bile duct-pancreatic fluid flow (BDPF) by a mechanism that is not understood. However, the increase in BDPF appears to be unrelated to the time course and the extent of liver necrosis and elevation of serum glutamic-oxaloacetic transaminase (SOOT) (Harms et al., 1976; Imamura et al., 1977; Peterson and Fujimoto, 1976~. Chronic Effects In long-term studies of carbon tetrachloride toxicity, Alumot et al. (1976b) fed rats diets that had been fumigated with carbon tetrachloride. Doses of 150 to 275 ppm had no effect on the growth of weanlings, but 520 ppm depressed weight gain significantly within 6 weeks. Females of a similar age were unaffected by the treatment. Triglycerides and total lipids were significantly higher in animals that were treated with the carbon tetrachloride, with the exception of those that were fed 150 ppm. The results of growth, fertility, and biochemical tests of rats that received 80 and 200 ppm for 2 years were not significantly different from those of controls. Pathological studies were not conducted. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARL) The following calcula- tions are for noncarcinogenic effects only. Twenty-Four-Hour Exposure Animal data indicate that 0.25 ml/kg (0.4 g) is the lowest dose that produces a toxic effect. Using a safety factor of 1,000 and assuming that 100% of the exposure will come from 2 litersiday intake, the 24-hr SNARL is calculated as follows: 0.4 g/dlay x 70 kg 1,000 x 2 liters = 14 mg/liter. Seven-Day Exposure Assuming that repeated daily intake of carbon tetrachloride produces cumulative effects, the 7-day SNARL value should be one-seventh of the 24-hr exposure limit. This is 2 mg/liter.

98 DRINKING WATER AND H"LTH Chronic Exposure This value will not be calculated because carbon tetrachloride is a carcinogen in animals. Ethylene Dibromide (1,2-Dibromoethane) (BrCH=CHBr) Ethylene dibromide, also known as 1,2-dibromoethane and ethylene bromide, is a clear, colorless, heavy liquid at room temperature and has a distinctive odor. The water solubility of ethylene dibromide is 0.43 g/ 100 g of water at 30°C (National Institute for Occupational Safety and Health, 1977~. Its half-life in water at 20°C (pH 7) is approximately 14 years. The U.S. production of ethylene dibromide increased from an estimated 29 million kg in 1940 to over 150 million kg in 1973. Approximately 85% of the ethylene dibromide that has been produced has been used in gasoline as a constituent of antiknock mixtures containing tetraethyl lead. It is also used as a fumigant-insecticide, nematocide, and specialty solvent for resins, gums, and waxes (U.S. Environmental Protection Agency, 1976~. METABOLISM Ethylene dibromide is a reactive molecule that may form covalent bonds under certain conditions. From the data of Edwards et al. (1970) and Plotnick and Conner (1976), an approximate biological half-life using ~4C-labeled ethylene dibromide in mice and in guinea pigs can be estimated to be less than 48 hr. Twenty-four hours after intraperitoneal administration of 30 mg/kg of ethylene dibromide to guinea pigs, 65% was excreted as metabolites in the urine and 12% unchanged in the expired air. Similar findings were reported for mice. After administering intraperitoneal injections of i4C-ethylene dibro- mide to rats and mice, Shih (1976) found that the compound was widely distributed, with concentrations in the liver, kidney, and small intestine. At 24 hr the liver and kidney contain irreversibly bound i4C in RNA, DNA, and protein. Ethylene dibromide appeared to be a substrate for at least two different enzymes including glutathione-S-transferase. HEALTH ASPECTS Observations in Humans There are several case reports of human exposure to ethylene dibromide. In one, death occurred 54 hr following oral ingestion of 140 mg/kg of ethylene dibromide by a 43-year-old female. Postmortem examination disclosed major toxic elects in the liver

Toxicity of Selected Drinking Water Contaminants 99 and kidneys. Other reports indicate severe contact irritation both to the skin on typical exposure and lungs following inhalation. Systemic manifestations include malaise and vomiting. Observations in Other Species A cute Elects Single oral dose lethality studies in several animal species revealed LDso's ranging between 420 mg/kg and 55 mg/kg. Rabbits were the most sensitive and mice the least sensitive (Rowe et al., 1952b). Schlinke (1969) produced symptoms and/or death in sheep and calves by giving them single oral doses of 25 to 50 mg/kg. A 10 mg/kg dose in calves produced no ill effects. No autopsies were performed. Chronic Elects Chronic toxicity has been studied mostly by inhala- tion exposures. On the basis of data that were gathered from repeated inhalation exposures in animals, McCollister et al. (1956) estimated maximum safe concentrations and exposure times for humans that are exposed to undiluted ethylene dibromide to be 192 mg/m3 for a 7-fur single exposure. In chronic feeding studies, Alumot et al. (1968) gave feed that was contaminated with ethylene dibromide to hens in order to assess effects on weight gain and egg production. Concentrations as low as 40 ppm of ethylene dibromide administered for 10 weeks produced one or more adverse effects. Other investigators have reported significant egg weight reduction in chickens that were fed 5 to 7.5 ppm ethylene dibromide in grain (Bond) et al., 1955~. Amir and Volcani (1965) observed abnormal spermatozoa in Friesian bulls that had been fed an average of 1 mg/kg/day of ethylene dibromide from the age of 4 days until they were approximately 24 months old. Mutagenicity Buselmaier et al. (1972) reported that ethylene dibro- mide is mutagenic in the host-mediated assay and in in-vitro studies with Salmonella typhimurium. The authors concluded that ethylene dibromide did not require activation through in-vivo metabolism to exert its mutagenic ejects on Salmonella typhimurium, nor did metabolism of ethylene dibromide sufficiently deactivate its mutagenic potential, since both the host-mediated assay and the in-vitro plate tests were positive. dibromide was mutagenic in fruit flies (Drosophila Ethylene dibromide was mutagenic melanogaster) in tests by Vogel and Chandler (1974) Carcinogenicity A number of investigators have reported ethylene dibromide to be carcinogenic (National Institute for Occupational Safety and Health, 19771. Olson et al. (1973) exposed rats and mice via chronic

100 DRINKING WATER AND HEALTH oral intubation 5 times per week to experimentally predetermined maximally tolerated doses (MTD) and one-half MTD. They found a high incidence of squamous cell carcinoma of the stomach of both species. lathe initial doses were 80 and 41) mg/kg/day for rats and 120 and 60 mg/kg/day for mice. Weisburger (1977) reported that oral ethylene dibromide caused stomach tumors with many metastases in both rats and mice. Teratogenicity No available data. Carcinogenic Risk Estimate In a recent study by Olson et al. (1973), rats and mice were Savaged with doses of ethylene dibromide chloroethylene ranging from 40 to 120 mg/kg body weight for as long as 78 weeks. There was a high incidence of metastatic carcinoma of the stomach in both the mice and the rats. Each set of dose-response data was used to estimate statistically both the lifetime risk and an upper 95% confidence limit on the lifetime risk at the low dose level. These are estimates of lifetime human risks that have been corrected for species conversion on a dose/surface area basis. The risk estimates are expressed as a probability of cancer after a lifetime consumption of 1 1iter/day of water containing 1 ,ug/liter of the compound. For example, a risk of 1 x 10-6 implies a lifetime probability of 2 x 10-5 of cancer if 2 liters/day were consumed and if the concentration of the carcinogen was lO,ug/liter. This means that during a lifetime exposure, this compound would be expected to produce one excess case of cancer for every 50,000 persons exposed. Assuming the population of the United States to be 220 million people, this translates into 4,400 excess deaths from cancer or 62.8 per year. For ethylene dibromide at a concentration of 1 ,ug/liter the estimated lifetime risk for humans is 7.0 x 10-6. The upper 95% confidence estimate is 9.1 x 10-6. Both of these estimates are the averaged risks calculated from the male and female mice. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARL) Carcinogenicity and mutagenicity studies strongly implicate ethylene dibromide. Although dose studies in laboratory animals are limited, the carcinogenicity data seem to require that water contaminated with ethylene dibromide not be consumed. This would include acute spills, 7- day, or chronic situations. Furthermore, the data concerning acute and lo.

Toxicity of Selected Drinking Water Contaminants 101 subacute noncarcinogenic actions were insufficient to permit estimates of SNARL's for 24-hr and 7-day exposure limits. Dichlorodifluoromethane (Cl2CF2) Dichlorodifluoromethane (Freon 12) is a relatively inert, nonflammable liquid that is used mainly as a refrigerant. Its molecular weight is 120.92, its vapor pressure is 6.43 atm at 25°C, and it is quite insoluble in water (5.7 ml/100 ml at 26°C). METABOLISM Apparently there are no data regarding the metabolic fate of Freon 12 following oral exposure. The fluorocarbon compounds are lipid-soluble and thus are generally well absorbed through the lung. Absorption after ingestion is 35 to 48 times lower than after inhalation (Charlesworth, 1975~. The fluorocarbons are particularly stable compounds that do not appear to be metabolized (Blake and Mergner, 1974~. Inhalation experiments in laboratory animals have demonstrated that the half-life of fluorocarbons is very short (minutes) (Clayton, 1967) and that they are eliminated via the lung. Adir et al. ( 1975) reported that complete recovery of an inhaled dose in dogs occurs within 40 to 60 min. HEALTH ASPECTS Observations in Humans The threshold limit value (TLV) for inhaled dichlorodifluoromethane recommended by the American Conference of Governmental Industrial Hygienists (ACGIH) is 1?000 ppm for 8 hr (American Conference of Governmental Industrial Hygienists, 19751. This is approximately 2057O by volume. There are no data on the oral route of administration in humans. Inhalation studies in humans (Adir et al., 1975; Azar et al., 1972) indicated good tolerance to the TLV level (1,000 ppm, 25 mg/min). Even a 10,000-ppm exposure resulted only in a small (7S'o) reduction in the standardized psychometer test score indicating that this high dose given for 2.5 hr would not pose a serious threat to an individual's health. At 1,000 ppm, the Freon 12 level in venous blood was 1.2 ~g/ml. This is much less than the 22.8 ~g/~1 that is necessary to sensitize the dog's heart (Azar et al., 19739. Acute human inhalation exposures indicate that concentrations in

102 DRINKING WATER AND HEALTH excess of 50,000 ppm (5%) should be avoided and that 100,000 ppm (10%) produces unconsciousness (Largent and Largent, 1955~. Observations in Other Species Acute Effects The inhalation LDso in mice during a 3-fur exposure was 62% by volume, or 3,348 mg/liter (Shugaev, 1963~. The LD~oo was 6637O by volume, or 3,564 mg/liter. Shugaev observed the earliest symptoms (motor stimulation, placidity, rapid respiration, immobility) at 3037O by volume, or 1,620 mg/liter. At higher doses the mice quickly developed tremors and were narcotized. After acute exposure by inhalation to a 13.5% concentration for 30 s, the myocardium in unanesthetized dogs was sensitized to subsequent injection of epinephrine (Azar, 19711. In contrast, a 2.5% concentration that was inhaled 6 hr/day for 5 days resulted in no cardiac sensitization in dogs. When rats were intubated with 1,000 mg/kg (maximum feasible dose), Freon 12 produced no deaths (Clayton, 19671. Symptomatology was not described. Chronic Elects Sayers et al. (1930) conducted inhalation experiments in dogs, monkeys, aIld guinea pigs. They exposed the animals 6 to 8 hr/day 5 to 6 days/week, for up to 12 weeks. They investigated overt symptoms, growth, complete blood count, mortality, and gross patholo- gy. Although the reliability of the data must be questioned because of the very small number of animals that were tested, the authors concluded that 20% vapor was tolerated by these species with only transient symptomatology and no accumulative or permanent deleterious effects. Inhalation studies with dichlorodiRuoromethane were conducted in mice, rats, and dogs by Smith and Case (19731. Dosages were 970 mg/kg/day for 23 months (mice), 164 mg/kg/day for 93 days (rats), 700 mg/kg/day for 93 days (dogs), 560 mg/kg for 90 days (dogs), and 2,240 mg/kg/day for 1 year (dogs). There were no changes in electrocardio- grams of the dogs, whose high doses were approximately 200 times greater than those normally given in clinical use. No microscopic changes in tissue were found in any of the species. Blood chemistry and hematology and urinalyses were all normal. Clayton (1967) gave dichlorodifluoromethane to male and female rats and dogs in oral dose levels of 160 to 379 mg/kg (rats) and 84 to 85 mg/kg (dogs, for 90 days. He observed no effects on growth, hematolo- gy, behavior, or on tissue that had been examined both grossly and

Toxicity of Selected Drinking Water Contaminants 103 microscopically. Other studies, in which Clayton gave male rats 430 mg/kg for 10 days, also produced no ejects on tissue or toxicity. Dichlorodifluoromethane was administered by gavage to rats at 15 or 150 mg/kg/day for 2 years and in food to beagle dogs at 8 or 80 mg/kg/day for 2 years (Sherman, 1974~. Except for a slight adverse effect on growth of the rats that had received the high doses, no clinical, biochemical, urinary, hematological, or histopathological changes were found in either species. No evidence of a carcinogenic response was observed. Mutagenicity In chronic exposure studies, male and female rats were given Freon 12 in doses of 15 or 150 mg/kg by intubation. The rats were then bred and evaluated for fertility, corpora lutea, implantation sites, resorption sites, and number of live fetuses per litter. No dominant lethality was found at either dose level (Sherman, 1974~. Carcinogenicity For 2 years, groups of 50 male and 50 female albino rats were given Freon 12 orally by intubation in doses of approximately 15 or 150 mg/kg daily. Control groups of 50 rats of each sex received the vehicle alone by intubation. Indices studied included growth, clinical signs, mortality, hematology, urine, blood biochemistry, organ weight, and tissue pathology. No evidence of a carcinogenic effect was found (Sherman, 1974~. Teratogenicity In studies of inseminated Wistar albino rats and albino rabbits, Paulet et al. (1974) administered a mixture of 90~o Freon 12 (dichlorodifluoromethane) and 10% Freon 11 by inhalation for 2 hr/day. Rats were exposed on days 4 to 16 of gestation, and rabbits on days 5 to 20. The mixture was administered in a 2037O concentration (200,000 ppm). No indications of any embryotoxic, fetotoxic, or teratogenic changes were found when dams were sacrificed and fetuses removed on day 20 of gestation (rats) or day 30 of gestation (rabbits). No oral teratology data were found. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARLJ Twenty-Four-Hour Exposure The above acute toxicity data character- ize a compound of relatively low toxicity. For example, an oral single dose of 1,0()0 mg/kg in rats resulted in no mortality; inhalation in mice at 1,620 mg/liter of air produced pharmacologic symptoms but no

104 DRINKING WATER AND HEALTH mortality; and inhalation by dogs of 135-mg/liter of air resulted in no cardiac sensitization even after 6 hr/day for 5 days. The acute side effect of most concern is cardiac sensitization. The oral rat data indicate that a cardiac toxic dose may exceed the maximum (limited) water solubility (307.8 mg/liter). An uncertainty factor of 100 was used. Assuming 100% exposure from water during this period and a 2 liters/day intake by a 70-kg human: 1, mg/kg x 70 kg 100 x 2 liters 350 mg/liter. Seven-Day Exposure The 10-day experiment of Paulet-et al. (1974) in male rats (cited above) showed that an oral dose of 430 mg/kg/day was free of toxicity. This would translate, by direct extrapolation, into 30,100 mg/kg/day for a 70-kg human, or 15,050 mg/liter of water, assuming a consumption of 2 liters/day. Even a 100-fold safety factor would allow for 150 mg/liter in water. Based on the 10-day rat study and using the same assumptions as above: 430 mgJkg x 70 kg 100 x 2 liters = ISO mg/liter. Chronic Exposure Two-year studies in rats and dogs by the Haskell Laboratory indicate that the highest dose levels used (150 mg/kg/day in rats and 80 mg/kg/day in dogs) were free of any clinical, biochemical, hematological, or histopathological effects (including tumors). An uncertainty factor of 100 was used. A 20% intake from water and consumption of 2 liters/day of water by a 70-kg human was assumed: 80 mg/kg x 70ikg x 0 2 = 5.6 mg/liter. These data demonstrate that Freon 12 has remarkably low chronic toxicity even at high doses. Its very short half-life and rapid excretion may be a major factor in this low toxicity. 1,2-Dichloroethane (CICH2'H2CI) This compound was evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 7231. The following material, which became available after that 1977 publication, updates and, in some instances, reevaluates the information in the earlier report. Also included are some references that were not assessed in the original report.

Toxicity of Selected Drinking Water Contaminants 105 METABOLISM The pharmacokinetics of 1,2-dichloroethane (DCE) has been the subject of a limited number of investigations. Morgan et al. (1970), in a study of the uptake and elimination of a series of chlorinated hydrocarbons in humans, found DCE: vapor to be rapidly absorbed but rather slowly exhaled in relation to other compounds in the series. Morgan and his colleagues (1970, 1972) attributed rapid absorption and retention of the chlorinated hydrocarbons largely to their lipid solubility. Consequently, it appears that DCE has a potential for accumulation in the body. Apparently no data have been published regarding pharmacokinetics of ingested DCE in humans or in laboratory animals. The only in-vivo study of metabolism of DCE was conducted in mice by Yllner (19714. He found that DCE was metabolized through a 2-chloroethanol intermedi- ate to chloroacetic acid and related conjugation products. The propor- tion of DCE that was metabolized, as opposed to the DCE that was exhaled unchanged, varied inversely with dose. In in-vitro studies Van Dyke and Wineman (1971) confirmed that DCE can be dechlorinated by a rat liver microsomal system. Despite reports that liver alcohol dehydrogenase in the rat (Johnson, 1967) and human (Blair and Vallee, 1966) can catalyze oxidation of 2-chloroethanol to chloroacetaldehyde, the nature of DCE metabolism and its role in DCE toxicity remain largely speculative. HEALTH ASPECTS Observations in Humans Acute Effects A number of reports of human fatalities resulting from ingestion of DCE have been reviewed by the National Institute for Occupational Safety and Health (1976c). Estimated quantities of the chemical consumed range from "a sip" to as much as 100 ml. Quantities from 20 to 50 ml were commonly reported to be lethal. These figures are compatible with the established oral LD50 in rats, which is 0.77 ml/kg (Smyth et al., 19691. Published case histories of acute human exposures to DCE reveal that symptoms and signs of poisoning are quite similar, whether the DCE was inhaled or ingested. Nausea and vomiting are experienced by many persons who have been subjected to large quantities of the chemical by either route of exposure. Another important, recurring clinical finding in such patients is the presence of hemorrhagic lesions of most organs. An explanation for the finding is offered by Martin et al. (1969), who

106 DRINKING WATER AND HEALTH reported depletion of a number of clotting factors and thrombocytopenia in a DCE-poisoned patient. The hypocoagulability is attributed to disseminated intravascular coagulation. Schoenborn et al. (1970) were not successful in preventing loss of clotting factors and subsequent circulatory collapse by administering heparin to a patient 5.5 hr after DCE ingestion. However, Przezdziak and Bakula (1975) reported the successful use of heparin in the treatment of a man exhibiting both consumption coagulopathy and hepatic coma. Severe hepatotoxicity may augment hypocoagulability. Apparently, this occurred in the case of a 14-year-old who died from ingestion of approximately 15 ml of DCE (Yodaiken and Babcock, 19731. Chronic Effects Accounts of chronic exposure of humans to low levels of DCE via inhalation and/or skin contact in occupational settings indicate that toxic manifestations may resemble those in persons who have been poisoned by acute exposures to high concentrations of DCE. The recommended U.S. standard for occupational exposure to DCE is 5 ppm. This standard is based largely upon reports of adverse effects in humans in European industrial settings (National Institute for Occupa- tional Safety and Health, 1976c). Among other neurological abnormali- ties, these ejects include nausea and vomiting, changes in appetite and mood, depression, gastric pain and upset, and liver and kidney dysfunction. Workers experiencing such difficulties were exposed for various lengths of time to DCE vapor at levels ranging from as little as 10 ppm to as much as 200 ppm. Several of these accounts, which were reviewed by National Institute for Occupational Safety and Health (1976c), apparently suffer from inadequate or inappropriately designed atmospheric sampling. However, there are a sufficient number of reports of adverse health effects to suggest that prolonged inhalation of 10 to 15 ppm DCE may be harmful. A single study, which was conducted under controlled laboratory conditions with humans, suggests that inhalation of as little as 1.5 to 3.0 ppm DCE for 30 s to 1 min can produce transient vasoconstriction and altered breathing patterns (Borisova, 1957~. Observations in Other Species Acute Effects Studies of the acute toxicity of DCE in laboratory animals have for the most part involved inhalation exposures. The acute oral LD50 of 1,2-dichloroethane has been established at 0.77 (0.67-0.89) ml/kg in rats (Smyth e' al., 1969~. LC50 values for vapor inhalation are 12,000 ppm in 0.53 hr. 3,000 ppm in 2.75 hr. and 1,000 ppm in 7.20 hr in

Toxicity of Selected Drinking Water Contaminants 107 rats (Spencer et al., 1951~. In one study with rabbits, the LD50 for skin penetration was determined to be 3.89 (3.4~54.46) ml/kg. Signs of acute poisoning in animals are quite similar to those in humans. The toxic effects of single acute exposures to 1,2-dichloroethane were central nervous system depression, lung irritation, and injury to the liver, kidneys, and adrenals (Gohlke and Schmidt, 1972~. A single study by Plaa and Larson (1965) concerns acute exposure of laboratory animals to DCE by a route of administration other than inhalation. These investigators gave the compound to mice by intraperitoneal injection and found modest renal injury only with a near-lethal dose (i.e., 0.4 ml/kg). The renal injury was indicated by alteration of only one of three variables tested. Chronic Effects Based upon the chronic exposure studies of Heppel et al. (1946), Hofman et al. (1971), and Spencer et al. (1951), the highest no- e~ect vapor level in a variety of species would be approximately 100 ppm. Even at this level, increase in liver weight and diminished body weight gain have been noted in certain species (Hofman et al., 1971; Spencer et al., 19519. No mention was made in these studies of measurement of pharmacokinetic parameters. Results of two long-term DCE feeding studies have recently been published. In the first, Alumot et al. ( 1 976a) fed male and female Leghorn chickens diets containing 250 and 500 ppm DCE for up to 2 years. They observed erects during this time on growth, fertility, spermatogenesis, or clinical chemistry. There was some reduction in egg weight and egg production. Therefore, Alumot et al. (1976a) proposed 100 ppm as a maximum tolerance limit for contamination of poultry feed and 5 mg/kg body weight as an acceptable daily intake for chickens. In the second chronic, low-level feeding study with aliphatic halocar- bons, Alumot et al. (1976b) fed carbon tetrachloride or DCE to male and female rats for up to 2 years. Results of their fertility and reproduction studies are discussed in the section on teratogenicity. The minimum toxic dietary levels of carbon tetrachloride and DCE in the feed, as revealed by increases in hepatic total lipids and triglycerides after 5- to 7-week feeding periods, were 275 and 1,600 ppm, respectively. No significant alterations in body weight gain or in clinical chemistry indices were observed in male or female rats after 2 years of consumption of carbon tetrachloride (80 and 200 ppm) or DCE (250 and 500 ppm). Based on these established "no-effect levels," Alumot et al. (1976b) proposed that an acceptable daily intake of DCE for animals was 25 ml/kg body weight and a tolerance level for feed contamination was 250 ppm. These investigators also proposed a tolerance of 10 ppm for DCE in human

108 DRINKING WATER AND HEALTH food, considering that a safety factor of at least 100 should be applied to data from experiments in animals. Mutagenicity 1,2-Dichloroethane has been found to be a weak mutagen in several nonmammalian and in-vitro test systems. Brem et al. (1974) reported a mutagenic effect of 1,2-dichloroethane in Salmonella typhimurium (TA 1530, TA 1535 and TA 1538) and DNA polymerase- deficient Escherichia colt. However, it was the weakest of the series of haloalkanes tested. In other studies, DCE has been found to be weakly mutagenic in Drosophila (Shakarnis, 1969, 1970) and in two strains of S. typhimurium (McCann et al., 1975b; Rannug and Ramel, 1977~. Noting that DCE and 1,2-dibromoethane are mutagenic in Drosophila, Vogel and Chandler (1974) suggested that each compound, upon loss of a halogen atom, might act as a bifunctional alkylating agent capable of introducing cross-links into biological molecules such as DNA. Although DCE itself is apparently a weak mutagen in certain nonmammalian test systems, certain of its metabolites would appear to be more hazardous. Two possible human metabolites, 2-chloroethanol and chloroacetaldehyde, are reported to be relatively potent mutagens in certain bacterial test systems, but not in others (McCann et al., 1975a; Rannug et al., 1976; Voogd and Van der Vet, 1969~. McCann et al. (1975a) and Rannug et al. (1976) demonstrated that chloroacetaldehyde is many times more potent than 2-chloroethanol, while chloroacetate is inactive. McCann et al. (1975a) could enhance the mutagenicity of chloroethanol by incubating it with rat liver microsomes, but could not similarly activate DCE. However, Rannug and Ramel (1977) reported a fivefold increase in mutagenicity of DCE, when a 9,000-g, NADPH- deficient rat liver fraction is added to DCE in their S. typhimurium screening system. Carcinogenicity There is recently published evidence that high doses of DCE may be carcinogenic in certain species of animals. Theiss et al. (1977), in their evaluation of the ability of DCE to produce pulmonary adenomas in Strain A male mice, reported that repeated intraperitoneal injections of 100 mg/kg body weight of the chemical elicited an increased tumor incidence at 24 weeks, although this incidence was not significantly (statistically) different from controls. At the National Cancer Institute (1978), female and male mice and rats were given large oral doses of DCE (5~300 mg/kg) daily (5 times/week) for 78 weeks. Mammary tumors were observed in DCE-treated female rats, while stomach tumors were seen in some DCE-exposed male rats. Male mice showed a dose-dependent increase over controls in incidence of

Toxicity of Selected Drinking Water Contaminants 109 hepatocellular carcinoma, and both male and female mice that had been subjected to DCE exhibited an increase in lung tumors. McCann et al. (1975a) speculated that vinyl chloride, DCE, and 2-chloroethanol all may prove to be carcinogenic, since each may be metabolized via a similar pathway. They mentioned chloroacetaldehyde as a potential carcinogenic metabolite that is common to each of these three com- pounds. Teratogenicity Although DCE and its metabolites have been demon- strated in certain nonmammalian test systems to be mutagenic, no one has apparently found DCE to be teratogenic in vivo. Heppel et al. (1946) failed to report abnormalities in the offspring of guinea pigs and rats that had been subjected to long-term DCE inhalation exposures during pregnancy. Alumot et al. fed chickens (1976a) and rats (1976b) diets containing 250 and 500 ppm DCE for up to 2 years. In the chickens, there was no effect on sperm count or motility and no influence on fertility of either sex. No alteration of fertility or reproduction was seen in the rats. The average size, weight, and mortality of offspring were unaffected by DCE. Carcinogenic Risk Estimate In a recent study by the National Cancer Institute (1977), rats and mice were Savaged with doses of 1,2-dichloro- ethane ranging from 47 to 299 mg/kg/day for as long as 78 weeks. There was a dose-related incidence of squamous cell carcinomas of the forestomach in the male rats, but not in the females. The mice also exhibited dose-related cancer at several sites including mammary glands and endometrium, and there were alveolar/bronchiolar adenomas in both sexes. Each set of dose-response data was used to make statistical estimates of both the lifetime risk and an upper 95% confidence bound on the lifetime human risks. These estimates have been corrected for species conversion on the basis of dose/surface area. The risk estimates are expressed as a probability of cancer after a lifetime consumption of 1 liter/day of water containing 1 ~g/liter of the compound. For example, a risk of 1 x 10-6 implies a lifetime probability of 2 x 10-5 of cancer if 2 liters/day were consumed and the concentration of the carcinogen was 10 ~g/liter. At a concentration of lO ~g/liter during a lifetime exposure, this compound would be expected to produce one excess case of cancer for every 50,0()0 persons that are exposed. If the population of the United States is taken to be 220 million people, this translates into 4,400 excess lifetime deaths from cancer, or 62.8 per year. At a DCE concentration of l ~g/liter, the estimated lifetime risk for humans is 3.7 x 1O-7. The upper 95% confidence estimate is 7.0 x 10-7.

110 DRINKING WATER AND HEALTH Both of these estimates are the averaged risks, which have been calculated from the male and female rats and mice. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level fSNARL) Twenty-Four-Hour Exposure/Seven-Day Exposure DCE is acutely toxic to both humans and laboratory animals. Although its ingestion has proved fatal to a number of individuals, there is unfortunately little information pertaining to quantities of DCE consumed in nonfatal cases. Nor are there data on concentrations of the parent compound or its metabolites in the bodily fluids or organs of the poisoned victims. Thus, there is little basis for making a reasonable judgment as to the minimum toxic oral dose of DCE in humans. Little is known about potential toxic effects of exposure of animals to DCE by any route other than inhalation. While inhalation studies are valuable in determining manifes- tations of exposure to DCE and setting vapor exposure limits, inhalation toxicity studies include few data on uptake and distribution of the compound. Thus, it is very difficult to relate toxic effects at a given vapor concentration of DCE to injury that might result from ingestion of a given quantity of the chemical. The metabolism of DCE and the role of metabolism in activation or inactivation of the compound as a toxicant are largely speculative. Therefore, the subcommittee concluded that there is insufficient information to serve as a basis for recommending a 24-hr or a 7-day suggested no-adverse-response level (SNARL). Definitive acute and subacute toxicity studies in animals should be conducted using several dissimilar animal species; employing a range of oral doses to characterize dose-response relationships and ascertain the maximum no-effect level for both single ingestions and multiple ingestions over a 24-hr and a 7-day period; examining a variety of test parameters which are valid indices of injury to known or suspected target organs/cells/biochemical systems (organs to be examined should include the heart, lung, kidney, adrenal, and liver); and investigating uptake, distribution, metabolism, and excretion of DCE and its major metabolites. Chronic Exposure Chronic exposure to DCE is potentially hazardous. A variety of injurious effects in humans have been attributed to prolonged, low-level exposure to the chemical in occupational settings (National Institute for Occupational Safety and Health, 1976c) Unfortu- nately, it is difficult to relate findings in such accounts of inhalation

Toxicity of Selected Drinking Water Contaminants 111 and/or dermal exposures to situations involving chronic oral intake of DCE. Although a number of inhalation studies utilizing several species of animals indicate that 100 ppm of DCE is essentially a "no-e~ect level" on chronic intermittent exposure, extrapolation to continuous ingestion is tenuous. When considered alone, data generated by the chronic feeding studies of Alumot et al. (1976a,b) appear to be the best and most applicable for calculation of a chronic SNARL. They reported "no-effect dietary levels', of 250 and 500 ppm for rats and chickens fed DCE- contaminated diets for 2 years. These investigations included evaluation of indices of growth, blood chemistry, fertility, and reproduction. These data should not serve as the only basis for setting a chronic SNARL since a number of recent reports suggest that DCE may be a mutagen and/or a carcinogen. Additional long-term oral ingestion studies employing several species of animals are needed to determine if DCE is a carcinogen and, if so, which organs are involved in different species; the nature of uptake, metabolism, and accumulation of DCE and its metabolites; and minimum times and doses of DCE that are required to induce tumors. Populations of workers that have been exposed occupationally to DCE and to 2-chloroethanol might also be screened, as are vinyl chloride workers, for cancer. Vinyl chloride, 2-chloroethanol, and DCE may each be carcinogenic, since they may share a common metabolic pathway. Additional investigations need to be conducted using other indices of toxicity. Attention might be concentrated on histopathology and on sensitive tests to detect cardiac, adrenal, pulmonary, renal, and hepatic . . nJury. Epichlorohydrin (H 2C-CHCH2CI) Epichlorohydrin (ECH) is an important industrial chemical that is used frequently as a solvent for resins, gums, cellulose, and paints and, especially, as a raw material for the manufacture of epoxy resins and synthetic glycerine (International Agency for Research on Cancer, 1976; National Institute for Occupational Safety and Health, 1976b; U.S. Environmental Protection Agency, 1977a). In agriculture, it is used as an insect fumigant (Anonymous, 1978~. ECH is synthesized commercially from allyl chloride, allyl alcohol, dichlorohydrin-glycerine, or propylene. However, the major route of synthesis is apparently high temperature chlorination of propylene to allyl chloride, which is followed by chlorohydrination with hypochlorous acid to a mixture of isomeric glycerol chlorohydrins. These are subsequently dehydrochlorinated with alkali to yield the technical product (Lichtenwalter and Riesser, 1964~. In

112 DRINKING WATER AND HEALTH 1975 the total U.S. production of ECH was about 250 million kg; 325 million kg was produced in 1978. In 1969, an estimated 58% of the ECH was used in the manufacture of synthetic glycerine and 42% was processed to refine ECH. Refined ECH is used in the manufacture of epoxy resins, surface active agents, pharmaceuticals, insecticides, agri- cultural chemicals, textile chemicals, coatings, adhesives, ion-exchange resins, solvents, plasticizers, glycidyl esters, ethinyl-ethylenic alcohol, and fatty acid derivatives. Most epoxy resins are synthesized by alkylating bisphenol A with ECH. The compound is also a proposed intermediate in the metabolism of allyl chloride (Van Duuren, 1977~. ECH is a colorless liquid at room temperature. It has a distinctive odor, which has been described as "ethereal," "chloroform-like," and "garlic-like." The boiling point of ECH is 116°C at 20 mm of mercury. The low latent heat of vaporization of ECH (9,060 cal/mol) contributes to its relatively high volatility. ECH is soluble in water, 6.4% w/w at 20°C (International Agency for Research on Cancer, 1976~. ECH is characterized by two potentially reactive sites: the epoxide ring and the chlorine atom. The presence of the highly strained three- membered ring makes ECH a relatively reactive compound. It is hydrolyzed slowly at room temperature, but hydrolysis is greatly accelerated by heat or traces of acid or base. An important feature of the chemistry of ECH is its ability to form compounds containing two functional groups. ECH reacts in viva to form covalent bonds with alcohols, amines, thiols, and other nucleophilic biochemical constituents of the cell. The epoxide ring opens to form a new, stable, covalent carbon-hetero atom bond: HC-C CC1 + RXH = H R R H O X H X O H 1 1 1 1 HC- C CCI + HC- C CC1 H H H H H H H H H Reactions of epichlorohydrin with cellular nucleophiles (where X= an electronegative element such as 0, N. or S. and R = an aLkyl, aryl, or other organic group). The initial reaction products may undergo a second nucleophilic reaction to form stable, covalent cross-linking bonds between two molecules, either by direct displacement of the chlorine atom or through the formation of a nonstable, short-lived cyclic intermediate (Alexander et al., 1952~. These cross-linking bonds may have a high degree of chemical stability, such as that found typically in epoxy resins. The half-life of ECH in water at pH 7 is about 36.3 hr (National

Toxicity of Selected Drinking Water Contaminants 113 Institute for Occupational Safety and Health, 1976b). Presumably, ECH is hydrolyzed initially to form 1-chloro-2,3-propanediol (cx-chlorohydrin). Neither ECH nor a-chlorohydrin is a strong nucleophile and could be expected to react further (Jones et al., 1969~. Reaction of ECH at the less reactive chlorine rather than the epoxide is also known to occur (International Agency for Research on Cancer, 1976~. ECH has a pronounced effect on the organoleptic properties of water. Not only does it impart a specific odor, but it can also be irritating to the mouth. The threshold for odor perception of the compound is 0.5 to 1.0 mg/liter (133 to 265 ppm), while the threshold for its irritant action is below this level at 0.1 mg/liter (26.5 ppm) (U.S. Environmental Protection Agency, 1977a). In light of the toxicity and cumulative elects of ECH, Fedyanina (1968) recommended 0.01 mg/liter as a maximum permissible concentration in reservoir water. METABOLISM There has been little research on the metabolism of ECH (National Institute for Occupational Safety and Health, 1976b). Apparently, a- chlorohydrin is a hydrolysis product of ECH (Jones et al., 1969; National Institute for Occupational Safety and Health, 1976b). Jones et al. (1969) demonstrated the in-vivo formation of a-chlorohydrin by hydrolysis of ECH, since Wistar rats dosed orally or intraperitoneally with a-chlorohydrin or ECH yield some of the same urinary metabolttes, i.e., 2,3-dihydroxypropyl-S-cysteine and its N-acetate. Since ECH is a strong electrophile that is capable of reacting with cellular nucleophiles, it is probable that some ECH metabolites are also covalently bound to various tissue macromolecules. While there are few published reports on the metabolism of ECH, the fate of its hydrolysis product, ~x-chlorohydrin, has been investigated in some detail. Jones (1975) showed that c~-chlorohydrin is conjugated with glutathione in the rat and then excreted in the urine as S-~2,3-dihydroxy- propyl~cysteine and the corresponding mercapturic acid, N-acetyl-S-~2,3- dihydroxypropyl~cysteine. On the basis of the similar antifertility properties of ~x-chlorohydrin and glycidol (Jackson et al., 1970) and of the results of in-vivo and in-vitro metabolism experiments (Jones, 1975), a-chlorohydrin is thought to be converted initially in vivo to the reactive intermediate glycidol (2,3-epoxypropanol). Glycidol is metabolized by conjugation with glutathione. In a subsequer~t report, Jones and Murcott (1976) showed that a-chlorohydrin is also oxidatively metabolized to ,B- chlorolactic acid and then to oxalic acid by the rat. In the 24 hr after administration of 36Cl-labeled a-chlorohydrin, 16970 of the radioactivity

114 DRINKING WATER AND HEALTH was also eliminated in the urine as 36C1-. Presumably, this dechlorination occurs when a-chlorohydrin is converted to glycidol. HEALTH ASPECTS Observations in Humans Several reports of industrial exposure to ECH have been documented (National Institute for Occupational Safety and Health, 1976b; U.S. Environmental Protection Agency, 1977a). Eye and throat irritation, nausea, vomiting, headache, and dyspnea were the initial effects after a worker inhaled an unknown amount of ECH. An enlarged liver was reported within 2 days of exposure, and bronchitis and liver damage were present 2 years later. ECH has also been associated with persistent burns following dermal exposure of workers (National Institute for Occupational Safety and Health, 1976b; U.S. Environmental Protection Agency, 1977a). Several reports (Kucerova and Polivkova, 1976; Kucerova et al., 1976, 1977; Sram et al., 1976) have indicated that ECH has a dose-related mutagenic effect on human lymphocytes. Kilian et al. (1978) reported that an unknown substance (presumably a metabolite), which induced mutations in Salmonella typhimurium, was present in the urine of individuals who were accidently overexposed to ECH. The hygienic standard for ECH [8-hr time-weighted average (TWA) in mg/m3] is 5 in East Germany and 18 in West Germany. In the USSR, the maximum acceptable ceiling concentration is 1 mg/m3 (Winnell, 19751. The United States recently lowered its standards to a TWA of not greater than 2 mg/m3 for a 40-hr workweek, with a ceiling concentration of 19 mg/m3 (National Institute for Occupational Safety and Health, 1976b). Previously, the U.S. standard was a TWA of 19 mg/m3 (Winnell, 19751. The National Institute for Occupational Safety and Health (1976b) estimates that 50~000 employees may currently be exposed to ECH in the United States. Observations in Other Species Acute Elects Reports of acute oral LD50 values for ECH range between 90 and 240 mg/kg for rats, mice, and guinea pigs (Christensen et al., 1976; Lawrence et al., 1972; National Institute for Occupational Safety and Health, 1976b). The acute LD50in mice, rats, guinea pigs, and rabbits after intraperitoneal administration was between 118 and 165 mg/kg. The LD50 in rats and mice following intravenous injection was 154 and 178 mg/kg, respectively (Patty, 1963~. The subcutaneous LD50

Toxicity of Selected Drinking Water Contaminants 115 of ECH in the mouse was 720 mg/kg and the dermal toxicity in the rabbit was 1,300 mg/kg (Christensen et al., 1976~. The acute toxicity of the ECH metabolite a-chlorohydrin was similar to that of ECH. Its oral LD50 for the rat and mouse was 150 and 160 mg/kg, respectively (Christensen et al., 1976~. Because of the volatility of ECH and the many opportunities for frequent industrial exposure to the chemical, considerable work has also been done on its inhalation toxicity. Inhalation exposure or mice to 16,600 and 8,300 ppm ECH for 30 min produced 100% mortality, whereas no deaths were observed in mice that had been exposed to 2,370 ppm for 1 hr (National Institute for Occupational Safety and Health, 1 976b). Exposure of rats to 250 ppm ECH for 4 hr killed two to four out of six rats (Carpenter et al., 1949~. The lowest published lethal concentration (LCLo) for a 30-min inhalation exposure of mice was 7,400 ppm (Christensen et al., 1976~. Acute exposure to ECH can cause central nervous system depression, irritation of the respiratory tract, profuse nasal discharge, weight loss, leukocytosis, and kidney damage (International Agency for Research on Cancer, 1976; National Institute for Occupational Safety and Health, 1 976b). Death generally results from depression of the respiratory center. Nephrotoxicity is a cumulative effect of ECH poisoning, and renal insufficiency occurs within 2~48 hr in approximately 80% of the rats that have been given 125 mg/kg of the compound. ECH produces extreme irritation when tested intradermally, dermally, or intraocularly in rabbits (Lawrence et al., 19721. Subacute Effects Gage (1959) exposed five groups of eight albino Wistar rats (four males and four females) to ECH vapor at 9, 17, 24, 56, and 120 ppm for 6 hr/day, 5 days/week, for a total of 11 to 19 days. To facilitate comparison of the data the National Institute for Occupational Safety and Health (1976b) calculated the total amount of inhaled ECH from the concentration of ECH in the inhaled air and the duration of exposure. This calculation does not include an estimated percent absorption of the inhaled ECH. All rats inhaling ECH at a concentration of 120 ppm for 11 exposures (total amount inhaled approximately 1,132 mg/kg) experienced labored breathing, profuse nasal discharge, a marked loss of weight, leukocytosis, and an elevation in urinary protein excretion. Microscopic examination of the kidney and liver showed leukocyte infiltration, atrophy, and necrosis, while the lungs were congested and edematous. Rats that were exposed 8 times to ECH at 56 ppm (total amount inhaled approximately 417 mg/kg) had mild nasal irritation and

116 DRINKING WATER AND HEALTH some abnormal lung histopathology. At the lowest concentration tested, 17 ppm for 19 days, which is equivalent to an exposure of 227 mg/kg, there were no apparent effects. Lawrence et al. (1972) gave intraperitoneal injections of 1 1.2 and 22.4 mgJkg/day to groups of 12 Sprague-Dawley rats for 30 days. At the end of the exposure period hemoglobin values were decreased significantly at the high dose, whereas there was a significant increase at the lower dose. Leukocytosis was also observed in the group given the high dose. The kidney: body weight and brain: body weight ratios increased sig- nificantly at both ECH doses. Microscopic examination did not reveal any abnormalities in any organs except lungs where lesions were evident in all ECH-exposed groups. In a second experiment male Sprague-Dawley rats received 0, 11.2, 22.4, and 56.0 mg/kg, injected intraperitoneally on Mondays, Wednes- days, and Fridays for 12 weeks (Lawrence et al., 19721. At the highest dose level there was a significant reduction in body weight, hemoglobin, and hematocrit values. This was accompanied with an increase in the weights of the heart, kidney, and liver, and a decrease in brain weight. The only adverse effect in animals receiving 11.2 mg/kg was a reduced hematocrit. Grigorowa et al. (1974) exposed two groups of 60 male albino rats to ECH at 15.6 or 156 ppm by inhalation for 4 hr/day for 8 days. At the highest exposure level (estimated at 89 mg/kg/day), there was a significant decrease in body weight. They observed no significant adverse ejects after exposures to 15.6 ppm (8.9 mg/kg total daily intake). Some of the animals were exposed at 35°C for 45 min on each of the 8 days, but the investigators concluded that heat stress had no eject on the toxicity of ECH. The toxicological properties of ECH were studied by Fedyanina (1968), who administered daily oral doses of 20 mg/kg to groups of 18 albino rats over 2 months. Body weights of control and exposed animals were the same throughout the treatment. Rats that were exposed to ECH showed decreased numbers of reticulocytes and leukocytes and an altered differential leukocyte count. Blood glutathione levels were increased and were primarily in the oxidized for. Gage (1959) also exposed two groups of two rabbits to ECH vapor at concentrations of 16 and 35 ppm for approximately 6 hr/day, 5 days/week. At the higher levels, the animals were exposed 20 days (approximately 439 mg/kg total amount inhaled). Inhalation exposure to ECH at 35 ppm produced nasal irritation, but postmortem examination failed to show any abnormalities. The second group of rabbits was exposed for two periods at 16 ppm (total amount inhaled approximately

Toxicity of Selected Drinking Water Contaminants 117 20 mg/kg), but this exposure was reduced to 9 ppm and continued for 20 more days (approximately 113 mg/kg total amount inhaled). Thus, the second group of rabbits is estimated to have inhaled a total dose of 133 mg/kg. Nasal irritation was observed during the period in which the animals were exposed to 16 ppm; however, following exposure for a longer period to 9 ppm, no adverse effects were detected. Fedyanina (1968) also administered daily oral doses of 80 mg/kg ECH to 13 rabbits over 2 months. Two rabbits died toward the end of the experiment, and the weight of the experimental rabbits was significantly below that of the control animals. Exposed animals exhibited a reduced leukocyte count, an altered differential count, decreased blood gluta- thione levels, and various changes in blood chemistry. Chronic Effects Kremneva and Tolgskaya (1961) exposed a group of 10 rats for 3 hr daily to ECH vapor at 5.2 to 15.6 ppm for up to 6.5 months. This exposure is equivalent to a daily ECH intake of between 2.2 and 6.7 mg/kg/day. No deaths or signs of intoxication were observed in the exposed animals, but the gain in body weight was less than that of controls. Exposed animals also showed evidence of hyperexcitability. The investigators observed no significant variations in the composition of the peripheral blood, but histopathological examination of the tissues showed some abnormal lung pathology. Fomin (1966) exposed three groups of 15 male albino rats to ECH vapor for 24 hr/day for 98 days at concentrations of 0, 0.05, O.S, and 5.2 ppm (equivalent to 17.2, 172, and 1,722 mg/kg total inhaled dose). Rats exposed to the highest dosage level, 5.2 ppm (17.6 mg/kg/day), showed reduced weight gains and prolonged latent periods in the motor defense reaction. The number of leukocytes in the blood was increased, and there was an elevated concentration of coproporphyrin in the urine. Gross and microscopic examination disclosed emphysema, bronchopneumonia, edematous areas, and an altered histopathology of the blood vessels in the lungs. Cloudy swelling in the epithelium of the convoluted tubules of the kidney, interstitial hemorrhage in the heart, and severe lesions in the brain were also evident. Animals inhaling ECH at the intermediate concentration of 0.52 ppm (1.76 mg/kg/day) showed leukocytosis and a decrease in blood nucleic acids, while animals exposed to the lowest concentration of ECH, 0.05 ppm (0. 176 mg/kg/day), produced no abnormalities Fedyanina (1968) administered ECH orally to groups of 1S albino rats at doses of 0, 0.005, 0.05~ and 5 mg/kg/day for 6 months. The author then studied the effects of ECH exposure on the conditioned reflexes of the animals using an "accelerated" variant of the motor-alimentary

118 DRINKING WATER AND H"LTH method. Dose-dependent weakening of stimulation and inhibition processes were seen in animals receiving 5 and 0.05 mg/kg/day. Doses of 0.005 mg/kg/day only diminished the strength of reflexes to light stimulus in the ECH-exposed animals, suggesting that this dose is liminal. However, supplementary studies on the effect of doses of 0.005 and 0.0005 mg/kg/day revealed a decreased synthesis of hippuric acid after loading with sodium benzoate in animals that were treated with 0.005 mg/kg/clay. The authors concluded that a daily dose of 0.0005 mg/kg ECH had no adverse effect on the functions that were investigated. The same investigator (Fedyanina, 1968) also conducted longer studies in rabbits in which groups of 11 rabbits received daily oral doses of 0, 0.005, 0.05, and 0.5 mg/kg of ECH for 6 months. The most pronounced changes in the treated rabbits were a reduced reticulocyte count, reduced blood glutathione levels, and increased albumin in the plasma of animals that had been dosed with 0.5 mg/kg/day. Levels of oxidized and reduced glutathione in blood and ascorbic acid in the kidney and liver were reduced in animals that had received 0.05 mg/kg/day ECH. Rabbits that had been treated with the lowest concentration showed no significant changes. Reproductive Studies ECH-induced sterility has been reported in male animals. Hahn (1970) gave oral doses of 15 mg/kg of ECH for 12 days (total dose of 180 mg/kg) to adult male rats of demonstrated fertility. Within 1 week males became infertile, but fertility was restored after 1 week of daily dosing when ECH was terminated. Histological examination of the testes, epididymis, prostate, and seminal vesicles showed no difference between treated and control animals. Jones et al. (1969) investigated the antifertility effects of ECH on Wistar rats. A single oral or intraperitoneal dose of 50 mg/kg produced infertility effects resembling those produced by the ECH metabolite, a- chlorohydrin. Cooper et al. (1974) gave oral doses to adult male rats. One group, which received five daily doses of 20 mg/kg (100 mg/kg total), displayed loss of fertility during the first 2 weeks, but recovered completely by the third week. A second group, which received five daily 50 mg/kg doses (total dose 250 mg/kg) in the same manner, was rendered completely sterile throughout a 10-week period. The last group of rats, which received one 100 mg/kg oral dose, experienced sterility within 1 week. By the 12th week following this treatment, permanent sterility had probably occurred. -v -O. There is considerable information regarding the antifertility properties

Toxicity of Selected Drinking Water Contaminants 119 of the ECH metabolite a-chlorohydrin, which is an orally active, reversible male antifertility agent in the rat, hamster, guinea pig, ram, boar, and monkey (Brown-Woodman et al., 1974;.Cooper et al., 1974; Coppola, 1969; Dixit et al., 1974; Ericsson, 1970; Ericsson and Baker, 1970; Ericsson and Norland, 1970; Ericsson and Youngdale, 1970; Johnson and Pursel, 19731. However, in the rhesus monkey chronic administration of a-chlorohydrin also produces damage to bone marrow (Kirton et al., 1970~. The minimal effective daily oral dose of a- chlorohydrin in the male rat that is necessary to prevent fertilization in the female following mating was 2.5 mg/kg/day (Vickery et al., 1974~. cr-Chlorohydrin has two effects on the reproductive tract of the male rat. Five consecutive 10 mg/kg daily doses cause immediate and reversible infertility by inhibiting sperm glycolysis. Conversion of a- chlorohydrin to its 1-phosphate ester within the spermatozoa and inhibition of glycolysis by this ester through inactivation of phospho- glyceraldehyde dehydrogenase has been suggested as the cause of antifertility action of a-chlorohydrin (Mohri et al., 1975~. The low-dose effect may also involve the epoxide metabolite of a-chlorohydnn, glycidol (Jackson et al., 19701. A single high oral dose (100 mg/kg) of a-chlorohydrin in the rat produces epididymal lesions or spermatoceles (Cooper et al., 1974~. These lesions occlude the ductuli efferentes, blocking the passage of testicular sperm and producing prolonged or even permanent infertility. Jones and Murcott (1976) have proposed that this and observed renal toxicity are due to the in-vivo oxidation of cx-chlorohydrin to ,l]-chlorolac- tic acid and, ultimately, oxalic acid. Jackson and Robinson (1976) found that the R(-)isomer of a- chlorohydrin exhibited neither reversible antifertility activity nor pe~ma- nent sterilizing capacity, but the toxicity of the R(-)isomer was higher than that of the racemic mixture D~-a-chlorohydrin. Conversely, the S(+)isomer of a-chlorohydrin has almost twice the antifertility activity of the racemic mixture and a reduced toxicity (Jackson et al., 1977~. Mutagenicity ECH is considered a potent mutagen and its mutagenic properties have been the object of numerous investigations (Fishbein, 1976; National Institute for Occupational Safety and Health, 1976b; U.S. Environmental Protection Agency, 1977a.) One of the earliest studies on the mutagenicity of ECH was undertaken by Rapoport (1948) who reported that it induced 0.7% mutations in Drosophila melarlogaster. The mutagenicity of ECH was subsequently shown in Escherichia cold (Strauss and Okubo, 1960), in Klebsiella pneumoniae (Voogd, 1973), and in plants (Loveless, 19511. A 150 mg/kg intraperitoneal dose did not

120 DRINKING WATER AND HEALTH induce dominant lethal mutations in ICR/Ha Swiss mice (Epstein et al., 19724. Mukai and Haw~yluk (1973) found that ESCH induced a Enfold increase in revertants over controls in both E. cold and S. typhimurium. In studies aimed primarily at testing the mutagenicity of vinyl chloride and its metabolites, Elmore et al. (1976) examined the mutagenicity of ECH since it is a methylene homolog of the chloroxirane metabolite of vinyl chloride. Tests with S. typhimurium strain TA-100 (a base-pair substitu- tion mutant) showed that ECH was highly mutagenic. These and subsequent tests were conducted in the absence of a mammalian S-9 fraction. However, ECH was less active than its congener, the chloroxi- rane metabolite of vinyl chloride. In contrast, the chemical did not inhibit a recombination repair of a deficient strain of Bacillus subtilis (strain MC-1), whereas the chloroxirane compound was highly active. These workers suggested that ECH produced a different type of DNA lesion than did the chloroxirane compound. Similar studies by Sram et al. (1976) demonstrated high mutagenic activity of ECH in S. typhimuri- um tester strains G-46 and TA-100 at exposures of 1 to 50 mM/hr. Three hours after doses of 50 and 100 mg/kg, ECH was found to increase the frequency of back mutations using Salmonella strains G-46, TA-100, and TA-1950 in a host-mediated assay. ECH induced chromosome abnormalities in the bone marrow of mice when injected intraperitoneally in single doses ranging from 1 to 50 mg/kg or in five doses of 5 to 20 mg/kg, and when given orally in single doses ranging from 5 to 100 mg/kg or in five doses of 20 mg/kg (Sram et al., 1976~. The frequency of chromosome change was dose-dependent. ECH did not induce dominant lethal mutations in mice when injected intraperitoneally in doses ranging from 5 to 40 mg/kg or in five doses of 1 to 10 mg/kg, and when given orally in single doses ranging from 20 to 40 mg/kg or in five doses of 4 to 20 mg/kg. Human peripheral lymphocytes that were exposed to 10-~i to 10-4 M ECH in vitro for 24 hr produced chromosomal abnormalities (Kucerova and Polivkova, 1976; Kucerova et al., 1976; Sram et al., 1976), but ECH was 4 to 5 times less mutagenic than the polyfunctional alkylating agent triethylenephosphoramide (TEPA) when tested in this test system at similar concentrations. The chromosomal changes were dose-dependent, and the most common type of aberration produced by ECH were chromatic breaks, followed by chromosomal breaks. Chromatid ex- changes were rare and chromosomal exchanges extremely rare. A subsequent report by this group (Kucerova et al., 1977) summarizes a prospective cytogenic study on 35 workers who had been occupationally

Toxicity of Selected Drinking Water Contaminants 121 exposed to ECH. Blood samples were obtained from them after the first and second year of exposure and cultivated for 56 to 58 hr. The percentage of cells with chromosomal aberrations in these blood samples was 1.37 before exposure, 1.91 after the first year of exposure, and 2.69 after the second year of exposure. These differences were highly significant at P < 0.0001. Particularly frequent were chomatid and chromosomal breaks after exposure. Kilian et al. (1978) noted that an unknown substance (presumably a metabolite) that can induce mutations in S. typhimurium was present in the urine of individuals who were accidentally overexposed to ECH. Mutagenic activity was also detected in the urine in mice after oral administration of 200 to 400 mg/kg ECH. The reactive constitutent in urine is likely to be a-chlorohydrin, which is a product of hydrolysis of ECH (National Institute for Occupational Safety and Health, 1976b). a- Chlorohydrin is a highly reactive electrophile that is capable of forming conjugates with sulfhydryl compounds in viva (Jones et al., 1969), and, as is the case with ECH, also has potent male antifertility activity (Cooper et al., 1974; Dixit et al., 1974; Vickery et al., 1974~. Sram et al. (1976) estimated the genetic risk to humans from occupational exposure to ECH. The expected exposure to ECH for workers in the plastics industry was estimated to be approximately 7.1 mg/kg/year. In their evaluations of individual genetic risk of humans, based on their measurements of mutagenic activity, the investigators concluded that the 7.1 mgJkg yearly dose of ECH would result in detectable level of genetic injury. c'-Chlorohydrin is an alkylating agent in vitro (Jones et al., 1969), but, like ECH (Sram et al., 1976), it failed to produce dominant lethal mutations (Jones et al., 19694. Carcinogenicity Kotin and Falk (1963) reported that a single subcu- taneous injection of ECH (0.5 ma) resulted in one skin papilloma, a single hepatoma, two lung adenomas (in one mouse), and four malignant lymphomas in a population of 30 mice. However, the control group also developed a number of hepatomas, lung adenomas, and lymphomas, but no skin papillomas. Therefore, results of this study were somewhat . . nconc uslve. Weil et al. (1963) reported that although skin painting with undiluted ECH (3 times/week for 25 months) produced no tumors in 40 mice, ECH was the most systematically toxic of 60 epoxy compounds that were tested. Van Duuren et al. (1972, 1974) investigated the effect of dermal application, subcutaneous injection, and intraperitoneal injection

122 DRINKING WATER AND HEALTH of ECH on female mice. A 580-day skin painting study (2 mg of ECH applied 3 times/week to 50 mice) produced no papillomas or carcino- mas. Another group of 50 mice received one skin application of 2.0 mg followed 2 weeks later with thrice-weekly applications of phorbol myristate acetate (PMA), a tumor promoter, for the duration of the study (385 days). Nine mice developed papillomas, and one developed a carcinoma. Three of 30 mice in a control group that received PMA alone developed papillomas. The subcutaneous injection of ECH (1.0 ma) into 50 mice weekly over 580 days resulted in the development of sarcomas in six mice and an adenocarcinoma in one. In the intraperitoneal experiment, 30 mice received weekly injections of 1.0 mg into the lower abdomen over 450 days. None developed local sarcomas, but 11 mice had papillary lung tumors. On the basis of these studies, the International Agency for Research on Cancer (1976) believes that ECH is "carcinogenic in mice by subcutaneous injection" and that it is "active as an initiator in a two- stage skin carcinogenesis study in mice." The lack of long-term oral ingestion or inhalation studies is a serious deficiency in any analysis of the carcinogenicity of ECH. Nevertheless, the experiments that have been cited do raise some concern about the risks of continuous exposure, especially in the workplace. Nelson ( 1977, personal communication) reported the preliminary results of several long-term inhalation studies with ECH. In the first experiment, rats received 30 exposures of 6 hr each to 100 ppm ECH. Twenty-eight of the initial 40 rats had died by the time the preliminary results were released. Three of them had developed squamous cell cancers of the nasal epithelium, and a fourth had a squamous cell papilloma. A second experiment, started after the first, investigated the effect of chronic exposure to 100, 30, and 10 ppm ECH. At the time of Nelson report, 2 of the 12 rats that were still alive in the 100-ppm group had developed nasal masses that resembled those observed during life in animals with subsequent histopathologically confirmed cancer. Nelson placed the results in perspective when he noted that the "occurrence of these tumors in relatively small numbers should be viewed against the background of experience of this laboratory in which thousands of rats had been used in inhalation studies without the observation of a single squamous cell carcinoma of the nose in control animals." Nelson's preliminary conclusion was that "ECH must be regarded as carcinogenic for the nasal epithelium of the rat." Teratogenicity The subcommittee located no teratogenic studies with ECH (National Institute for Occupational Safety and Health, 1976b).

Toxicity of Selected Drinking Water Contaminants 123 CONCLUSIONS AND RECOMMENDATIONS ECH is a highly reactive alkylating agent that has been shown to be mutagenic and carcinogenic in a variety of test systems. Although ECH is moderately toxic upon acute oral exposure, having LD50 values ranging between 90 and 240 mg/kg, subacute or chronic exposure to this chemical produces a variety of other toxic effects, depending on the route of exposure. Tissue damage occurs primarily in the lungs after inhalation exposure, but ECH that is absorbed via the lungs or from the gastrointestinal tract also elicits blood abnormalities including anemia, leucocytosis or leukemia, liver and kidney damage, and antifertility effects in male animals (Fomin, 1966; Lawrence et al., 1972; National Institute for Occupational Safety and Health, 1976b). ECH is also a potent mutagen (Fishbein, 1976; Sram et al., 1976) and carcinogen (International Agency for Research on Cancer, 1976; Nelson, 1977, personal communication). Because of the well-established carcinogenic and mutagenic properties of the compound, no long-term exposure estimate has been made. Suggested No-Adverse-Response level (SNARLJ Data that are suitable for use ire estimating a 1- or 7-day no-e~ect exposure level for ECH are very sparse. Nevertheless, some attempts at such calculations have been made. Twenty-Four-Hour Exposure Lawrence et al. (1972) observed a significant increase in pentobarbital-induced sleep in ICR mice that had been given a single intraperitoneal 17 mg/kg dose of ECH. Since intraperitoneal and oral LDso values for ECH in the same strain of mice were 170 and 236 mg/kg, respectively, a comparable increase in pentobarbital-induced sleep would probably have occurred with an oral dose of 24 mg/kg. Using an uncertainty factor of 1,000 and assuming that all of the ECH intake is from 2 liters/day consumption of water during this period, the following calculation can be made to yield a suggested no-adverse-e~ect level (SNARL) in drinking water of 0.84 mg/liter for a 1-day exposure to ECH: 24 mg/kg x 70 kg = 0.84 mg/liter Seven-Day Exposure Hahn (1970) found that male rats given daily 15 mg/kg oral doses of ECH became infertile after 1 week of treatment. Using an uncertainty factor of 1,000 and assuming that all of the ECH is

124 DRINKING WATER AND H"LTH ingested from 2 liters/day of drinking water during this period, the following calculation can be made to yield a suggested no-adverse-effect level for drinking water of 0.53 mg/liter for a 7-day exposure to ECH: 15 mg/kg Xli70 kg = 0.53 mg/1iter An almost identical estimate (0.46 mg/liter) may be obtained using the results of Gage (1959), who observed no significant adverse ejects in rats receiving inhalation exposure to ECH at 13 mg/kg/day (National Institute for Occupational Safety and Health, 1976b). The calculated 7-day and 1-day SNARL's in drinking water are only estimates that completely ignore the well-established mutagenic and suspected carcinogenic properties of ECH. Most of the long-term exposure data on ECH are derived from inhalation experiments. Since estimation of long-term risk to humans from ECH in drinking water makes much use of such inhalation data, there is a pressing need for conducting subacute and chronic experi- ments in animals that have been orally exposed to ECH. Only then can assumptions used to extend inhalation results be validated. Methylene Chloride (CH2C12) This compound was evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 743~. The following material, which became available after that 1977 publication, updates and, in some instances, reevaluates the information in the earlier report. Also included are some references that were not assessed in the original report. METABOLISM Exposure to methylene chloride produces an elevation of blood carboxyhemoglobin levels in humans (Stewart en al., 1972a,b), rabbits (Roth et al., 1975), and rats (Rodkey and Collison, 1977~. Biological conversion of methylene chloride to carbon monoxide has now been well established (Rodkey and Collison, 19774. DiVincenzo and Hamilton (1975) administered 412 to 930 mg/kg [~4C]-methylene chloride intra- peritoneally in corn oil to Sprague-Dawley rats. Methylene chloride was largely eliminated unchanged in the expired air during the first 2 hr. After 24 hr. only 2% of the original dose remained in the body. This 2% was found mostly in the liver, kidneys, and adrenal glands. Rodkey and Collison (1977) administered methylene chloride by vaporization in a

Toxicity of Selected Drinking Water Contaminants 125 closed rebreathing system to rats. Addition of methylene chloride to the gas phase caused an initial increase in carbon monoxide production to about 35 times the normal endogenous rate at all doses given (8~800 ,uM/kg). The amount of carbon monoxide that accumulated approached 50% of the dose of methylene chloride that was administered. The data suggest that a low dose of methylene chloride causes substrate saturation of the enzyme system involved in carbon monoxide production. Large doses of methylene chloride caused a progressive increase in the rate of carbon monoxide production over several hours, suggesting that sub- strate-induced enzyme formation occurs. The elevated carboxyhemoglo- bin levels resulting from methylene chloride exposure have a biological half-life that is twice that of the carboxyhemoglobin produced from exposure to carbon monoxide, because the absorbed methylene chloride is released slowly from storage sites in the body tissue. The addition of methanol to paint remover formulations appears to extend the biological half-life of carboxyhemoglobin derived from methylene chloride. Elevation of carboxyhemoglobin, which is caused by inhalation of methylene chloride, was reduced significantly in rabbits pretreated with carbon tetrachloride (Roth et al., 1975~. The biological significance of carboxyhemoglobin levels derived from methylene chlo- ride has not been fully defined. Stewart and Hake (1976) reported that blood carboxyhemoglobin elevations induced by methylene chloride persisted longer than when similar levels of carboxyhemoglobin were induced by breathing carbon monoxide. The literature shows that there is a wide variation in concentrations of carboxyhemoglobin in persons that have been exposed to a given concentration of methylene chloride for a given period and a lack of consistent correlation in measurement for calculation of carboxyhemo- globin. This apparent lack of correlation may have originated in the variations in rate of carbon monoxide formation and the existence of other carboxyhemoglobin-forming events' such as smoking, and/or individual differences in carboxyhemoglobin-eliminating events, includ- ing pulmonary function. HEALTH ASPECTS Observations in Humans Raleigh (1974) studied 562 workers, 103 of whom comprised the highest exposure group, having been exposed to methylene chloride at 50 to 100 ppm. There was no increase in the incidence of cardiovascular, gastrointestinal (including liver), genito- urinary, or central nervous system disease in the exposed groups, as compared with a nonexposed worker population.

126 DRINKING WATER AND HEALTH Winneke and Foder (1976) reported that the central nervous system function was impaired during methylene chloride exposure. Acute exposure to 500 ppm produced a state of reduced activity, which was accompanied by short periods of sleep. The effects of exposure to a paint remover (80% methylene chloride, 20~o methanol by weight) during rest and exercise were reported by Stewart and Hake (1976~. No untoward responses occurred during the 24-hr period following each exposure. None of the four subjects in- this study found the paint remover vapor to be irritating to the eyes, nose, or throat. Nor were there observed abnormalities in electrocardiogram or blood chemistry values; however, the above-cited authors and others have commented on the potential risks of elevated carboxyhemoglobin levels to those individuals with preexisting cardiovascular stress (angina, myocardial ischemia, etc.~. The difficulty in extrapolating from exposure levels of methylene chloride to symptoms is reminiscent of the occasional poor correlation between symptoms of carbon monoxide toxicity and carboxyhemoglobin levels. Langehennig et al. (1976) reported an apparent paradox between greater than 20% carboxyhemoglobin levels resulting from methylene chloride exposure and lack of symptoms. A recent report by Goldbaum et al. (1976), which showed that administered carboxyhemoglobin does not produce manifestations of carbon monoxide toxicity, indicates that carboxyhemoglobin levels may not fully define the toxicity of methylene chloride. Roberts and Marshall (1976) reported a case involving the ingestion of two pints of a paint remover that contained methylene chloride, methanol, cellulose acetate, triethanolamine, paraffin wax, and deter- gent. No carboxyhemoglobin levels were measured. The major acute features of the intoxication included hemolysis, hemorrhage of the gastrointestinal tract, and metabolic activation. They found no hepatic, renal, or cardiac injury. The National Institute for Occupational Safety and Health (1976d) recommended that occupational exposure to methylene chloride not exceed 75 ppm, a time-weighted average exposure for up to a leer workday. Observations in Other Species Acute Effects The acute oral LD50 values are 1.6 to 2.3 ml/kg for rats (Kimura et al., 1971~. In these same studies, Kimura et al. determined the approximate dose that induces the first observable signs of toxic actions (i.e., dyspnea, ataxia, cyanosis, and/or coma). For adult rats the oral dose was 0~001 ml/kg. The intraperitoneal LD50 values are 1.50 ml/kg

Toxicity of Selected Drinking Water Contaminants 127 for mice and 0.95 ml/kg for dogs (Klaassen and Plaa, 1967~. The toxic hazard to the eye from methylene chloride as liquid or vapor was assessed in rabbits by Ballantyne et al. (19761. They observed a small transient increase in corneal thickness and intraocular tension along with evidence of inflammation of the conjunctive. The concentrations required for these ejects were equivalent to a splash contamination of the eye with 0.01 ml of methylene chloride. Chronic Effects In a chronic study, Bornmann and Loeser (1967) reported no adverse effects in rats that had been maintained on drinking water containing methylene chloride at 2.25 g/18 liters for 91 days. In a study on the effects of methylene chloride inhalation, Heppel et al. (1944) reported no adverse effects on dogs and rabbits in a 6-month exposure to 5,000 ppm; however, a slight weight reduction was observed in guinea pigs. Some liver injury was found after 7.5 weeks at 10,000 ppm. Studies by Balmer e' al. (1976), involving exposure of guinea pigs for 5 days to approximately 500 ppm of methylene chloride plus high concentrations of ethanol, suggest that ethanol may potentiate the effects of methylene chloride in the liver (i.e., fatty livers). Mutagenicity Methylene chloride was negative in a Drosophila mutagenicity test (Filippova et al., 1967) but quite mutagenic in the Salmonella typhimurium (TA-100 without S-9 mix) assay (Simmon et al., 1978~. ~ 1 Carcinogenicity Pulmonary tumor response in strain A mice was negative with doses of 160 to 800 mg/kg, which were given 16 to 17 times (Theiss et al., 1977~. Teratogenicity. Methylene chloride vapor was not teratogenic in rats and mice at 1,250 ppm (Schwetz et al., 1975~. Both species were exposed daily for 7 hr on days 6 through 15 of gestation. The investigators found no fetal toxicity or teratogenicity, although there was an increase in the incidence of variation in the development of the sternum. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SlIARLj Twenty-Four-l~our Exposure Data on oral dosing of animals indicate LD,o values ranging between 1.6 and 2.3 ml/kg, which is similar to the intraperitoneal LD50 values of 0.9 to 1.5 ml/kg. The cause of death in

128 DRINKING WATER AND HEALTH these studies is not clear. The minimal-e~ect dose in rats is 1 ml/kg. The water solubility of methylene chloride is approximately 2 ml/dl, which means that humans could theoretically consume a lethal dose in an acute spill situation. Data on a no-effect oral dose do not exist. Using the minimal-e~ect acute oral dose for the rat (1 ml/kg), assuming 2 liters/day of drinking water as the only source, and employing a safety factor of 1,000: 1.3 g/kg x 70 kg 1,000 x 2 liters = 35 mg/liter. Seven-Day Exposure No data are available for calculation. Using the acute 24-hr level of 35 mg/liter and dividing by 7 days: 35 mg/liter 7 = S melter. Chronic Exposure This cannot be calculated due to a lack of adequate chronic (lifetime) exposure data. Polychlorinated Biphenyls (CI2 5H~3C<C6H~3C12 5) These compounds were evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 756~. The following matenal, which became available after that 1977 publication, updates and, in some instances, reevaluates information in the earlier report. Also included are some references that were not assessed in the original report. Polychlorinated biphenyls (PCB's) are a mixture of chlorinated biphenyls that have been produced commercially by the chlorination of biphenyl. PCB's have important heat-resistant properties and have been used in the production of capacitors and transformers. Commercially prepared PCB mixtures often contain between 40 and 60 different chlorinated biphenyls (Aroclors). PCB's are well-known environmental contaminants, and the actual number of chlorinated biphenyls reaching the environment probably is nearly 100 (Pomerantz et al., 1978~. The chemistry of these materials has been reviewed by Hutzinger et al. (1974) and Pomerantz et al. ( 19781. The environmental fate and occurrence of the PC13's are functions of their water solubility, volatility, bioaccumulation, photos/ability, and biodegradability. Some PCB mixtures are known- to be contaminated with varying amounts of chlorinated naphthalenes and chlorinated

Toxicity of Selected Drinking Water Contaminants 129 dibenzofurans. These substances have potent biological actions of their own, which may confound the observed toxicological properties of PCB mixtures. In addition to their occurrence in surface water, PCB's have been detected in U.S. drinking water where their concentration is limited by their solubility (Deinzer et al., 1978~. METABOLISM The biological and toxicological properties of PCB mixtures may vary, depending upon their isomeric composition. The metabolism and biochemical toxicity of the PCB's have recently been reviewed by Matthews et al. (1978~. Radio-labeled chlorobiphenyls with increasing chlorine numbers have a corresponding increase in biological half-life in mammals. The rate of elimination of the chlorobiphenyls is related to the extent of metabolism (Matthews and Anderson, 1975~. The lipophilicity of the chlorobiphenyls increases with increasing chlorine content and also is a factor in their biological disposition. The degradation and elimination of PCB congeners depend upon the hepatic microsomal enzyme system. Two possible mechanisms for biotransformation have been suggested by Ecobichon (1976~. The first and most rapid mechanism involves the formation of an arene oxide intermediate and requires the presence of unsubstituted adjacent carbon atoms in the nucleus. The second and much slower mechanism uses a different hydroxylation system for isolated unsubstituted positions, such as those found in highly chlorinated biphenyls. Two adjacent unsubsti- tuted carbon atoms appear to be important in metabolism. Their presence facilitates the formation of arene oxides by the hepatic mixed- function oxidases. Single-dose oral administration of PCB's to rodents, monkeys, swine, and sheep has indicated that intestinal absorption is rapid and approximately 90% complete (Albro and Fishbein, 1972; Allen and Norback, 1973; Berlin e! al., 1973; Borchard et al., 1975~. The feces provide the major route of excretion. Only traces of PCB's could be found in the urine of animals. In mice, a single oral dose of t~4C]-pentachlorobiphenyl rapidly entered the circulation and was distributed to the liver, kidneys, lungs, and adrenals. Within 24 hr. radioactivity was redistributed to the fat, which remained the major reservoir of unchanged PCB's in the body until only traces remained after 32 days (Berlin et al., 1975a,b). Polychlorinated biphenyls are strong inducers of hepatic mixed- function oxidase enzymes. The potency increases with increasing

130 DRINKING WATER AND HEALTH chlorination of the biphenyl rings. The enzyme-inducing properties of the PCB's are unusual in that the mixtures share the inducing properties of both phenobarbital and 3-methylcholanthrene (3MC) (Alvares et al., 1973~. Phenobarbital induction is characterized by an increase in cytochrome P-450 and an increase in a wide range of enzyme activities, while 3MC induces a more specific range of enzyme activities and cytochrome P~-450 (P-448), a cytochrome differing in several physical and chemical properties from cytochrome P-450. These effects of the PCB's can modify the toxicity of other agents, suggesting that secondary toxication or detoxication reactions that occur in PCB-exposed animals will be a function of the level of exposure, the nature of the secondary agent, and the mechanisms by which the toxicity is produced. HEALTH ASPECTS Observations in Humans PCB's are liver toxins and produce chloracne and, possibly, peripheral neuropathy in humans (Mural and Kuroiwa, 19711. Humans may be exposed to PCB's from occupational exposure, dietary exposure, or exposure to air or water containing PCB's. Cordle et al. (1978) have reviewed exposures of humans to PCB's and the reported effects. Among the diseases that are attributed to PCB's is Yusho, an acute outbreak of which occurred in Japan in 1972. This disease has been ascribed to the ingestion of rice oil containing PCB's although Kuratsune et al. (1976) suggest that the rice oil was contaminated with polychlori- nated dibenzofurans, which may have played a significant role in the observed toxicity. Consequently, the human health effects that occurred in the Yusho incident cannot be attributed entirely to the PCB's (Cordle et al., 19781. Recent surveys have indicated that PCB's can be found in the milk of nursing mothers. The highest reported level was 10.6 ppm. The mean for all samples was 1.8 ppm (U.S. Environmental Protection Agency, 19761. Observations in Other Species Acute Elects PCB's have low acute toxicity. Acute oral LD50 values in rats? rabbits and mice range from 1 to 10 g/kg of body weight. Kimbrough et al. (1978) compiled LD50 values for a variety of PCB mixtures and pure isomers. In addition to the well-known effects on mixed-function oxidase enzymes, high doses of the PCB's may cause enlargement of the liver and liver cells. Other commonly reported effects

Toxicity of Selected Drinking Water Contaminants 131 include decreased reproductive function, renal histopathological change, hepatic porphyria, and an increase in bile flow with a concomitant increase in bilia~y excretion of certain materials such as thyroxine, immunosuppression, thymus atrophy, and lymphopenia. Subchronic and Chronic Elects Because of the bioaccumulation of PCB's, subchronic and chronic effects of these agents are of more concern than acute toxicity. Tucker and Crabtree (1970) reported deaths in rats fed Aroclor at 1 g/kg for 28 to 53 days. Rabbits were given repeated daily oral doses of 300 mg of Aroclor 1221 and 1254 for 14 weeks. Aroclor 1254 produced liver enlargement and damage and one death. Aroclor 1221 produced only minor changes (Koller and Zinkl, 1973~. Allen et al. (1974) administered doses of 25 mg/kg Aroclor 1248 in the diet of six rhesus monkeys for 2 months. This resulted in facial edema, loss of hair, and acne 1 month after onset of feeding. Aulerich et al. (1973) produced 100% mortality in mink within 6 months by feeding them diets containing PCB at 30 mg/kg (10 mg/kg each of Aroclor 1242, 1248, and 1254~. In a study by Ringer et al. (1972), female mink that were fed a diet supplemented with Aroclor 1254 in 5 mg/kg doses for 9 months failed to produce offspring. The oral administration to rats of Aroclor 1242, 1254, and 1260 at 1, 10, and 100 mg/kg for 18 months (Keplinger et al., 1971) produced adverse effects only at the higher dosage. Aroclor 1242 and 1254 caused an increase in liver weight and a reduction in litter survival at 100 mg/kg. Kimbrough et al. (1972) reported experiments in which male rats survived Aroclor 1260 at 1 g/kg for 8 months, but 8 of 10 females died at this dosage. With both Aroclor 1254 and 1260, there was a significant dose-dependent increase in liver weight in male rats produced by doses as low as 20 mg/kg in the diet; in female rats, liver enlargement occurred only at 500 mg/kg and higher. The dietary ingestion of 100 ppm of Aroclor 1248, 1254, or 1262 for 52 weeks by male Sprague-Dawley rats produced no effect on gross appearance or weight gain (Allen et al., 1976~. However, there was an increase in total serum lipids and cholesterol and a transient increase in triglycerides. These changes were accompanied by morphologic alter- ations in the liver such as generalized liver hypertrophy and focal areas of hepatocellular degeneration. Zinkl (1977) produced varying degrees of dermatitis involving the ears and, later, the nose, tail, and feet of rats by feeding them continuously with Aroclor 1254. This lesion was found in 15 of 60 animals fed 100 ppm of the PCB mixture for 10 to 20 weeks. Oral administration of

132 DRINKING WATER AND HEALTH Aroclor 1248 at 2.5 and 5.0 mg/kg produced periorbital edema, alopecia, erythema, and acneiform eruptions within 1 to 2 months (Allen, 1975; Allen and Norback, 1973; Allen et al., 1974~. The feeding of Aroclor 1242 or 1254 to weakling swine and sheep at 20 ppm decreased food efficiency and rate of weight gain over 91 to 105 days of exposure. However, gross and/or microscopic lesions were minor (Hanson et al., 1976~. Mutagenicity Aroclor 1242 and 1254 administered orally to Osborne Mendel rats in five daily doses of 125 or 250 mg/kg and 75, 150, or 300 mg/kg, respectively, failed to produce dominant lethal effects. Treated rats were mated with untreated females for 10 to 11 weeks following treatment (Green et al., 1975a). Furthermore, Aroclor 1242 and 1254 failed to produce cytogenetic effects in bone marrow and spermatogoni- um cells of rats (Garthoff e' al., 1977; Green et al., 1975b). Aroclor 1232, 1254, and 1268 were tested for their ability to inhibit testicular DNA synthesis in mice. These compounds were found to stimulate incorpora- tion of tritiated thymidine into testicular DNA following a single intraperitoneal dose of 500 mg/kg. These changes were not considered significant evidence of an adverse eject (Seller, 1977~. Furthermore, administration of Aroclor 1254 at a dose of 50 mg/kg daily for 7 days failed to cause any significant chromosomal damage or alteration in spermatogenesis for up to 30 days following exposure (Dikshith et al., 1975). Carcinogenicity In a recent review of a number of carcinogenicity studies in which mice and rats had been fed various mixtures of PCB's, Kimbrough et al. (1978) described liver tumors produced by Aroclor 1254 (Calandra, 1975; Kimbrough and Linder, 1974~. Kimbrough et al. (1975) claimed that Aroclor 1260 produced hepatocellular carcinomas in rats. Tumor production, as noted in these studies, was the result of feeding dietary levels of 100 ppm or more. The possible carcinogenicity of Aroclor 1254 was tested in Fischer 344 rats that received diets containing 25, 50, or 100 ppm for 104 to 105 weeks. Twenty-four rats of each sex received each diet, and matched controls consisted of 24 males and 24 females. After 2 years of exposure to PCB's, the treated rats showed a trend toward an increased incidence of lymphomas and leukemias, but the incidence was not statistically different from controls. Some treated animals were found to have hepatocellular adenomas and carcinomas, but these changes were not seen in controls. A high incidence of hyperplastic nodules was also observed in the animals that had been exposed to Aroclor 1254.

Toxicity of Selected Drinking Water Contaminants 133 Hyperplastic nodules are regarded by many to be preneoplastic. The incidences of tumors were not significant. Consequently, the investiga- tors concluded that Aroclor 1254 was not carcinogenic in Fischer 344 rats; however, the high incidence of hepatocellular proliferative lesions was believed to be related to administration of the chemical (National Cancer Institute, 1978~. Teratogenicity Several studies have demonstrated that PCB's can cross the placenta, but no studies have detected defects of embryo-fetal development (Curley et al., 1973; Kato et al., 19721. In contrast, other studies have reported that the placenta acts as an effective barrier to the transfer of PCB's (Baker et al., 1977~. In addition, Funatsu et al. (1972) and Miller (1971) have linked maternal ingestion of PCB with dark- brown staining of the skin of newborn babies. PCB's have no known or clearly defined teratogenic effects in mammals although the transfer of PCB's across the placenta suggests the potential for some form of fetal toxicity (Kimbrough et al., 1978~. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARL) Twenty-Four-Hour Exposure Because the PCB's are a complex mixture of isomers and impurities }raving various biological activities and environmental fates, it is risky to suggest no-adverse-response limits. However, one of the most sensitive indicators of exposure to the PCB's is the induction of the mixed-function oxidase enzymes of the liver of mammals. Using this effect, the minimal no-adverse-response to Aroclor 1254 in rats is 2 to 5 ppm in the diet, which is the apparent threshold for these effects (Grant et al., 19741. Moreover, as little as 1 to 2 mg/kg of Aroclor 1254, given within 24 hr. is sufficient to stimulate mixed-function oxidase enzymes in rats (Bruckner et al., 1977~. If this end point is used as an indicator of minimal toxicity, the following 24-hr SNARL value may be calculated for the PCB class (without reference to any particular isomer or contaminant). Assuming that 100% of exposure is from water during this period and that a 70-kg human consumes 2 liters of water per day, and using an uncertainty factor of 100: 1 mg/kg x 70 kg 0 35 mg/liter The uncertainty factor of 100 was chosen because of the large number of studies that substantiate the effect and dosage described above.

134 DRINKING WATER AND HEALTH Seven-Day Exposure Using the same argument, a 7-day SNARL may be calculated: 24-hr SNARL = 0.35 mg/liter = 005 mg/liter Chronic Exposure A reliable chronic SNARL may not be calculated for the PCB's for the reasons noted above and also because certain PCB's are suspected carcinogens. Although there are considerable data on the toxicity of mixtures of PCB's, there is a paucity of data on the pure congeners that are present in these mixtures. Whether chronic toxicity is related to the metabolism of the PCB's remains to be determined. Considerably more attention must be directed to the detection of impurities in PCB's at very low concentrations. Polychlorodibenzofurans may constitute only one of several significant contaminating compounds that are responsible for PCB toxicity. Populations at special risk- both the industrially exposed and those heavily exposed by the ingestion of contaminated foods should be evaluated carefully. Despite the lack of evidence in the United States that dietary PCB's have any deleterious effects on health, there is a growing concern about the long-range effects of the contamination of our ecosystem with these chemicals. There is an urgent need for epidemiological studies of exposed populations, more precise identification of all sources of PCB contamination, and efforts directed at the control of disposal of PCB's. Because of the demonstrated carcinogenic potential, studies on individu- al congeners both those metabolized and those stored by humans are urgently needed. Tetrachloroethylene (Cl2C=CCl2) This compound was evaluated in Drinking Water and H. alth (National Academy of Sciences, 1977, p. 769~. The following material, which became available after that 1977 publication, updates and, in some instances, reevaluates information in the earlier report. Also included are some references that were not assessed in the original report. METABOLISM The pharmacokinetics of tetrachloroethylene (ICE) has been studied extensively in humans who have been exposed to vapors of the chemical. Unfortunately, little work has been done to delineate the uptake,

Toxicity of Selected Drinking Water Contaminants 135 distribution, metabolism, and excretion of oral doses of TCE. Studies have demonstrated that TCE is rapidly absorbed through the lung (Fernandez et al., 1976) and skin (Stewart and Dodd, 19644. Since TCE is a small, uncharged, lipophilic molecule, one would expect that it would be rapidly and completely absorbed from the gastrointestinal tract as well. Blood levels of 2 to 4 ppm TCE have been measured in persons who have breathed 100 to 150 ppm continuously for several hours (Hake and Stewart, 1977). As discussed below, 100 ppm appears to be the minimum effective concentration in vapor that is responsible for subjective complaints and minor necrologic dysfunction in humans. The majority of systemically absorbed TCE is rapidly eliminated by expiration of the unchanged compound. A small proportion, which is believed to accumulate primarily in adipose tissue, is eliminated much more slowly. A relatively minor portion of the total TCE that is absorbed into the body is metabolized and slowly excreted in the urine (Fernandez et al., 19761. Ikeda (1977) calculated that the respiratory half-life in humans is approximately 65 hr. while the urinary half-life is about 144 hr. Fernandez et al. (1976) detected TCE in expired air of subjects for more than 2 weeks following an 8-hr, 200-ppm exposure, and Hake and Stewart (1977) measured 4 ppm TCE in expired air of a man 25 days after he was overcome by TCE fumes. Therefore, the chemical might be expected to accumulate in the body if insufficient time is allowed to elapse between subsequent exposures. Such an accumulation was reported by Stewart et al. (1970a) during 5 days of daily 7-fur exposures of humans to 100 ppm TCE. Ikeda (1977) stated that TCE accumulates in the body at 3 to 4 times the rate of trichloroethylene, a relatively long- lived alkyl halide. Only a small proportion of systemically absorbed TCE is metabolized. Fernandez et al. (1976) estimated that only approximately 1.85% of the TCE absorbed by humans during an 8-hr, 150-ppm inhalation exposure is eliminated in the urine as trichloroacetic acid (TCA). The primary metabolites of TCE in the mouse (Yllner, 1961), rat (Daniel, 1963), guinea pig (Sakamoto, 1976), and human are TCA and chloride. Some investigators have reported trichloroethanol as a secondary metabolite in human urine (Ikeda, 1977; Ikeda et al., 1972), although others (Fernan- dez et al., 1976; Hake and Stewart, 1977) did not confirm this. Still other investigators observed? but could not positively identify, in human urine (Ogata et ale 1971) and guinea pig urine (Sakamoto, 1976) what may be trichloroethanol. This alcohol is undoubtedly more toxic (Mikiskova and Mikiska, 1966) than the relatively innocuous TCA (Woodard et al., 1941), although only small quantities of either metabolite would be expected to be formed from TCE. Despite wide individual variability in

136 DRINKING WATER AND HEALTH urinary TCA and trichloroethanol levels, Ikeda (1977) noted that the capacity of humans to metabolize TCE is quite limited, even at low-level exposures. Stimulation of TCE metabolism may enhance its toxicity. Results of studies of this phenomenon are conflicting, in that pretreatments of rats with ethanol (Cornish and Adefuin, 1966) and phenobarbital (Cornish et al., 1973; Moslen et al., 1977a) fail, but polychlorinated biphenyl (PCB) (Moslen et al., 1977a) succeeds in potentiating TCE hepatotoxicity. Moslen and her coworkers found that while phenobarbital enhances urinary excretion of TCE metabolites, PCB is even more effective. The initial step in TCE metabolism is believed to involve formation of TCE oxide (Ikeda, 1977~. Thus, production of increased quantities of the epoxide and/or subsequent cytotoxic products might account for toxicity potentiation. Sakamoto (1976) reported that TCE oxide is relatively toxic and may be carcinogenic or mutagenic, although Greim et al. (1975) noted that it is relatively stable and is not mutagenic in their microsomally supplemented Escherichia cold screening system. HEALTH ASPECTS Observations in Humans Although TCE is a potent depressant of the central nervous system, residual organ damage is not commonly observed in humans who have been exposed to large quantities of the compound. TCE was formerly used widely as an intestinal anthelmin- thic. Lambert (1933) reported that oral doses of 2.8 to 4.0 ml given for this purpose were quite effective and safe. Inebriation was the only troublesome side effect that was noted in 46~000 treated patients. Inhalation of levels of TCE sufficient to produce inebriation and unconsciousness has failed to elicit hepatic, renal, or hematological abnormalities in some individuals (Method, 1946; Patel et al., 1977~. However, in other cases, mild (Hake and Stewart, 1977; Saland, 1967; Stewart et al., 1961a; Stewart, 1969) to severe (Meckler and Phelps, 1966) hepatotoxicity has been diagnosed. In most such instances, liver injury was not manifest until several days after exposure. Recovery was uneventful, but sometimes prolonged, particularly in the more severe cases. TCE was quite slowly eliminated, in that approximately 1 ppm TCE was measured in the breath of victims as long as 11 to 12 days after exposure (Stewart et al.. 1961a; Stewart 19691. Little evidence of kidney injury or damage of any other organ was noted in any of the aforementioned cases, other than the nephrotoxicity that was reported by Hake and Stewart (19771.

Toxicity of Selected Drinking Water Contaminants 137 Experiments involving intentional exposure of humans to TCE vapor demonstrate that low levels of the chemical can cause irritation of mucous membranes and intoxication, but not permanent injury. Stewart et al. ( 1961 b), like Rowe et al. ~ I 952a), reported the minimum intoxicating vapor level of TCE to be approximately 100 ppm. Stewart and his coworkers found that the concentration of TCE in the blood rises gradually during a 3-hr, 194-ppm exposure to a maximum level of approximately 2.5 ppm (TCE ir1 blood). The investigators noted that although no TCE is detectable in the blood 30 min after exposure, the compound is measurable in expired air for up to 94 hr. Such findings raise some questions as to the efficiency/sensitivity of their blood extraction and analyses. Stewart et al. (1961b) reported that there were no abnormalities in laboratory tests for hepatotoxicity in their test subjects. A subsequent investigation by Stewart et al. (1970a) confirmed findings in the previous study, namely, that 100 ppm TCE was a threshold level for induction of early depression of the central nervous system. Altered electroencephalogram patterns suggestive of drowsiness were seen in a majority of male and female subjects who breathed 100 ppm of TCE for 7.5 hr (Hake and Stewart, 1977~. Extensive evaluation of subjects, including blood chemistry, urinalyses, pulmonary function testing, electrocardiogram, and visual acuity testing, failed to reveal adverse elects during a 5-day, 100-ppm exposure regimen (Stewart et al., 1 970a). Such experimental and epidemiological evidence led to the recommendation that 50 ppm (as a time-weighted average) be the occupational exposure standard (National Institute for Occupational Safety and Health, 19761. The National Institute for Occupational Safety and Health believed that this standard would protect against not only tissue injury but also temporary neurological dysfunction and respiratory tract irritation. Observations in Other Species Acute Effects Toxicological investigations with laboratory animals have confirmed observations in humans, namely, that TCE in sufficient quantities can depress the central nervous system but has quite limited ability to damage organ systems. The acute oral LD50 was shown to be 4,000 mg/kg in dogs and 5.000 mg/kg in rabbits (National Institute for Occupational Safety and Health, 1976~. A number of investigators have demonstrated that TCE is one of the least hepatotoxic and nephrotoxic of a series of alkyl halides when administered as a single dose by

138 DRINKING WATER AND HEALTH intraperitoneal injection to several species of animals (Cornish et al., 1973; Klaassen and Plaa, 1966, 1967; Plaa and Larson, 1965~. "Near- lethal" doses of TCE were usually required to produce significant tissue injury. Although alterations in organ function tests, serum enzyme levels, and histopathology were observed, the extent of these changes was relatively modest. The lowest reported toxic intraperitoneal doses of TCE, as indicated by increased serum enzyme levels, were approximately 0.3 to 0.5 ml/kg in rats (Cornish et al., 1973) and 0.7 ml/kg in dogs (Klaassen and Plaa, 1967~. Similarly, Moslen et al. (1977b) observed that a single oral dose of 0.75 ml/kg elicited elevated levels of serum glutamic oxaloacetic transaminase in rats. Kylin et al. (1963) found that 200 ppm was the lowest vapor level of TCE to produce fatty infiltration of the liver of mice following a 4-fur exposure. Chronic Effects Investigations of chronic toxicity of TCE in animals have all involved inhalation exposure, with the exception of a recently concluded assessment of carcinogenesis, which involved oral dosing (National Cancer Institute, 1977~. Male and female mice and rats that were gavaged with doses of TCE ranging from 386 to 1,072 mg/kg/day for as long as 78 weeks exhibited a very high incidence of nephrotoxicity but little evidence of hepatotoxicity (National Cancer Institute, 1977~. There was a high incidence of hepatocellular carcinoma in the mice, but not in the rats. A preliminary report by Pegg et al (1978) of a 6 hr/day, 5 days/week, 12-month exposure of rats to 600 ppm TCE stated that no organ toxicity was observed. Mutagenicity The literature contains only one investigation concern- ing the mutagenic potential of TCE. In the presence of a microsomal activating system, TCE was unable to induce mutations in Escherchia cold (Greim et al., 19751. In contrast, vinyl chloride was a quite potent mutagen under these conditions. The authors, assuming that both TCE and vinyl chloride are initially oxidized to epoxides, proposed that the epoxide of vinyl chloride was relatively unstable and might, therefore, more readily alkylate biological molecules and induce mutagenesis and carcinogenesis. Carcino~;enicity Results of the limited number of studies of the carcinogenic potential of TCE are conflicting. Although thorough gross and histopathological evaluations of tissues of several species were performed in subacute and chronic studies, no mention of tumor induction has been made (Kylin et al., 1963; Pegg et al., 1978; Rowe et al., 1952a). In an assessment of the ability of TCE to produce pulmonary

Toxicity of Selected Drinking Water Contaminants 139 adenomas in male mice, Theiss et al. (1977) failed to find a carcinogenic effect. However, the NCI (National Cancer Institute, 1977) has reported TCE-induced hepatocellular carcinomas in male and female mice, but not in male or female rats. In the NCI study, daily oral doses ranging from 386 to 1,072 mg/kg were administered for up to 78 weeks. A high tumor incidence was observed at both low and high dose levels. The quantities of TCE given were so large that marked, dose-dependent mortality in both species occurred throughout the study period. Teratogenicity Schwetz et al. (1975) examined the teratogenic ejects of TCE in rats and Swiss Webster mice. The animals were exposed at 300 ppm for 7 hr/day on days 6 through 15 of gestation. The investigators concluded that TCE had little eject on embryonic and fetal develop- ment, and that it was not teratogenic. Nevertheless, there were a number of modest but statistically significant deviations from controls, including increased weight of maternals, decreased body weight of mouse fetuses, increased fetal resorptions, and increased incidence of split sternebrae, subcutaneous edema, and delayed ossification of skull bones in mouse fetuses. Carcinogenic Risk Estimate In a recent study by NCI (1977), rats and mice were Savaged with TCE doses ranging from 386 to 1,072 mg/kg/day for as long as 78 weeks. There was a high dose-related incidence of hepatocellular carcinoma in the mice, but none in the rats. Each set of dose-response data was used to make statistical estimates of both the lifetime risk and an upper 95% confidence bound on the lifetime risk at the low dose level. These estimates are of lifetime human risks and have been corrected for species conversion on a dose/surface area basis. The risk estimates are expressed as a probability of cancer after a lifetime consumption of 1 liter/day of water containing 1 ~g/liter of the compound. For example, a risk of 1 x 10-6 implies a lifetime probability of 2 x 1O-5 of cancer if 2 liters/day were consumed by a 70-kg human and the concentration of the carcinogen was 10 ppb. This means that at a concentration of 10 ppb during a lifetime exposure, this compound would be expected to produce one excess case of cancer for every 50,000 persons exposed. Assuming the population of the United States to be 220 million, this translates into 4,400 excess lifetime deaths from cancer or 62.8 per year. For humans exposed to 1 ~g/liter TCE, the estimated lifetime risk is 6 2 x 10-~. The upper 95% confidence estimate is 1.4 x 1O-7. Both of these estimates are the averaged risks, which have been calculated from the male and female mice.

140 DRINKING WATER AND HEALTH CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level fSNARLJ Although acute expo- sure to high levels of TCE can result in marked depression of the central nervous system, the chemical apparently has quite limited capacity to cause tissue injury. The majority of the information that is relative to human and animal exposure concerns inhalation of TCE. Although duration and levels of exposure to TCE in vapor have been reported, body burdens of the chemical have generally not been determined. Therefore, it is quite difficult to relate toxicological findings in these reports to either blood levels or toxic effects that might result from ingestion of TCE. No accounts of toxicity resulting from ingestion of TCE by humans were found in the literature. The subcommittee has only the historical perspective of Lambert (1933) on the use of TCE as an anthelminthic, when oral doses of 2.8 to 4.0 ml apparently elicited little more than inebriation. Similarly, inhibition of psychophysiological function would appear to be the most sensitive index of exposure to TCE via inhalation. Twenty-Four-Hour Exposure In light of the lack of definitive informa- tion regarding the quantity of TCE that must be ingested to depress psychophysiological function, it seems most appropriate that calcula- tions for a SNARL be based upon quantities of the chemical that are required to produce tissue injury. The lowest reported intraperitoneal doses of TCE that produce increased serum transaminase levels in various species range from 0.3 to 0.7 ml/kg. A single oral administration of 0.75 ml/kg results in an apparent elevation in serum transaminase activity in rats. Therefore? the 0.3 ml/kg (0.49 g/kg) dose appears to be a reasonable "minimum toxic dose" from which to calculate a 24hr SNARL for contamination of drinking water, assuming that the sole' source of TCE dunug this period will be from 2 liters/day of drinking water consumed by a 70-kg human. A safety factor of 100 is applied: 490 mg/kg x 70 kg = 172 mg/1iter. The above considerations ignore the possibility that TCE may be carclnogemc. Seven-Day Exposure Unfortunately, there have been no studies of the subacute toxicity of ingested TCE in laboratory animals or in humans. Investigations involving five consecutive daily exposures of humans to

Toxicity of Selected Drinking Water Contaminants 141 100 ppm of TCE vapor have failed to reveal evidence of organ damage (Stewart et al., 1970a). These studies did reveal some propensity of the chemical to accumulate in the body. Nevertheless, in view of TCE's relative lack of toxicity, a 7-day standard for drinking water contamina- tion, which was obtained by dividing the 24-hr standard by 7 (172 mg/liter/7 days = 24.5 mg/liter), should protect against adverse effects by the chemical. This standard is based upon the assumption that the sole source of TCE during this period will be drinking water. These considerations ignore the possibility that TCE may be carcinogenic on short-term exposure. Chronic Exposure From the limited number of chronic studies it is unclear whether TCE is a carcinogen. NCI (1977) reported that it induces hepatocellular carcinomas in one strain of mice, but not in rats. The findings of this study should be interpreted with caution, recognizing the limitations of the experimental design (e.g., massive doses of TCE, large volumes of oil vehicle, marked nephrotoxicity, diminished life- span). The absence of histopathological change in organs other than the kidney in this 78-week study supports the assumption that TCE is relatively nontoxic, even when ingested repeatedly in large quantities. Unfortunately, due to the failure of the NCI study to determine a "no- effect level" and to monitor a sufficiently broad battery of sensitive toxicological indices, it is not possible to establish a "minimum-effect level" for chronic, noncarcinogenic toxicity. Although TCE is not overtly toxic, appropriately designed studies should be conducted with several species of animals to determine the minimum toxic dose for both single ingestions and multiple ingestions over a 1-week period. Attention should be focused on the liver and kidney, but care should be taken not to overlook effects on other organs. Assuming that inhibition of psychophysiological function is the most sensitive index of TCE-exposure in humans, studies of oral doses in volunteers might be conducted to determine whether small quantities of the chemical are intoxicating. In the absence of such information, one can only guess at the maximum no-effect dose in short-term exposure situations. Chronic, low-level feeding studies should also be conducted to establish with some degree of certainty whether and under what conditions TCE may be a toxicant, mutagen, teratogen, and/or carcino- gen. Such investigations should be conducted with several species of animals, should establish maximum no-effect levels, and should examine a range of TCE concentrations that realistically approximates antici- pated potential exposures. Although the majority of systemically

142 DRINKING WATER AND HEALTH absorbed TCE is exhaled unchanged, with only a small proportion excreted as urinal metabolites, it appears pertinent to determine whether TCE or its metabolites can accumulate as a result of prolonged, low-level ingestion. Thorium (Thy Thorium occurs naturally in a number of minerals including monacite and thorite (Anonymous, 1959~. Its concentration in the water of rivers, lakes, seas, and oceans varies between 10-5 and 10-9 g/liter. About half of the amount is present as carbonate complexes (Lyarskii et al., 1970~. It is the parent element of a naturally occurring radioactive family. It decays through a number of isotopes such as mesothorium, radiothorium, thorium, thorium emanation bodies, thoron gas, and, eventually, lead (208Pb), emitting alpha and beta particles and gamma rays. Most of the radiation is in the form of alpha particles (4.0 4.2 mev). The half-life of 232Th is 1.4 x 10~° years. Thorium is used principally as a fertile material for nuclear breeder cores, in the manufacture of incandescent lamps and mantles, as a catalyst in organic syntheses, and as an alloy in welding electrodes (Anonymous, 19591. METABOLISM The distribution of thorium compounds depends upon an interplay between the concentration of the thorium and mode of administration (Pavlovskaya, 1969~. Distribution of thorium is primarily dependent upon concentration; i.e., at concentrations less than 10-5 g/ml, thorium is distributed primarily to the skeleton, independent of the route of administration, while at concentrations greater than 10-3 g/ml, distribu- tion depends upon route of administration. At higher doses (> 10-3 g/ml) following intravenous administration, thorium (dioxide or citrate) is distributed primarily to the reticuloendothelial system with about 60570 to 903tO in the liver and the remainder in the spleen and marrow (Berenbaum and Birch, 1953~. Following inhalation or intratracheal administration, 7097O to 80% of the thorium dose is found in the lungs. Following intramuscular, intraperitoneal, or subcutaneous injection, more than 90% of the dose remains at the site of injection. Following oral doses, the bulk of the thorium is distributed to the skeleton since the absorption coefficient is very low (not more than 0.05%) (Pavlovskaya et al., 1971~. Administration of lower doses of thorium (~1 mg/ml) always leads to accumulation in the skeleton. Thus, thorium entering via the diet or drinking water of humans tends to be distributed to the skeleton.

Toxicity of Selected Drinking Water Contaminants 143 HEALTH ASPECTS The salts of thorium are known to be associated with chemical as well as radiological hazard (Tendon et al., 1975~. Some reports have stated that thorium causes pathological changes in the spleen, liver, lungs, and hematopoietic organs. Long-term effects include production of malig- nant neoplasms (Pavlovskaya, 1969~. Observations in Humans There have been numerous reports of clinical carcinogenicity of thorium compounds, especially thorium dioxide (Thorotrast), which was used as a radiopaque contrast medium (Schmitz et al., 1970~. Cancers that were produced reflect the distribution and dose of the thorium that was administered. Consequently, after intravenous administration of thorium dioxide, most cancers were tumors of the reticuloendothelial system especially involving the liver with an equal distribution of sarcomas and carcinomas. Injection of thorium dioxide into the sinus cavities results in carcinomas. The latent period for tumor development is long (approximately 15 years), and tumorigenesis arises from a-radiation. Observations in Other Species Acute Toxicity McClinton and Schubert (1948) reported that the intraperitoneal LD50 of thorium nitrate in albino, female Sprague- Dawley rats was 68 + 12 mg/kg. Subacute Toxicity There are few reports on subacute toxicity of thorium. Tandon et al. (1975) applied daily 1-ml doses of of 5%, 104370, and 2097O thorium nitrate to the skin in the lateroabdominal and scrotal areas of male albino rats for 15 days. The animals were sacrificed 15 days later. None of the animals showed symptoms of morbidity or mortality during or after treatment. There was no gross abnormality in skin, liver, kidneys, or testes of any of the rats. However, upon microscopic examination, mild degenerative changes were found in the skin and testes of animals receiving the 20% thorium solution. Skin lesions consisted of mild acanthosis and thickening of epithelial lining. Testicu- lar lesions consisted of mild edema of seminiferous tubules and interstitium and desquamation of the spermatogenic cells. A few tubules carried spermatid-type giant cells. Chronic Toxicity No available data.

144 DRINKING WATER AND HEALTH Mutagenicity No available data. Carcinogenicity See Observations in Humans. Teratogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS Because of the paucity of data on acute and chronic oral toxicity of thorium, and since thorium produces cancer in humans, estimates for exposure limits will not be calculated. However, the following standards have been set for thorium by various groups: 1. International Commission on Radiological Protection (ICRP) 0.5 g/liter soluble 232T~; 0.5 to 9 g/liter for insoluble compounds (Lyarskii et al., 19703. 2. Council for Mutual Economic Aid-4 mg/liter for 232Th; 2.7 mg/liter for natural thorium (Lyarskii et al., 1970~. 3. Sanitary Regulations (USSR) 0.1 mg/liter for both 232Th and natural thorium, based on consumption of 2.2 liters of water per day and the skeleton as the critical organ (Lyarskii et al., 1970~. 4. Desired Maximum Ambient Environmental Levels for Elemental Forms Based on Chemical Toxicity (Battelle Laboratories, 19703 0.0005 mg/liter of potable water. This value was calculated from acute median lethal dose (LD50) by intraperitoneal injection and biological half-life. There is a lack of data concerning acute and chronic oral toxicity of thorium compounds. Although injected thorium produces a significant risk of carcinogenicity, available data do not provide evidence of risk from thorium in drinking water. Whether the carcinogenic effect is due to the radioactive properties of this element needs to be assessed. 1, 1,1-Trichloroethane (Cl3C CH3) 1, 1,1-Trichloroethane (TCE), or methyl chloroform, is used widely as an industrial chemical for such purposes as a cleaner and degreaser of metals, a spot remover, and a solvent of lipophilic substances. It is a clear, colorless liquid at room temperature; its solubility in water is 4,400 mg/liter at 20°C (Verschueren, 1977~; its boiling point is 74°C; and its vapor is heavier than air and nonflammable. Dioxane is commonly added to promote its stability. TCE has been identified in drinking water

Toxicity of Selected Drinking Water Contaminants 145 supplies in the United States (U.S. Environmental Protection Agency, 1978c). METABOLISM There have been extensive studies on the uptake and distribution of TCE in humans and laboratory animals that have been exposed to the chemical by inhalation. Unfortunately, there is little information on the pharmacokinetics of ingested TCE. One would expect that this com- pound would be readily absorbed from the gastrointestinal tract in light of the report by Stewart and Dodd (1964) who found that it penetrated intact human skin. Astrand et al. (1973) reported rapid absorption of inhaled TCE and measurable levels of the chemical in the arterial blood of human subjects after inhaling 250 ppm TCE for 10 s. The arterial blood contained substantially higher TCE levels than the venous blood throughout a 2-fur exposure, indicating ready uptake of the compound from blood into tissues. Blood concentrations of 3 to 5 ppm have been measured in humans breathing 350 ppm TCE (Astrand et al., 1973; Gamberale and Hultengren, 1973; Stewart et al., 1961c), the current threshold limit value for occupational exposure in the United States. Apparently, there are no data on blood levels of the compound following oral exposures. The majority of systemically absorbed TCE is eliminated via the lungs. Hake et al. (1960) reported that about 98.7% of a 700 mg/kg dose of radio-labeled TCE, which was injected intraperitoneally into rats, was exhaled unchanged within 25 hr. They also observed small amounts of radio-labeled carbon dioxide in expired air and of the glucuronide conjugate of 2,2,2-trichloroethanol in urine. Later studies revealed trichloroacetic acid to be a second metabolite, although the amounts that were formed in the urine of both rats (Eben and Kimmerle, 1974; Ikeda and Ohtsuji, 1972) and humans (Stewart ei al., 1969) were substantially less than those for trichloroethanol. No chloral hydrate was detected in the blood or tissues of rats by Eben and Kimmerle (1974~. In studies by Van Dyke and Wineman (1971) and Ikeda and Ohtsuji (1972) on the metabolism of a series of alkyl halocarbons, TCE was one of the least extensively metabolized compounds. Nevertheless, it is known to exhibit type I binding characteristics with cytochrome P450 (Pelkonen and Vainio, 1975) and to be capable of inducing microsomal enzyme and P-450 activity (Fuller et al., 1970~. A progressive increase in urinary output of trichloroethanol was observed in humans that had b~een subjected to five daily inhalation exposures to TCE. This indicates that TCE induces its own metabolism (Stewart et al., 1969~.

146 DRINKING WATER AND HEALTH Upon termination of TCE exposure, the chemical is rather quickly eliminated in exhaled air. Levels in the blood and alveolar air decrease exponentially, showing an initial rapid fall, followed after several hours by a somewhat slower decline (Astrand et al., 1973; Stewart et al., 1969~. This latter stage probably reflects slow mobilization of the agent from lipoidal tissues. Stewart et al. (1969) reported that a slight amount accumulated in humans who inhaled TCE at 500 ppm, 6.5 to 7 hr/day for 5 days, despite the relatively rapid loss of TCE from the body. These investigators also found TCE in breath samples from one subject 1 month after exposure. Thus, it appears that TCE can accumulate in the body if intake is frequent enough and/or of sufficient magnitude. The relative lack of toxicity of TCE, in comparison to certain other alkyl halocarbons, can be attributed to its relatively rapid elimination and stability. This compound is much more volatile (Ikeda and Ohtsuji, 1972) and, therefore, more readily excreted via the lungs (Morgan et al., 1970) than the more toxic congener 1,1,2-trichloroethane. Although neither halocarbon is metabolized to a significant degree, Carlson (1973) observed that microsomal enzyme induction with phenobarbital potenti- ates hepatotoxicity of both TCE and 1,1,2-trichloroethane. Thus, it appears that the metabolitets) of each compound is responsible for cytotoxicity, although the identity and mechanism of the actual toxicantts) remain unknown. HEALTH ASPECTS Observations in Humans The primary toxic effects in humans who have been subjected to short-term, high-level exposure to TCE are manifesta- tions of depression of the central nervous system. In the majority of reports of human fatalities resulting from inhalation of TCE, death is attributed to a functional depression of the central nervous system. Levels of TCE in the victims' blood vary considerably, generally ranging from 60 (Hatfield and Maykoski, 1970; Stahl et al., 1969) to 720 ppm (Hall and Hine, 1966~. As might be predicted, the highest concentrations of TCE are found in the brains of victims (Caplan et al., 1976; Stahl et al., 19691. Due to problems that are inherent in analyses of volatile toxicants in autopsies, it is difficult to establish lethal TCE concentra- tions in blood or tissue. Inhalation of high concentrations of TCE can cause irritation of the respiratory tract and minimal organ damage, as well as depression of the central nervous system. Acute pulmonary congestion and edema typically found in fatalities result from inhalation of TCE (Bonventre et

Toxicity of Selected Drinking Water Contaminants 147 al., 1977; Caplan et al., 1976~. There are also scattered reports of modest fatty vacuolation in the liver (Caplan et al., 1976; Hall and Hine, 1966; Stahl et al., 1969~. In most such instances there probably would have been insufficient time between exposure and death for hepatotoxicity to be fully expressed. Stewart (1971) reported the case histories of four individuals who were monitored clinically after being overcome by TCE vapors. In each case, recovery from depression of the central nervous system was quite rapid and largely uneventful. However, one of the four patients exhibited elevated urinary urobilinogen but no alteration of other indices of hepatotoxicity. These studies indicate that TCE possesses a limited capacity to exert hepatic injury in cases of acute, high-level inhalation exposure. Clinical experience and scientific investigations suggest that acute high-level inhalation of TCE can adversely affect the cardiovascular system of humans. Dornette and Jones (1960) used concentrations of 10,000 to 26,000 ppm TCE to anesthetize surgery patients. They noted that both induction of and recovery from anesthesia were quite rapid. No evidence of respiratory depression or hepatotoxicity was seen. However, there were disturbing cardiovascular effects including dimin- ished systolic pressure, premature ventricular contractions, and, in one patient, even cardiac arrest. Bass (1970) reported a syndrome termed "sudden sniffing death" in persons dying abruptly while inhaling volatile solvents for self-intoxica- tion. TCE was one of the most frequently implicated solvents in such incidents. The fatalities were tentatively attributed to cardiac arrhyth- mias that resulted from a combined action of the solvent and endoge- nous biogenic amines. Recent investigations of the phenomenon with laboratory animals are discussed below. A single account of ingestion of TCE by a human has appeared in the literature (Stewart and Andrews, 1966~. A 47-year-old male mistakenly consumed 1 oz of TCE (approximately 0.6 g/kg). He became nauseated within 30 min and developed progressively severe vomiting and diarrhea over the next few hours. Clinical evaluation following gastric ravage revealed neither drowsiness nor difficulty with coordination. Urinalysis and clinical chemistry tests revealed evidence of only minimal hepato- renal injury early in the course of hospitalization. After resolution of the vomiting and diarrhea, the patient was asymptomatic during a 2-week observation period. Since depression of the central nervous system is the predominant effect of TCE on humans, certain manifestations of the depression should be the most sensitive indices of the physiological action of small quantities of,the solvent. Early studies with volunteers indicate that

148 DRINKING WATER AND HEALTH inhalation of 500 ppm TCE for several hours has no significant effect other than transient, mild eye irritation (Stewart et al., 1961a,b; Torkelson et al., 1958~. Stewart and his coworkers (1969) concluded in a later study that 500 ppm may be excessive for persons who are particularly susceptible to the chemical's depressant effects on the central nervous system. In an even more recent investigation, inhalation of 350 ppm TCE for 4 hr was not effective, but 450 ppm elicited subjective complaints of transient eye irritation and dizziness (Salvini et al., 1971a,b). Although a battery of psychophysiological tests did not reveal a statistically significant degree of functional inhibition, lower scores resulted when tests were conducted during TCE exposure than when under control conditions. Results of an investigation by Gambe- rale and Hultengren (1973) indicated that inhalation of 350 ppm TCE can significantly inhibit psychophysiological functions of humans. Blood levels in these "inhibited" subjects averaged approximately 3 to 4 ppm, although the investigators noted wide intersubject differences in blood and alveolar air concentrations. Gamberale and Hultengren concluded that it would be difficult, with any degree of accuracy, to set a threshold for the vapor concentration of TCE that would not alter function of the central nervous system. Their tests of psychophysiological function are certainly more sensitive and objective than the indices used in the earlier studies of Torkelson et al. (1958) and Stewart et al. (1961, 1969~. Nevertheless, the current U.S. threshold limit value for occupational exposure to TCE remains at 350 ppm. This standard is designed to protect the majority of workers from mucous membrane irritation and performance inhibition. One interesting facet of the studies by Torkelson et al. (1958) and Stewart et al. (1969) is their failure to find any evidence of organ damage in humans that were subjected to acute TCE inhalation regimens. Short-term exposure to TCE appears to be no more harmful to humans or laboratory animals than is acute exposure. Stewart et al. (1969) exposed humans via inhalation to 500 ppm TCE for 6.5 hr daily for 5 consecutive days. They observed some objective and subjective signs of depression of the central nervous system, but no evidence of toxicity upon examination for neurological, respiratory, and hepatorenal function. There were also a small accumulation of TCE and an increase in urinary trichloroethanol levels. Observations in Other Species Acute Elects Overall results of animal experimentation confimn the previously described findings in humans-namely, that TCE is relatively

Toxicity of Selected Drinking Water Contaminants 149 nontoxic upon short-term exposure. The acute oral LD50 for TCE, as determined in several species of animals, is reported by Torkelson et al. (1958) to range from 5.7 to 14.3 g/kg. Unfortunately, little other toxicological data involving oral dosing are available. LD50 values that were derived upon administration of TCE by routes other than oral illustrate the difficulty in using such data to predict consequences of ingestion of the chemical. In contrast with an oral LD50 value of 11 g/kg in the mouse (Torkelson et al., 1958), the LD50 is approximately 16 g/kg for subcutaneous injection (Plea et al., 1958) and approximately 4.9 g/kg for intraperitoneal injection (Klaassen and Plaa, 1966~. By administering equivalent intraperitoneal and oral doses of carbon tetrachloride to rats, Nadeau and Marchand (1973) demonstrated that significantly higher hepatic concentrations of carbon tetrachloride and more extensive hepatotoxicity are manifested in the intraperitoneally dosed animals. Despite the problems inherent in extrapolating data from one route of chemical exposure to another, we may gain qualitative insight into the toxicity of TCE by examining information from studies in which the oral route was not used. Plaa and his colleagues found TCE to be the least hepatotoxic of a series of aLkyl halocarbons that were given subcutane- ously (Plea et al., 1958) and intraperitoneally (Klaassen and Plaa, 1966) to mice and intraperitoneally to dogs (Klaassen and Plaa, 1967) and rats (Klaassen and Plaa, 1969~. Near-lethal quantities of TCE were generally required to produce hepatotoxicity. They observed little to no evidence of nephrotoxicity. In contrast to TCE (ED50 - 2.5 ml/kg for SGPT elevation in mice), its congener 1,1,2-trichloroethane was much more toxic (ED50= 0.1 ml/kg), and tetrachloroethylene was of equivalent potency (ED50 = 2.9 ml/kg). In laboratory animals, as well as in humans, the primary hazard of inhalation of high concentrations of TCE is excessive depression of the central nervous system. Adams et al. (1950) reported the 3-fur LC50 in rats to be 18,000 ppm. They observed that recovery of several test species of animals from marked depression of the central nervous system was rapid and uneventful. The lowest and shortest exposure that elicited histologi- cal change in tissues of rats was 8,000 ppm for 7 hr. This produced an increase in liver weight and fatty vacuolation of hepatocytes. Distur- bance of vestibular function in rabbits infused intravenously with TCE was observed by Larsby et al. (1978) when blood levels of TCE in the rabbits exceeded 75 ppm. Levels of TCE in the cerebrospinal fluid were approximately one-third of that in the blood. Although this vestibular disturbance is physiologically significant, it should be recalled that Gamberale and Hultengren (1973) observed inhibition of psychophysio- logical function in humans with blood levels of only 3 to 5 ppm TCE.

150 DRINKING WATER AND HEALTH A second hazard that is associated with acute exposure to vapor containing high concentrations of TCE is cardiovascular toxicity. The aforementioned accounts of cardiotoxic effects of TCE in humans (Bass, 1970; Dornette and Jones, 1960) have been confirmed in studies of dogs. Reinhardt et al. (1973) found TCE to be more potent than trichloroethyl- ene in inducing arrhythmias in dogs concomitantly dosed with epineph- rine. The lowest elective concentration of TCE was 5,000 ppm. However, Egle et al. (1976) did not detect adverse cardiovascular effects in freely moving dogs that had been exposed to 5,000 and 10,000 ppm TCE in a Freon propellant. They attributed the disparity between their own findings and those of Reinhardt et al. (1973) to differences in experimental design. Herd et al. (1974) found TCE to exert a Aphasic action on the cardiovascular system of anesthetized dogs, which was characterized by an initial decrease in blood pressure that was associated with peripheral vasodilation as well as reflex chronotropic and inotropic ejects on cardiac function, and subsequent depression of cardiac function. In a study of the biochemical mechanism of TCE's cardiotoxic- ity, Herd and Martin (1975) observed inhibition of respiratory function and alteration of permeability characteristics in mitochondria that were isolated from rats. Herd et al. (1974) emphasized that studies are needed to determine whether low-level exposure to TCE may be injurious to the cardiovascular system. In contrast to previous findings of microsomal enzyme induction in mice (Lal and Shah, 1970) and rats (Fuller e! al., 1970) that inhaled 3,000 ppm TCE for 24 hr. inhibition of microsomal drug metabolism was observed in rats that had been given approximately 1.4 g/kg orally (Vainio e! al., 1976) and in mice that had been given 1.0 mI/kg of undiluted TCE intraperitoneally (Shah and Lal, 1976~. Shah and Lal (1976) further demonstrated that dilution of the TCE with olive oil reduced the inhibitory effect, while TCE that was diluted with dimethyl sulfoxide (DMSO) potentiated the effect. These investigators suggested that the olive oil inhibited the systemic absorption of TCE and that the DMSO potentiated TCE's hepatotoxicity. Chronic Effects McNutt et al. (1975) exposed mice continuously to 250 and 1,000 ppm TCE for up to 14 weeks. Serial sacrifices were performed at weekly intervals to ascertain the development of any histopathological abnormalities. Hepatocytic vacuolations and sig- nificant increases in liver weight and triglyceride content were observed throughout the study in the animals exposed to 1,000 ppm. After 4 weeks of exposure to 1,000 ppm TCE a number of ultrastructural alterations were observed in centrilobular hepatocytes, including proliferation of

Toxicity of Selected Drinking Water Contaminants 151 smooth endoplasmic reticulum. Such a structural alteration would be expected in light of the reports of microsomal enzyme induction by Fuller et al. (1970) and Shah and La] (1976~. McNutt et al. (1975) saw a return to normal of each of the indices at 2 and 4 weeks after exposure. Quite modest ultrastructural alterations and increases in liver weight and triglyceride were occasionally observed in the animals that were exposed to 250 ppm during the Tweak study. Thus, this exposure level Night be considered a threshold for a biological effect of TCE in the mouse. Platt and Cockrill (1969) studied biochemi- cal changes in rat livers in response to a series of aliphatic halocarbons. They found seven daily oral doses of 1.65 g/kg to enhance cytoplasmic and microsomal protein content and to exert no hepatotoxicity. Savolainen et al. (1977) reported slight decreases in brain RNA and liver microsomal P-450 in rats inhaling 500 ppm TCE for 6 hr daily for 4 or 5 days. The significance of these latter findings is uncertain. The only lifetime feeding study that has been reported was conducted as a part of the National Cancer Institute (NCI) Bioassay Program (National Cancer Institute, 1977~. In an initial range-finding study, oral doses ranging from 1,000 to 10,000 mg/kg TCE in corn oil were given to male and female mice and rats 5 days weekly for 6 weeks. The highest "no-e~ect" dose for rats was 3,160 mg/kg while that for mice was 5,620 mg/kg. Indices of toxicity that were evaluated included body weight and gross evidence of organ damage. A chronic dosing study was then initiated but had to be discontinued because of undefined intoxication in rats receiving 3,000 mg/kg. In the final chronic dosing study, male and female rats received 750 or 1,500 mg/kg TCE in corn oil by Savage 5 times weekly for 78 weeks. Similarly, male and female mice were given doses that were increased during the study when it became apparent that larger quantities of the chemical could be tolerated. The time-weighted averages for the two dose levels in mice for the 78-week regimen were approximately 2,800 and 5,600 mg/kg. Diminished body weight gain and decreased survival time were manifest in both mice and rats. Surprising- ly, the incidence of histopathological change was no greater for TCE- dosed than for control animals of either species. No other indices of toxicity were evaluated. A number of long-te:~ animal studies of the toxic potential of inhaled TCE have been conducted over the last 20 years; These studies have been directed largely toward assessing potential hazards of TCE in occupational exposure situations. Daily exposure of a variety of species to 500 ppm of TCE over a 6-month period elicited no recognizable adverse effect, but 1,000 ppm produced fatty changes and increased weight of livers of guinea pigs (Torkelson et al., 19581. Rowe et al. (~1963)

152 DRINKING WATER AND HEALTH reported similar findings when testing a solvent mixture consisting of approximately 75% TCE and 25% tetrachloroethylene. However, guinea pigs in the latter study did show some decrease in body weight gain, wh ch was attributed to reduced food consumption, as well as an increase in liver weight. In studies of responses to even lower concentra- tions, Prendergast et al. (1967) exposed rats, guinea pigs, dogs, rabbits, and monkeys to TCE vapor continuously for 90 days. They observed depressed body weight in rabbits and dogs inhaling 370 ppm, but no adverse effects in any species inhaling 135 ppm. Eben and Kimmerle ( 1974) detected no evidence of hepatorenal injury, hematological change, or histopathologic alteration in rats that received 200 ppm TCE for 8 hr daily, 5 days weekly, for 14 weeks. Mutagenicity Simmon et al. (1978), when conducting a mutagenesis screen of 71 chemicals that had been identified in U.S. drinking water, found TCE to be very weakly mutagenic in vitro for Salmonella typhimurium. Microsomal activation appears to have little effect on its potency. Carcinogenicity The only study of the carcinogenic potential of TCE conducted to date failed to reveal any evidence of carcinogenicity (National Cancer Institute, 1977~. Teratogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level fSNARL) Twenty-Four-Hour Exposure The literature indicates that TCE is one of the least toxic of the commonly used alkyl halocarbons. Since depression of the central nervous system is its predominant effect when inhaled, it appears that loss of manual dexterity, coordination, percep- tion, etc., may be the most sensitive indices of exposure. Unfortunately, it is unclear whether significant inhibition of psychophysiological functions will occur in humans who ingest the chemical. The 0.6 g/kg of TCE reportedly ingested by the patient of Stewart and Andrews (1966) might be considered to be a minimum oral hepatorenal toxic dose. However, the nausea, vomiting, and diarrhea experienced by the patient are toxicologically significant manifestations that must be avoided, although the gastrointestinal upset may have resulted from consumption

Toxicity of Selected Drinking Water Contaminants 153 of undiluted TCE. Moreover, the vomiting, diarrhea, and gastric ravage may have prevented systemic absorption of a portion of the TCE. A single case history is obviously not sufficient to serve as a basis for setting an exposure level for acute ingestion of TCE. However, a similar quantity of TCE in laboratory animals appears to be what might be termed a "minimum-e~ect level." Vainio et al. (1976) found that a single oral dose of approximately 1.4 g/kg depresses some hepatic microsomal metabolic indices in rats. In light of reports of hepatic microsomal enzyme induction in mice and rats following inhalation of TCE, it is possible that oral doses lower than 1.4 g/kg might also stimulate xenobiotic metabolism. Nevertheless, it seems appropriate that calcula- tions for a suggested 24-hr SNARL for contamination of drinking water by TCE be based upon a minimum (oral) erect level that is derived from actual experimentation, namely 1.4 g/kg. This SNARL is based upon the assumption that the sole source of TCE during this time will be drinking water and that a 70-kg human consumes 2 liters/day (an uncertainty factor of lOO is applied): . . . 1.4 g/kg x 70 kg 100 x 2 liters = 490 mg/1iter. Seven-Day Exposure TCE appears to be no more hazardous upon short-term exposure than it does upon acute exposure. A study in which humans were subjected to 500 ppm of TCE vapor on 5 consecutive days revealed no evidence of toxicity (Stewart et al., 1969~. Platt and Cockrill (1969) reported seven daily oral doses of 1.65 g/kg not to be hepatotoxic in rats, but to enhance hepatic microsomal and cytoplasmic protein content. However, their use of liquid paraffin as a vehicle may have markedly retarded systemic absorption of the TCE. Thus, because of the lack of more definitive information regarding short-term minimum- or no-effect levels, the suggested 7-day SNARL for drinking water contamination is obtained by dividing the 24-hr SNARL by 7. Assuming that the sole source of TCE during this period will be drinking water: 490 mg/1iter = 70 mg/liter. Definitive studies in several species of animals should be undertaken using a range of oral doses of TCE to characterize dose-effect and dose- response relationships for both single and multiple ingestions over a 1- week period. A variety of tests, which are valid indices of injury to potential target organs (e.g., heart, liver, kidneys), should be monitored. As no data pertaining to the uptake, distribution, metabolism, and

154 DRINKING WATER AND HEALTH excretion of TCE upon ingestion are available, pharmacokinetic studies should be undertaken using a range of oral doses in several animal species. Vehicles for administration should be carefully selected to avoid discrepancy from actual exposures via drinking water or foods. Limited studies of ingestion of small quantities of TCE might be undertaken in humans. These studies appear warranted since the toxic end points that might serve as a basis for setting standards include subjective (e.g., nausea) as well as subtle objective (e.g., performance) indices. It would be valuable to determine the quantity of TCE that must be consumed to produce a blood level of 3 to 4 ppm, the level that Gamberale and Hultengren (1973) associated with inhibition of psycho- physiological function. Other indices that might be evaluated include cardiovascular function and microsomal xenobiotic metabolism. Chronic Exposure TCE seems to be no more toxic upon long-term exposure than it is upon acute or short-term exposure. Quite large quantities of the chemical given orally to mice and rats 5 times weekly for 78 weeks elicited little apparent histopathological change of any organ in either species (National Cancer Institute, 1977~. However, decreased body weight gain, obvious ill health, and diminished survival time in certain of these animals suggest that more sensitive and/or appropriate tests may reveal adverse effects by comparable ingestion regimens. Indeed, McNutt et al. (1975) reported increased liver weight, liver lipid content, and ultrastructural alterations in hepatocytes of mice that had been subjected for weeks to a vapor concentration as low as 250 ppm. Unfortunately, since no information on TCE blood or tissue levels was presented by these investigators, it is difficult to extrapolate their data to oral exposure. In the absence of more definitive information regarding the chronic toxicity of ingested TCE, the lowest dosage level that was administered to either species in the NCI (1977) study (i.e., 750 mg/kg to rats) will be used as a basis for calculating a chronic SNARL. Depression in body weight gain in males and diminished survival time in both males and females have been observed in rats that were maintained on the 750 mg/kg oral dose. The following calculation assumes that a 70-kg human consumes 2 liters of water per day and that 20~o of the total TCE intake is provided by water. An uncertainty factor of 1,000 is used and the dose is multiplied by 5/7 to convert from the 5- to 7-day exposure: 750 mg/kg x 5 days x 0.1 liter x 70 kg 7 days x 1,000 = 3.8 mg/l~ter. The study from which this value was calculated did not provide a no

Toxicity of Selected Drinking Water Contaminants 155 observed-adverse-effect level. The large uncertainty factor Is used for this reason. Because of the expected high use of this compound, the subcommittee considered it important to provide some provisional guidelines. Appropriately designed oral, long-term dosing studies using several species of animals and a range of doses of TCE should be conducted in order to establish minimum toxic dose levels with accuracy. Vehicles for administration should be selected to assure that artificial exposure conditions are not created, e.g., use of large quantities of corn oil, as in the NCI (1977) study. Sensitive indices of aliphatic halocarbon exposure should also be selected carefully. Since only one investigation of the carcinogenic potential of TCE has been reported to date, additional research should be conducted with doses of TCE that do not shorten the life-span of the subjects. Vehicles for TCE dilution/administration should not create highly artificial exposure conditions. Appropriate studies to investigate the mutagenic and teratogenic potential of TCE should also be conducted. Trichloroethylene (CIHC=CC12) This compound was evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 777~. The following material, some of which became available after that 1977 publication, updates and, in some instances, reevaluates information in the earlier report. Also included are some references that were not assessed in the original report. METABOLISM The pharmacokinetics of trichloroethylene (TCE) have been studied extensively in humans and laboratory animals that were exposed to the chemical by inhalation. Unfortunately, there is little information concerning the uptake and distribution of ingested TCE. One would anticipate that it would be readily absorbed from the gastrointestinal tract since it is a small, uncharged lipophilic molecule. Stewart and Dodd (1964) found that TCE penetrated intact human skin. Astrand and Ovrum (1976) reported rapid absorption of TCE vapor from the lung and a retention of approximately 55% of the amount that was inhaled by men. The arterial blood of these subjects was found to contain substantially higher TCE levels than did venous blood, indicating rapid uptake of the chemical from blood into tissues. TCE would be expected to localize in tissues of the body according to their lipid content (Sato et

156 DRINKING WATER AND HEALTH al., 1977), although the liver should accumulate an even greater proportion of ingested TCE because of its position in the portal circulation. Although a portion of systemically absorbed TCE is exhaled un- changed, a substantial amount is metabolized. Ogata et al. (1971) found that 65% to 75~o of the TCE that was retained by humans during an inhalation exposure was excreted eventually as urinary metabolites. The proportion of metabolized TCE versus the proportion that was exhaled unchanged would be expected to vary inversely with dose. The details of TCE metabolism will not be related here since they have been reviewed in detail by the National Institute for Occupational Safety and Health (1973), Uehleke and Poplawski-Tabarelli (1977), and Uehleke et al. (19771. The implications of epoxide formation and reaction/inactivation are discussed below under mutagenicity. Suffice it to say that the major, isolable TCE metabolites in humans and laboratory animals are trichloroethanol glucuronide and trichloroacetic acid (TCA) (Ikeda and Ohtsuji, 19721. TCE is metabolized quite rapidly in humans, since both trichloroethanol and TCA appear in the blood quickly upon exposure to TCE (Muller et al., 1974~. Trichloroethanol is readily eliminated in the urine. The peak urinary excretion occurs from 1 to 3 hr after exposure to TCE (Ogata et al., 1971~. In contrast, renal clearance of TCA is delayed due to its high degree of plasma protein binding. Trichloroethanol is the predominant TCE metabolite in both humans and rodents, in that 3 to 4 times more trichloroethanol than TCA is excreted in the urine (Ikeda and Ohtsuji, 19721. Negligible quantities of both metabolites are eliminated in the feces. Measurement of both urinary metabolites has been used as an index of occupational exposure to TCE, although the procedure is of limited quantitative value due to marked individual variability (Monster et al., 19764. HEALTH ASPECTS Observations in Humans There are a number of accounts of poisoning of individuals who ingested TCE. Clinical signs and symptoms in victims are principally those of gastrointestinal upset, narcosis, and occasional cardiac abnormalities. Hepatorenal involvement is uncommon, although Kleinfeld and Tabershaw (1954) recounted a fatal case in which severe hepatorenal injury resulted from accidental ingestion of an undeter- mined amount of TCE. Two persons who each consumed 15 to 25 ml of TCE experienced vomiting and abdominal pain, followed by inebriation and transient unconsciousness (Stephens, 19451. Less severe depression

Toxicity of Selected Drinking Water Contaminants 157 was observed in a 15-year-old girl who swallowed approximately 15 ml of TCE just after having eaten a large meal (Naish, 1945~. A 4.5-year-old boy who ingested an estimated 7.6 g of TCE became inebriated within a few minutes, but recovered after approximately 3 hr (Gibitz and Plochl, 1973~. More profound depression of the central nervous system, often accompanied by cardiac abnormalities, has been experienced by persons who consume even larger quantities of TCE (Dhuner et al., 1957; Todd, 1954; Tomasini, 1976~. Dhuner and colleagues (1957) reported the case histories of two patients who, upon drinking 350 and 500 ml TCE, were rendered unconscious for 4 and 8 days, respectively. Hypotension and cardiac arrhythmias were delayed in onset, but were quite serious in nature. In a number of instances, cardiac arrhythmias have been observed in patients being anesthetized with TCE and in subjects who were exposed occupationally. Those reports have been reviewed by the National Institute for Occupational Safety and Health (1973~. Experiments involving intentional, acute exposure of humans to TCE vapors reveal that inhalation of low levels of the chemical can result in mucous membrane irritation and impairment of psychophysiological functions. Complaints of eye and throat irritation and fatigue are made by some persons after exposures to 100 to 200 ppm TCE. Objective measures of psychophysiological performance generally show little evidence of inhibition (Gamberale et al., 1976; Nomiyama and Nomiya- ma, 1977; Vernon and Ferguson, 19691. In contrast, Salvini et al. (1971b) reported that an 8-hr, 100-ppm exposure to TCE can inhibit performance in human subjects. Stopps and McLaughlin (1967) noted a slight decline in performance of subjects inhaling 200 ppm. This became increasingly pronounced at the 300- and 500-ppm exposure levels. TCE blood levels are approximately 1 ~g/:T~1 in subjects inhaling 100 ppm (Muller et al., 1974) and 2 ~g/ml in subjects inhaling 200 ppm (Vesterberg and Astrand, 19761. In light of the adverse effects of high concentrations on cardiac function, it is of interest that slight depressions in pulse rate and/or blood pressure have been observed in persons inhaling levels of TCE in doses as low as 170 ppm (Ogata et al., 1971) and 200 ppm (Nomiyama and Nomiyama, 19771. Vernon and Ferguson (1969) reported that inhalation of 1,000 ppm TCE for 2 hr had no erect on blood cell count, BUN, SOOT, or urinalysis. A limited number of inhalation studies, in which subjects were exposed to TCE for 1 to 2 weeks, indicated that the chemical is apparently no more hazardous upon repeated exposure than upon acute exposure. However, TCE does have a potential for accumulation in the body. Ikeda (1977) estimated that its urinary biological half-life in humans is 41 hr. in contrast to values of 144 hr for tetrachloroethylene

158 DRINKING WATER AND HEALTH and 7 hr for both toluene and xylene. Humans that had been exposed to 50, 100, or 200 ppm of TCE vapor for 6 to 7 hr daily on 5 consecutive days exhibited progressively greater blood levels of trichloroethanol (Ertle et al., 1972) and urinary excretion of trichloroethanol and tricholoracetate (Stewart et al., 1970b). Stewart and his colleagues could detect TCE in the exhaled air of one subject for as long as 88 hr after exposure and metabolites in urine for up to 12 days. They also observed a striking degree of individual variability in urinary excretion of TCE metabolites. There was no alteration in clinical chemistry, urinalyses, or hematological indices, nor any evidence of elects on neurological and performance tests. Several epidemiological studies of occupational exposures to TCE have been reviewed by the National Institute for Occupational Safety and Health (19731. There was a variety of subjective complaints and objective clinical findings, the majority of which involved irritant and depressant elects of the compound. Unfortunately, the accounts of chronic erects of TCE suffer from a number of deficiencies including the researchers' failure to determine accurate exposure levels and their inability to distinguish TCE-induced effects from those caused by other factors. In the absence of more definitive chronic exposure data, a time- weighted average (TWA) limit of 100 ppm for occupational TCE exposure was recommended in 1973. Selection of this value was based primarily upon findings in acute studies of inhibition of performance in humans. Largely on the basis of the carcinogenic potential of TCE, the National Institute for Occupational Safety and Health (1978) recom- mended that the TWA limit be reduced to at least 25 ppm, if not lower. It noted that ongoing epidemiological studies are looking for an association between occupational TCE exposure and cancer. Observations in Other Species Acute Effects A number of investigations of the acute toxicity of TCE have been conducted in various species of laboratory animals. Unfortu- nately, none involves oral administration of the chemical, except one in which the acute oral LD50 in rats was 4,920 mg/kg (Registry of Toxic Erects of Chemical Substances, 1975~. Klaassen and Plaa (1967) reported that the EDso for elevation of SGPT levels in mice that were given TCE by intraperitoneal injection was 1.6 ml/kg, while the acute intraperitoneal LD50 was only 2.2 ml/kg. In a subsequent study with dogs, the intrapentoneal EDso for SGPT elevation was 0.57 ml/kg, and the acute intraperitoneal LDso was 1.9 ml/kg (Klaassen and Plaa, 1967~. Studies with rats revealed modest elevations in SOOT and/or histo

Toxicity of Selected Drinking Water Contaminants 159 pathological changes in the livers of animals that were given 0.3 ml/kg by intraperitoneal injection (Cornish et al., 1973) or 477 mg/kg by subcutaneous injection (Wirtschafter and Cronyn, 1964~. TCE did not injure the kidneys of mice (Klaassen and Plaa, 1966) or dogs (Klaassen and Plaa, 1967) at any dosage level. In each of the aforementioned investigations, TCE proved to be one of the least hepatotoxic of a series of alkyl halides. TCE, like most other chlorinated aliphatic hydrocarbons, elicits depression of the central nervous system and inhibits cardiac function in animals upon acute exposure. Investigations of possible adverse elects of exposure to inhaled or injected TCE have been adequately reviewed by the National Institute for Occupational Safety and Health (1973~. However, it may be worthwhile to summarize the findings of a limited number of acute studies in which apparent "minimum-effect levels" of TCE were delineated. Mikiskova and Mikiska (1966) observed that a single intraperitoneal injection of 0.6 ml/kg of TCE caused a loss of muscle tone, depression of reflexes, and slowing of heart rate in guinea pigs. Tricholoroethanol, the principal metabolite of TCE in humans and laboratory animals, had similar effects and was at least 3 times as potent as TCE. On the basis of their studies of the potential of TCE to sensitize a dog's heart to epinephrine, Reinhardt et al. (1973) stated that the minimum effective vapor concentration was about 5,000 ppm. Kylin et al. (1963) reported an accumulation of 1,600 to 3,200 ppm in female mice that had been subjected to TCE inhalation for 4 hr. Lower vapor levels altered performance of rats. Inhalation of as low a concentration as 200 ppm TCE for several hours produced an increase in their spontaneous activity (Grandjean, 1960; Savolainen et al.; 1977~; 400 and 800 ppm inhibited their swimming performance; and 1,600 ppm diminished their motor activity (Grandjean, 19631. Chronic Elects Long-term studies of the toxicity of TCE indicate that the chemical is no more harmful upon repeated exposure than it is upon acute exposure of a variety of species of laboratory animals. One of the first comprehensive investigations of the chronic effects of TCE was conducted by Adams et al. (1951~. They exposed several animal species to various levels of TCE vapor 7 hr/day, 5 days/week, for 6 months. "Maximum no-effect" levels were 400 ppm for monkeys, 200 ppm for rabbits and rats, and 100 ppm for guinea pigs. In a later study, Prendergast et al. (1967) subjected rats, guinea pigs, monkeys, rabbits, and dogs to one of two inhalation regimens: 730 ppm TCE for 5 days/week, 8 hr/day, for 6 weeks, or 35 ppm continuously for 90 dlays. The only evidence of effects of TCE exposure in either regimen were

160 DRINKING WATER AND HEALTH occasional, slight body weight loss or depressed body weight gain. Seifter (1944) reported that 7 to 8 hr/day, 6 days/week exposures of dogs to 750 ppm TCE caused liver injury after 3 weeks. Although they stated that the TCE was 98% pure, these results should be interpreted with caution. In a more recent study Kylin et al. (1965) used a similar dosage schedule for 8 weeks. They found that 1,600 ppm TCE produced a slight accumulation of lipid in livers of mice. Fifty-five ppm TCE, inhaled 8 hr/day, 5 days/week, for 14 weeks, caused an increase in liver weight in rats (Kimmerle and Eben, 1973~. No ejects were noted in other organs, in the blood, in blood glucose levels, or in liver and kidney function. A progressive increase in urinary excretion of trichloroethanol was ob- served in these animals, but there was no evidence of accumulation of TCE, trichloroethanol, or trichloroacetic acid. This phenomenon led Kimmerle and Eben (1973) to speculate that TCE had induced its own metabolism. Goldberg et al. (1964) reported that a daily 200 ppm TCE exposure inhibited avoidance response in certain "sensitive" members of a group of rats, although no progression in severity of the inhibition in these animals was observed during a 2-week regimen of daily, Bohr inhalation exposures. There was a rapid development of tolerance to the depressant ejects of TCE in rats that had been exposed repeatedly to 4,380 ppm of the chemical. The only long-term toxicity studies employing oral administration of TCE were conducted as a part of the NCI (1976) investigation of TCE's carcinogenic potential. In an initial range-finding study, doses varying from 562 to 5,620 mg/kg TCE were given in corn oil to male and female rats and mice 5 days weekly for 6 weeks. Indices of toxicity that were monitored included body weight gain, food consumption, general appearance and behavior, grossly apparent lesions, and mortality. Based upon these criteria, the minimum-effect level appeared to be 1,780 mg/kg/day for rats and 3,160 mg/kg/day for mice. ~ chronic dosing study was then undertaken in which the animals were dosed by Savage 5 times weekly for up to 78 weeks. The time-weighted average doses were as follows: male and female rats, 549 and 1,097 mg/kg; male mice, 1,169 and 2,339 mg/kg; female mice, 869 and 1,739 mg/kg. Dose-dependent reduction in body weight gain, haggard appearance, and decreased survival time were manifest in male and female rats throughout much of the study. The only evidence of adverse effect that was observed in the mice was an increase in mortality in high-dose males. Although no histopathological change was detected in the livers of mice or rats, slight to moderate degenerative alterations of proximal tubular epithelium were apparent in kidneys of low- and high-dose male and female mice

Toxicity of Selected Drinking Water Contaminants 161 and rats. Findings pertaining to tumor induction are discussed below under carcinogenicity. Mutagenicity Recently, several groups of investigators have observed that TCE is mutagenic in in-vitro screening systems. Although Shahin and Von Borstel (1977) reported technical grade TCE to be a potent mutagen for yeast' other workers found the compound to be only weakly mutagenic in bacterial test systems (Greim e' al., 1975; Henschler et al., 1977; Simmon et al., 1978~. The purity of the TCE in each investigation was not established except in the study by Henschler et al. who demonstrated that technical grade TCE contained at least two mutagenic contaminants, epichlorohydrin and 1,2-epoxibutane. In contrast, Hen- schler et al. (1977) noted that purified TCE was very weakly mutagenic. Donahue et al. (1978), employing their microsomally activated Sal- monella typhimurium mutagenicity screening system, found that high pressure liquid chromatography purification resulted in a significant decrease in the mutagenicity of 4 of 11 chemicals tested. They suggested that experiments be conducted to determine whether highly purified TCE was mutagenic in the Ames test. A number of investigators have speculated that hepatotoxic, mutagen- ic, and carcinogenic effects of TCE are related to its conversion to a reactive intermediate by the microsomal mixed-function oxidase system. It is generally accepted that this reactive molecular species is 2,2,2- trichloroepoxide, an Intel immediate product during the metabolism of TCE to its first stable in-vivo metabolite, chloral hydrate. However, Bartsch et al. (1976) failed to detect epoxide formation from TCE by mouse liver microsomes and did not find human liver microsomes to activate TCE to a mutagen in a S. typhimurium test. Greim et al. (1975) demonstrated that the stability/reactivity of epoxides of a series of chlorinated ethylenes was dependent upon their pattern of chlorination. Unsymmetrically substituted ethylenes (e.g., 1,1- dichloroethylene, vinyl chloride, TCE) formed more unstable epoxides than did symmetrically substituted congeners (e.g., 1,2-dichloroethylene, tetrachloroethylene). Greim et al. (1975) observed that the aforemen- tioned unsymmetrically substituted ethylenes were indeed mutagenic to Escherichia colt, while the symmetrically substituted compounds were not. Vinyl chloride was the most potent mutagen. Since chlorinated ethanes are not as predisposed as ethylenes to form epoxide intermediates, they are probably less likely than ethylenes to react directly with biological nucleophiles, thereby producing toxicity, mutagenicity, or carcinogenicity. The recent findings of Uehleke et al. (1977) support this concept. They reported the binding of microsomally

162 DRINKING WATER AND HEALTH activated chlorinated ethylenes to hepatic microsomal P-450. Listed in order of decreasing magnitude of binding, they are: 1,1-dichloroethy- lene, TCE, 1,2-dichloroethylene, and tetrachloroethylene. There was no apparent binding in a series of chlorinated ethanes, while synthetic TCE epoxide was a quite strong binding agent. Binding of ~4C-TCE to mouse and rat hepatic microsomal proteins has been demonstrated both in vitro and in viva (Banerjee and Van Duuren, 1978; Bolt et al., 1977; Uehleke and Poplawski-Tabarelli, 1977; Van Duuren and Bane~ee, 19761. Uehleke and Poplawski-Tabarelli (1977) reported that lesser amounts of radio-labeled TCE were bound to mitochondrial and cytoplasmic proteins in the livers of mice. They also noted that the TCE that was used by Van Duuren and Banerjee (1976) contained substantial amounts of labeled impurities. The binding of contaminants to microsomal proteins in vitro without metabolic activa- tion was shown by Uehleke and Poplawski-Tabarelli (1977~. They observed that the binding of purified TCE was considerably less than that of the original sample. Each of the foregoing reports demonstrated that TCE microsomal protein binding, binding to cytochrome P-450, and in-vitro mutagenicity are dependent upon activation of the TCE by a microsomal mixed- function oxidase (MFO) system. In every instance, a microsomal enzyme-inducing agent was used to stimulate the metabolic capability of the mixed-function oxidase system. Uehleke and Poplawski-Tabarelli (1977) and Van Duuren and Banerjee (1976) noted that pretreatment with phenobarbital produced an approximately twofold increase in hepatic, microsomal covalent binding of TCE in mice. SKF 525-A, a microsomal enzyme inhibitor, blocked the covalent binding. The hepatotoxicity of TCE in rats is markedly potentiated by pretreatment with various microsomal enzyme inducers (Carlson, 1974; Moslen et al., 1977a,b). Moslen and her colleagues (1977b) observed a marked decrease in liver glutathione levels in phenobarbital-pretreated rats, while Van Duuren and Banerjee (1976) found that addition of gluta- thione to their induced in-vitro system inhibited covalent binding of TCE to hepatic microsomal proteins. Thus, it appears that reduced gluta- thione helps stabilize or convert reactive intermediatets) to less reactive, toxic metabolitefs). Additional evidence for the formation of a reactive intermediate (epoxide) was furnished by Van Duuren and BaneI3ee (1976), who fouIld that addition of 3,3,3-trichloropropene oxide, a potent inhibitor of epoxide hydrase, to their ir'-vitro system enhanced covalent binding of TCE. Reported strain, sex, and species differences in susceptibility to injury by TCE appear to be related to inherent differences in hepatic metabolic

Toxicity of Selected Drinking Water Contaminants 163 capability. As reported above, TCE produced liver cancers in male and female B6C3F1 mice, but not in Osborne Mendel rats in the NCI study (1976~. One basis for this species difference may be the relatively low activity of epoxide hydrase in the mouse (Oesch, 1973~. Simmon et al. (1978) reported that liver microsomes from B6C3F1 mice pretreated with polychlorinated biphenyl (PCB) were more effective than microsomes from pretreated Sprague-Dawley rats in activation of TCE to a mutagen in the Ames test. Similarly, Banerjee and Van Duuren (1978) found greater binding of TCE to microsomal proteins of B6C3F1 mice than to proteins of Osborne Mendel rats. Such binding was greater for Sprague- Dawley rats than for Osborne Mendel rats. Covalent binding of TCE to exogenous DNA and to hepatic microsomal protein was more pro- nounced in males than in females of the B6C3F1 hybrids. Female mice proved more resistant than males to the tumorigenic action of TCE in the NCI study (1976~. Carcinogenicity Results of the NCI study (1976) have prompted a second investigation of the carcinogenic potential of TCE. This investi- gation, sponsored by the Manufacturing Chemists Association, involves inhalation exposure of B6C3F1 mice and Charles River rats for 6 hr/day, 5 days/week, to 100, 300, or 600 ppm TCE for up to 2 years. A preliminary report (Manufacturing Chemists Association, 1977) of findings in the animals that were sacrificed after 2 years of TCE exposure or that died during this period indicates that there is a modest increase in incidence of both hepatocellular carcinomas and adenomas in male mice exposed to the highest vapor concentration. Scrutiny of the data reveals what appears to be a concentration-related incidence of tumor induction, although variance from controls is not statistically significant in most cases. No liver tumors have been observed in the rats. The findings of the NCI carcinogenicity study (1976) and ir'-vitro mutagenicity studies have stirred a good deal of controversy surrounding the true carcinogenic potential of TCE. Henschler et al. (1977) performed gas chromatography, mass spectrometry analyses of the same industrial grade TCE that was used by the NCI. They found the sample to contain approximately 0.2% of both epichlorohydrin and 1,2-epoxibutane, as well as smaller amounts of chloroform and carbon tetrachloride. Henschler et al. (1977) suggested that the epichlorohydrin and 1,2- epoxibutane were responsible for the reported carcinogenic effect, in that they found each contaminant to be a potent mutagen in a S. typhimurium screening test. In contrast, purified TCE was a very weak mutagen in this in-vitro system. Donahue et al. (1978) pointed out that most carcinogens are active when given orally to rodents on a daily basis in the mg/kg

164 DRINKING WATER AND HEALTH body weight range and that some of the more potent carcinogens are active in the ,ug/kg range. Although Donahue and his coworkers discounted the role of 1,2-epoxibutane, they believed that it was quite possible that contamination of TCE with epichlorohydrin could have been responsible for the carcinogenic response that was reported by the NCI (1976~. Other questions about the design of the NCI (1976) carcinogenesis bioassay of TCE have been raised. One concerns the daily administra- tion of extraordinarily large quantities of a chemical. In such cases, distribution, metabolism, and excretion of the chemical may be quite different from that which occurs at anticipated environmental exposure levels. In addition, repeated insult to the liver by quantities of the chemical sufficient to kill some hepatocytes will result in a continuing state of cellular regeneration. Cellular proliferation is implicated in chemical carcinogenesis in that partial hepatectomy may predispose rodents to tumor induction by various chemical agents (Cayama et al., 1978~. Other problems with the NCI (1976) study design include the administration of large volumes of oil as a vehicle for TCE and the housing of subjects in the same room as animals receiving a variety of other volatile chemicals. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARL) Twenty-Four-Hour Exposure Although acute exposure to large quan- tities of TCE can result in marked depression of the central nervous system, the chemical apparently has a limited capacity to cause residual tissue injury. The lowest oral dose of TCE that has been reported to produce inebriation is approximately 300 mg/kg body weight. However, in light of the significant degree of depression of the central nervous system, which was observed at this dose level, the "minimum-e~ect level" for inhibition of psychophysiological functions is doubtlessly lower than 300 mg/kg. Thus, it seems that the 300 mg/kg figure derived from human clinical experience is a reasonable dose on which to base calculations for a 24-hr SNARL for contamination of drinking water by TCE. A 100-fold safety factor is used recognizing that 300 mg/kg is not a "minimum-effect level" for ingestion of TCE as a single dose. It is assumed here that the sole source of TCE during the 24-hr period will be drinking water and that a 70-kg human consumes 2 liters/day. These considerations ignore the possibility that TCE may be carcinogenic on short-term exposure:

Toxicity of Selected Drinking Water Contaminants 165 300 mg/liter x 70 kg 100 x 2 liters = 105 mg/liter. Seven-Day Exposure TCE appears to be no more hazardous upon short-term exposure than it is upon acute exposure. Unfortunately, the only short-term studies that have been conducted with humans or with laboratory animals involve inhalation exposure to relatively low concen- trations of the chemical. No adverse effects on performance, neurological function, or clinical chemistry indices are reported in humans subjected several hours daily to as much as 200 ppm TCE for 5 consecutive days. These studies demonstrate that TCE, despite rather extensive metabo- lism, has some propensity for accumulation in the body at the exposure levels used. Nevertheless, in view of TCE's relative lack of toxicity, a 7- day SNARL for drinking water contamination, obtained by dividing the 24-hr SNARL by 7, should protect against adverse effects of the chemical. This standard is based on the assumption that the sole source of TCE during this period will be drinking water. These considerations ignore the possibility that TCE may be carcinogenic on short-term exposure: 105 mg/liter 7 = 15 mg/l~ter. Chronic Exposure TCE appears to be no more toxic (noncarcino- genic) on long-term exposure than it is upon acute or short-term exposure. Unfortunately, every long-term study, with the exception of the NCI (1976) carcinogenesis investigation, involves administration of TCE to laboratory animals by inhalation. The "minimum-effect level" reported in these inhalation studies, 55 ppm, produced increased liver weight in rats that were subjected to TCE vapor 8 hr/day, 5 days/week for 14 weeks. The lowest oral dose used in the NCI study (1976), 549 mg/kg/day, produced renal injury, haggard appearance, and decreased life-span in male and female rats. Thus, due to failure of the NC! study to determine a "no-effect level" and to monitor a sufficiently broad battery of sensitive toxicological indices, it is not possible to establish a '`minimum-effect level" for chronic, noncarcinogenic toxicity. Definitive oral administration studies in several species of animals should be undertaken using a range of doses of TCE to characterize dose-e~ect and dose-response relationships for both a single ingestion and multiple ingestions over a 1-week period. A variety of valid indices of injury to potential target organs should be monitored. Particular

166 DRINKING WATER AND HEALTH attention might be focused on ascertaining the "minimum-effect level" for inhibition of psychomotor performance and cardiac function, since these parameters seem to be among the most sensitive indices of human exposure to TCE. Studies involving potential interactions of TCE with other chemicals and drugs are appropriate, in view of the likelihood of exposure of persons to combinations of chemicals. As no data pertaining to the uptake, distribution, metabolism, and excretion of TCE upon ingestion are apparently available, pharmacokinetic studies should be undertaken. Vehicles for administration should be chosen carefully to avoid discrepancies between test exposures and actual exposures from drinking water or food. Additional long-term studies of toxicity, mutagenesis, and carcinogen- esis should be conducted with purified TCE in order to determine if TCE is a toxicant, mutagen, or carcinogen, and the minimum times and doses that are required to produce adverse effects. A range of doses, including anticipated, potential exposures, should be given orally to at least two species of animals. Trichlorofluoromethane (CCI3F) This compound was previously evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 781~. The following material, some of which became available after that 1977 publication, updates and, in some instances, reevaluates the information in the earlier report. Also included are some references that were not assessed in the original report. METABOLISM Theoretical metabolites of trichlorofluoromethane are dichlorofluoro- methane and tetrachlorodifluoroethane. No evidence of free-radical formation in rats or mice has been shown; nor is there evidence of significant metabolism of trichlorofluoromethane although it has been shown to interact with hepatic cytochrome P-450 (Cox et al., 19721. HEALTH ASPECTS Observations in Humans No available data.

Toxicity of Selected Drinking Water Contaminants 167 Observations in Other Species Acute Effects Unanesthetized beagle dogs inhaled various concentra- tions of Freon 11 for 10 min via face mask. Threshold concentrations in blood associated with cardiac sensitization were 28.6 ,ug/ml (arterial) and 19.7 ,ug/ml (venous) (Azar et al., 1973~. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARLJ There are insufficient oral data from which to determine with accuracy the effects from a large oral exposure or to estimate a threshold dose that might produce adverse effects. The solubility of Freon 1 1 in water is negligible at 21 °C, but is as much as 1 g/liter at 25°C. Thus, a person ingesting 2 liters of water per day after a spill could consume 2 g if the water was warm (close to 27°C). This would translate to approximately 30 mg/kg/day in a 70-kg human. In view of the data reported by Kudo et al. (1971), who gave oral doses to mice for 1 month at levels as high as 220 mg/kg without severe adverse effects, it seems very unlikely that the 30 mg/kg projected above would constitute an oral toxic hazard to humans in the short or long term. Although data for calculating a Bohr or 7-day SNARL are very limited, one can estimate the 24-hr value from the 2.5-g oral dose in rats by Slater (1965) and the 7-day value from the 220 mg/kg maximum tolerated dose for 1 month in mice by Kudo et al. (1971~. Assuming that a 70-kg human consumes 2 liters of water daily, with lOO~o of exposure from water, and using a safety factor of 1,000: Twenty-Four-Hour Exposure: 2,500 mg/kg x 70 kg 1,000 x 2 liters Seven-Day Exposure: 220 mg/kg x 70 kg 1,000 x 2 liters = 88 mg/liter. = 8.0 mg/liter. The paucity of data on this compound particularly by the oral route precludes an accurate risk assessment. The maximum dose, which theoretically could be consumed in warm water at 2 liters/day by a 7~kg

168 DR. N K. NO WATER AN D H "eLTH human, would be about 30 mg/kg/day. The meager data suggest that this is a nontoxic dose. Thus, the acute spill hazard would appear to be quite low. However, it seems very pertinent to compare concentrations in the blood of humans after exposure to known levels with concentrations that are known to produce cardiac sensitization in dogs, e.g., 19.7 g/ml (venous) and 28.0 ~g/ml (arterial). Toluene (C6~5CH3) This compound was evaluated in Drir~kirZg Water and Health (National Academy of Sciences, 1977, p. 770~. The following material, which became available after that 1977 publication, updates and, in some instances, reevaluates the information in the earlier report. Also included are some references that were not assessed in the original report. The biological erects of toluene and other organic solvents have been the subject of several recent reviews (Bruckner and Peterson, 1977; Dean, 1978; Hayden et al., 1977; Savolainen, 1977~. Except for toluene's depressant erects on the central nervous system, there are no conclusive data indicating specific target organ toxicity. However, many biological effects have been attributed to toluene in both humans and animals. Among these are neurotoxicity, kidney and liver damage, cardiac sensitization, and blood dyscrasias. METABOLISM An extensive reservoir for inhaled toluene is provided by body fat (Sato et al., 19741. This is illustrated by the fact that saturation of the liver and brain of mice is not reached despite inhalation of toluene vapor at a concentration of 4,000 ppm for 3 hr (Bruckner and Peterson, 1976~. Following inhalation or oral exposure of rats to radio-labeled toluene, similar amounts of radioactivity were found in several organs and tissues, although fat generally contained levels that were 10-fold higher than any other tissue measured. Radioactivity was rapidly eliminated from all organs and tissues, and only 1% or less of the initial radioactivity was found in tissues other than fat, which contained 3.5% to 5% (Carlsson and Lindqvist, 1977; Pyykko et al., 1977~. In humans, the uptake of inhaled toluene was influenced by body fat. Subjects with the least adipose tissue had the smallest uptake (Carlsson and Lindq~st, 1977~. Egle and Gochberg (1976) noted that retention of inhaled toluene in the respiratory tract was greater than that for benzene. In contrast the extent of percutaneous absorption of benzene and toluene appears to be similar (Wahlberg, 1976~. e

Toxicity of Selected Drinking Water Contaminants 169 The measurement of urinary hippuric acid has been used to estimate occupational exposure to toluene, both qualitatively and quantitatively (Brugnone et al., 1976; Caperos and Fernandez, 1977; Engstrom et al., 1976; Kira, 1977; Ogata et al., 1977; Szadkowski, 1975; Szadkowski et al., 1976~. However, these measurements may be of little quantitative value except for exceptionally high exposures because of the variable excretion of hippuric acid by unexposed individuals (Engstrom e' al., 1976). Toluene exposure may alter the metabolism, disposition, and biologi- cal effects of other agents (Hayden et al., 1977~. Noteworthy is the recent report by Andrews et al. (1977) concerning die effects of toluene on benzene disposition and toxicity. They observed that toluene ameliorates benzene toxicity. HEALTH ASPECTS Observations in Humans A variety of effects in humans include adverse mental changes, such as altered psychomotor performance, irritability, disorientation, and unconsciousness. Additionally, toluene abuse has reportedly been associated with cardiac arrhythmias and with liver and kidney dysfunction (Hayden et al., 1977; Weisenberger, 1977~. Female shoemakers who were exposed occupationally to toluene vapor concen- trations of 60 to 100 ppm had significantly higher urinary concentrations of hippuric acid than did controls and voiced some complaints of abnormal tendon reflexes, reduced grasping power, and decreased agility of the fingers (Matsushita et al., 19754. Mixtures of organic solvents that include toluene have been implicated as the cause of lens changes in car painters (Raipta et al., 1976~. However, the evidence is not sufficient to establish that these effects are due to toluene. In other occupational groups exposed to toluene, no chronic toxic effects could be detected (Rosensteel and Thoburn, 19751. Obserrations ir' Other Species Acute Elects The acute oral toxicity of toluene in 14-day-old rats is significantly greater than in young adult rats (Kimura et al., 1971~. In adult rats, the 4-fur LC50 for inhaled toluene concentrate (toluene, 45.89%; benzene, 0.06%; paraffins, 38.69%; naphthenes, 15.36~o) was 8,800 ppm (35 mg/liter). Rats tolerated 6.8 mg/liter for 4 hr. and dogs, 3.0 mg/liter for 6 hr. without signs of discomfort (Carpenter et al., 1976~. Short-term, high-dosage, oral administration of toluene to rats decreased slightly the apparent liver microsomal activity of some of the

170 DRINKING WATER AND HEALTH enzymes associated with the mixed-function oxidase system (Mungikar and Pawar, 1976a) and inhibited liver microsomal lipid peroxidation (Mungikar and Pawar, 1976b). In contrast, the intraperitoneal adminis- tration of toluene to male mice did not alter liver microsomal N- and O-demethylase activity or the various spectral characteristics of micro- somal cytochrome P-450 (Fabacher and Hodgson, 1977~. Single, large oral dosages of toluene were essentially nonhepatotoxic in guinea pigs. However, at the highest dosage, 1.2 ml/kg, some lipid accumulation in liver was noted (Divincenzo and Krasavage, 1974~. Inhalation of toluene vapor by mice and rats at concentrations of 2,600 to 12,000 ppm for up to 3 hr inhibited "performance" as assessed by a battery of animal performance tests. However, there was no evidence of organ damage, as determined by organ function tests, organ weights, and histopathology (Bruckner and Peterson, 1978~. Rats that were exposed to 1,000 to 4,000 ppm toluene vapor for 4 hr showed changes in their sleep cycle, which were detected by electro- encephalogram (EEG) (Takeuchi and Hisanaga, 1977~. Mabuchi et al. (1974) suggested that similar changes can be observed in the EEG's of patients that have been chronically exposed to organic solvents. Using a continuous flow bioassay system, the LC50 for toluene in freshwater was determined for goldfish. The 24-, 48-, 72-, 96-, and 72~hr LC50 values were 41.6, 27.6, 25.3, 22.8, and 14.6 ppm, respectively (Brenniman et al., 1976~. Chronic Effects Toluene applied daily to the skin of rats 4 hr/day for 4 months at lO g/kg increased plasmic and lymphoid reticular cells in bone marrow and impaired leukopoiesis, especially neutrophil matura- tion. Toluene at 1 g/kg was without effect (Yushkevich and Malysheva, 1975~. Daily intraperitoneal (0.25 to 1.0 ml/leg for 12 days) or subcutaneous (0.25 to 1.0 ml/kg for 3 weeks) dosages of toluene administered to male rats produced a dose-related increase in the number and total area of mitochondria per unit of cytoplasmic area in liver. A dose-related decrease in nuclear volume was also observed (Ungvary et al., 1976~. The specificity of these effects was not dete~ined and their significance is uncerta~n. Rats that were exposed to toluene vapors at 1,000 ppm for 8 hr/day for l week had slightly elevated SGOT and SGPT activities and showed metabolic acidosis (Tahti et al., 1977~. However, male and female rats that were exposed to toluene vapor at concentrations of 0, 30, 100, 300, or 1,000 ppm for 6 hr/day, 5 days/week, for 13 weeks (64 exposures) showed no significant alteration in body weight gain or in hematological

Toxicity of Selected Drinking Water Contaminants 171 or clinical chemistry measurements. Furthe~nore, no gross or histo- pathological alteration in any treated rat could be attributed to toluene (Rhudy et al., 1978~. Exposure of mice to 4,000 ppm toluene vapor, 5 times per week for up to 8 weeks, failed to reveal evidence of injury to the lung, liver, or kidney (Bruckner and Peterson, 1976~. Exposure of male and female Fischer 344 rats to 0, 30, 100, or 300 ppm doses of toluene vapor for 6 hr/day, 5 days/week, for 6 months caused only eye reddening and varying degrees of hair loss in some animals. At high doses, both male and female animals exhibited slightly, but significantly less weight gain than controls for the 6-month period. Hematological, histopathological, and clinical chemistry studies revealed no significant dose-related differences be- tween treated and control animals (Chemical Industry Institute of Toxicology, 1978~. Rats and dogs that inhaled 3.9, 1.9, or 0.95 mg/liter of toluene concentrate (toluene, 45.89%; benzene, 0.06%; paraffins, 38.69%; naph- thenes, 15.36~o) for 13 weeks, 6 hr/day, were not significantly different from their air-exposed controls in any of the criteria of injury that were monitored through hematology, clinical chemistry, and micropathology (Carpenter et al., 1976~. Exposure of mice to toluene vapors at concentrations of 1, 10, 100, or 1,000 ppm for 6 hr/day for 20 days induced changes in the composition of peripheral blood and decreased wheel-turning activity. Bone marrow hypoplasia was noted at 1,000 ppm toluene (Inoue, 19751. There were decreases in wheel-turning activity on the tenth day of exposure to 1 ppm toluene and earlier in the other treatment groups (Horiguchi et al., 1978~. Exposure of eight rats to toluene vapor at a concentration of 4,000 ppm, 2 hr/day, for 60 days appeared to decrease learning as measured by impairment of the acquisition of a differential reinforcement on a 12-s schedule. However, there was no elect of toluene as determined by the extinction of the fixed ratio schedule, memory in the continuous reinforcement schedule, spontaneous activity, or emotionality (Ikeda and Miyake, 19781. Mutagenicity There are no data on the mutagenicity of toluene in bacterial systems (Dean, 19781. Lyapkalo (1973) reported chromosome damage of bone marrow cells in rats that were injected with 1 g/kg of toluene (Lyapkalo, 19739. Studies of workers who had been exposed to toluene for many years failed to show a significant increase in chromosome aberrations (Form et al., 1971~. Chromosome damage induced by inhalation exposure to toluene (160 mg/m3) in bone marrow cells was still present 1 month after terminating exposure in rats

172 DRINKING WATER AND HEALTH (Dobrokhotov and Enikeev, 1977) although blood cell counts reportedly had returned to normal. These studies were also cited by Walker (1976~. Carcinogetzicity A skin carcinoma in one mouse and a skin papilloma in another were observed in a group of 30 mice to whose skin 16 to 20 al of toluene was applied topically twice a week for 72 weeks (Lijinsky and Garcia, 1972~. In contrast, application of toluene to the skin of mice 3 times a week for the lifetime of the mice failed to produce any carcinogenicity that could be attributable to the toluene (Poll, 1963~. The collagen content of the dorsal skin of the mice was reduced by three weekly paintings with toluene for 10 weeks. Compared with toluene as a solvent for 0.3% 20-methylcholanthrene, acetone caused a slower decrease of collagen and a longer latency of tumor development (Mazzucco, 1975~. Teratogenicity Exposure of pregnant rats to toluene vapors at a concentration of 600 mg/m3 for an unspecified period caused embryo toxicity but no teratogenicity (Hudak et al., 1977~. Women that were exposed to toluene and other agents through the occupational use of "organosiliceous" varnishes showed evidence of a fall in red blood cell and thrombocyte indices. There was a high incidence of menstrual disorders and reported effects on embryonic and fetal development such as more frequent fetal asphyxia, a greater number of newborns with low weight, and belated sucking of the maternal breast (Syrovadko, 1977~. CONCLUSIONS AND RECOMMENDATIONS Suggested JVo-Adverse-Response Level (SNARL) Twenty-Four-Hour Exposure Single, large (1.2 ml/kg), oral doses of toluene produced minimal lipid accumulation in the liver of guinea pigs (Divincenzo and Krasavage, 19741. Assuming that this dose represents the largest minimal-effect dose and using an uncertainty factor of 100, the following 24-hr SNARL may be calculated for a 70-kg human consuming 2 liters of water per day with lOO~o of exposure from water during this period: 1,200 mg/kg x 70 kg 100 x 2 liters = 420 mg/liter. The solubility of toluene in water is 470 mg/liter at 16°C.

Toxicity of Selected Drinking Water Contaminants 173 Seven-Day Exposure Daily intraperitoneal 1.0 g/kg doses of toluene administered to rats produced only minor mitochondrial changes of liver cells after 12 days of exposure (Ungvary et al., 1976~. These data suggest that rats can tolerate a dosage of 1.0 g/kg/day for 7 days with minimal effect. Thus, the following 7-day SNARL value may be calculated for a 70-kg human consuming 2 liters of water per day assuming an uncertainty factor of 1,000 and lOO~o of exposure from water during this period: 1,000 mg/kg/day x 70 kg 1,000 x 2 liters = 35 mg/liter. Chronic Exposure Exposure of male and female Fischer-344 rats to approximately 1,130 mg/m3 (300 ppm) toluene vapor 6 hr/day, 5 days/week, for 6 months caused minimal toxicity. Based on these data, a chronic SNARL value is calculated as follows for a 70-kg human consuming 2 liters of water per day. An uncertainty factor of 1,000 is assumed. Because these data refer to inhalation exposure, it is further assumed that a 70-kg human inhales 10 m3/day and that 30% of the inhaled toluene is absorbed from the lung into the blood. This chronic value also assumes that 20~o of the toluene comes from drinking water: 1,130 mg/m3 x 10 m3/day x 0.3 x 0.2 1,000 x 2 liters Uranium (U) = 0.34 mg/liter. Uranium is a silvery-white, heavy metal that occurs ubiquitously in the crust of the earth `(Merck Index, 19761. Natural uranium consists of three isotopes 23su, 235U, and 234U in the relative abundance of 99.27%, 0.72%, and 0.006%, respectively (Chen et al., 1961; Merck Index, 1976~. All are naturally radioactive, being cx-emitters, but on a weight basis the activity of 234U is 17,000-fold and that of 235U is 6-fold greater than that of 238U. Thus, radiotoxicity depends upon whether one is dealing with natural uranium or enriched uranium (Chen et al., 19611. Uranium has valences of 6, 5, 4, and 3, and it forms salts with many anions that also influence the solubility of uranium. 235U is used in atomic and hydrogen bombs. 234U and 235U are used as nuclear fuel in power reactors.

174 DRINKING WATER AND HEALTH METABOLISM The major exposure of the general populace to uranium is through diet and drinking water. Welford and Baird (1967) reported that the dietary intake of uranium in New York, Chicago, and San Francisco is 1.3, 1.4, and 1.3 ,ug/day, respectively. The U.S. Geological Survey Water Supply Paper 1812 (Durfor and Becker, 1964) indicates that the concentration of uranium in drinking water for the 100 largest U.S. cities is <0.1 ,ug/liter. However, in some cities, e.g., Long Beach, California, the level may reach 8.6 ~g/liter (Hamilton, 1972~. The International Commission on Radiological Protection (1959) estimates that the body burden of uranium in a 70-kg human is 20 fig. However, Hamilton (1972) estimated that the body burden ranges from 100 to 125 leg in the United Kingdom, where intake is approximately 1 ,ug/day. Disposition studies in humans were conducted by Bernard (1958) who administered 4- to 5-mg doses of hexavalent and tetravalent uranium intravenously to terminal patients suffering from brain tumors. With both chemical forms, 60% to 70% of the uranium dose was excreted within 24 hr. The hexavalent uranium tended to distribute about equally to the skeleton and kidney. Half-lives of approximately 300 days were calculated for both organs. However, tetravalent uranium was distrib- uted to the bones and liver. Studies by Hamilton (1972) reveal that the greatest amount of uranium in the body is found in the skeleton, but kidney burden was not measured in these studies. In small animals, the principal organ for storage of hexavalent uranium is bone (85~o3. Belova and Lun'kina (1967) studied the accumulation of uranium in the teeth of inhabitants from two sites in the Volga region of Russia. The amount of uranium in the drinking water at these sites varied from 5 x 10-5 g/liter to 3 x 10-7 g/liter. These authors found that accumula- tion of uranium in the human skeleton depends upon the daily amount ingested. In addition, they found that the accumulation of uranium increased up the biological chain, i.e., accumulations were lowest in plants and highest in humans. In other animal studies, McClellan et al. (1962) studied the distribu- tion of 233U in the milk of sheep following intravenous injection of uranium. The milk-to-plasma ratio was <0.1. Sikov and Mahlum (1968) found less than 1% of the 233U dose in the fetuses of rats receiving uranium between days 15 and 19 of gestation.

Toxicity of Selected Drinking Water Contaminants 175 HEALTH ASPECTS There is general agreement that chemical toxicity to uranium is most critical in the kidney (Bernard, 1958; Chen et al., 1961~. However, it should be understood that chemical toxicity of uranium is not restricted to the kidneys, but also involves the cardiovascular, endocrine, hemato- poietic, and immunological systems as well as hepatic dysfunction (Novikov, 1970~. 238Uranium could not be a radiological hazard in humans since the doses necessary to deposit enough uranium in bone to create radiation that is equal to 0.1 ,uCi of radium would be far in excess of uranium doses producing lethality by kidney destruction. With the more active a-emitting uranium isotopes, 233u, 234U? and 235U, the bone becomes the critical organ of concern for radiotoxicity (Chen et al., 1961). Observations ir' Humans Very few recent data deal with any aspect of the toxicity of uranium compounds in humans. In early studies, Luessenhop et al. (1958) found that a dose of 0.1 mg/kg of uranyl nitrate administered intravenously to terminal patients suffering from glioblas- toma multiforme produces nephrotoxic ejects that are characterized by catalasuria, albuminuria, and the appearance of casts in the urine. Neuman (1953) recommended a safe kidney burden of 2 to 3 ~g/g of uranium for chronic exposure. The nephrotoxic intravenous dose of 0.1 mg/kg would yield a kidney burden of about 6 ,ug/g. Novikov and colleagues (Novikov, 1967, 1970; Novikov and Abramova, 1969; Novikov and Yudina, 1970; Novikov et al., 1968) published a series of papers describing studies conducted in humans from two towns (A and B) in the Volga region of Russia. These towns were closely matched for climate, socioeconomics, demography, etc., except for the average concentration of uranium in the drinking water, which was 0.04 to 0.05 mg/liter in town A and 0.002 to 0.004 mg/liter in town B (approximately one-tenth that of town A). In general, the Russian investigators found that the relatively high concentration of uranium (0.0~0.05 mg/liter) in the drinking water of town A did not affect the health of the population. No differences were found in birthrate or death rate, deaths from malignant neoplasms, or incidence of cancer. Studies were also conducted to examine the distribution of uranium in various human tissues from these two populations. The greatest amount of uranium was found in the kidney and the bone; however, there were no significant differences between the two towns. The authors concluded that drinking water containing uranium in concentrations of 0.04 to 0.05 mg/liter did not cause an accumulation of

176 DRINKING WATER AND HEALTH the element in human tissues. Additional studies were conducted examining the eject of this uranium difference on various clinical biochemical and hematological tests. The only difference found was an alteration in the ratio of serum albumin to globulin. The inhabitants of town A had a decrease in albumin but an increase in globulins in comparison with the inhabitants of town B (Novikov et al., 1968~. Observations in Other Species Acute Effects The LD50 of uranium (salt not specified) in male albino rats is 750 mg/kg (route not specified) (Arsentyeva, 1972~. Stefanov and Yurukova (1964) gave a single intravenous injection of uranium nitrate to rabbits of both sexes and examined various clinical biochemistry tests as well as making histological examinations on days 1, 3, 5, 7, 10, 14, and 21 after administration of uranium. Weight losses were observed by day 5. The rabbits weighed only 70~o of their initial weights by day 21. Decreases in hemoglobin and erythrocytes were observed. Nonprotein nitrogen and urea were increased. Histological examination revealed nephropathology and hepatotoxicity. Subacute Effects Arsen'yeva (1972) administered uranium in doses of 0.6, 6.0, and 60 mg/kg to rats for 3 months. He observed alterations in alkaline phosphatase, blood urea, and urinary asparagine transaminase at the two higher doses of uranium. At the 60 mg/kg dose, he also noted an inhibition of weight gain. Chronic Effects Novikov and Yudina (1970) examined the effects of 0.02 to 2.0 mg/kg doses of uranium which were administered per os for 12 months to rabbits and in doses of 0.05 to 60 mg/kg in rats. They observed no changes in urea, creatinine, or serum chloride levels. At the high doses in both groups the metabolism of nucleic acids in kidney and liver were found to be altered. Mutagenicity No available data. Carcinogenicity Archer et al. (1973) studied the mortality of uranium mine workers (662) who were examined from 1950 to 1952 until the end of study in 1967. The total death rate was 104 (195~1967), the same as the expected rate; however, there were excess deaths due to malignant diseases of lymphatic and hematopoietic tissue. Complicating the study, there were other agents present in mills, e.g., vanadium, radium, and thorium, which may have been contributing factors. Data from animal

Toxicity of Selected Drinking Water Contaminants 177 studies suggested that excess malignancies may have resulted from irradiation of the lymph nodes by 232Th. Heuper et al. (1952) studied the metallotoxic reactions following the introduction of a suspension of powdered metallic uranium (25% U in lanolin; 50 mg U/0.05 cm3) into the right femur (first study) or into the pleural cavity (second study) of Osborne Mendel male and female rats of 4 and 6 months of age, respectively. In the first study, the investigators observed renal toxicity and necrosis of muscle tissue during the first 6 months. The major histological findings during the remaining 18 months of the study were tumors (sarcoma) in the tissues surrounding the injection site at the femur. In all, 11 of the 30 surviving rats had tumors. In addition, these tumors underwent metastasis to inguinal, lung, and lymph node sites. In the second study, major pathological findings were kidney damage, local tissue necrosis, and tumors at the site of injection. Teratogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS Suggested No-A dverse-Resporlse Level (SNARL) There is a paucity of recent human data for making a recommendation of a maximal permissible concentration for uranium in drinking water. Until 1962, the International Commission on Radiological Protection (1964) based its calculations of maximum permissible concentrations (MPC) for soluble uranium compounds on their radiation effects without taking into account their chemical toxicity (Novikov, 1970~. There is general agreement that calculations of the MPC should be based upon the chemical toxicity of the element and that the kidney should be regarded as the critical organ. Therefore, the following calculations are based upon exposure to 238U with nephrotoxicity as the critical effect. They do not consider potential effects of radiation. Twer~ty-Four-Hour Exposure This calculation is based upon data of Luessenhop et al. (1958) who observed nephrotoxicity in humans after an intravenous dose of 0.1 mg/kg and assumes that a 70-kg human drinks 2 liters/day with 100% of exposure from water during this period: 0.1 mg/kg x 70 kg 2 liters = 3.5 mg/liter.

178 DRINKING WATER AND HEALTH Seven-Day Exposure This is based upon data from Arsen'yeva (1972) who found no toxicity in rats given uranium at a dose of 0.6 mg/kg for 3 months. This calculation, which uses an uncertainty factor of 100, assumes that a 70-kg human consumes 2 liters/day and that lOO~o of exposure is from water during this period: 0.6 mg/kg xi 70 kg = 0.21 mg/liter. Chronic Exposure A chronic exposure- level will not be calculated because uranium and its compounds are suspected carcinogens. The chemical toxicity of uranium (238U) has received little attention in the United States. Apparently, even less is known about its potential for toxicity in drinking water. If considerable amounts of uranium are found in drinking water, there will be need for specific studies to evaluate the potential chemical toxicity of this element. Xylenes rC6H4(CH3~2] These compounds were evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 787~. The following material, which became available after that 1977 publication, updates and, in some instances, reevaluates information in the earlier report. Also included are some references that were not assessed in the original report. The biological effects of the xylenes (o-, m-, p-xylene) and other organic solvents have been the subject of several recent reviews (Bruckner and Peterson, 1977; Dean, 1978; U.S. Environmental Protec- tion Agency, 1978a). In commercial xylene the m-isomer predominates. Xylenes are ranked 1 3th out of 7,000 chemicals that were surveyed for occupational exposure: more than 4 million workers are believed to be exposed. Release of xylenes into the environment each year is estimated to be nearly 410 million kg (U.S. Environmental Protection Agency, 1978a). However, no reports of damage to the environment have been attributed to xylenes.

Toxicity of Selected Drinking Water Contaminants 179 METABOLISM Bioaccumulation has been predicted for xylenes because of their strong partitioning into the n-octanol phase of the n-octanol/water system (U.S. Environmental Protection Agency, 1978a). However, the rapid oxidation of xylenes to their corresponding polar metabolites seems to preclude bioaccumulation in higher animal systems. Other species such as the eel may show some xylene retention, possibly because of a less active metabolizing system (Ogata and Miake, 1975~. In humans who have been exposed to approximately 0.2 to 0.4 mg/liter xylene isomers (o-, m-, p-xylene) or a 1:1:1 mixture for up to 8 hr. Sedivec and Flek (1976a) determined that the pulmonary retention was 64%, which under the described conditions was independent of dosage or duration of exposure. After exposure, only 5% of the retained xylenes were eliminated in expired air. More than 95% of the retained xylenes were excreted by humans into the urine in the form of methylhippuric acids. Free toluic acids, toluylglucuronic acids, and hydroxytoluic acids were not detected. A small portion of the adminis- tered xylenes was excreted into urine as the corresponding xylenol. Engstrom et al. (1977) found that the percutaneous absorption rate of m-xylene in humans was approximately 2 ,ug/cm2/min through the skin of the hands. Percutaneously absorbed m-xylene was primarily excreted into urine as methylhippuric acid. A small amount of xylene was also detected in expired air. Measurements of urinary methylhippuric acids have been used successfully to ascertain qualitatively and quantitatively the extent of exposure to xylenes (Caperos and Fernandez, 1977; Engstrom et al., 1976; Kira, 1977; Ogata et al., 1977; Sedivec and Flek, 1 976b). When p-xylene was incubated with hepatic or pulmonary mierosomes from adult female rats, Harper et al. (1974) observed that hydroxylation occurred primarily on its one methyl group to for~ p-methylbenzyl alcohol. In the presence of an aldehyde or alcohol dehydrogenase, p- methylbenzoic (toluic) acid is formed. HEALTH ASPECTS Observations in Humans Mixtures of organic solvents, which include xylenes, have been implicated as the cause of lens change in car painters (Raipta et al., 1976~. However, the evidence is not sufficient to establish that these effects are due to xylenes.

180 DRINKING WATER AND HEALTH Observations in Other Species Acute Effects The acute inhalation toxicity of p-xylene was 4,740 ppm (LCso) in adult female rats (Harper et al., 1974~. This compares favorably with the 4-fur LC50 of 6,700 ppm for mixed xylenes (Carpenter et al., 1975~. Exposure of female rats to p-xylene vapors for 4 hr to 1,000 to 2,000 ppm caused dose-related increases in SGPT, SOOT, glucose-6-phosphate dehydrogenase, and glutathione reductase by 24 hr after cessation of exposure. This suggests impairment of liver function (Patel et al., 1976~. These exposures also caused depression of the central nervous system and irritation of mucous membranes. The acute intraperitoneal LDso of p-xylene in female rats was 3.8 mg/kg. Dosages of 0.1 mg/kg/day for 3 days caused moderate fatty infiltration of the liver (Harper et al., 1974~. The intraperitoneal administration of o-, m-, orp-xylene to male mice did not alter liver microsomal N- and O-demethylase activity or various spectral characteristics of microsomal cytochrome P450 (Fabacher and Hodgson, 1977~. Xylenes may cause liver injury in guinea pigs. Increases in serum ornithine-carbamyl transferase activity and liver lipids have been noted following intraperitoneal injection (Divincenzo and Krasavage, 1974~. By the use of a continuous flow bioassay system, the LC50 for xylene in freshwater was determined for goldfish. The 24-, 48-, 72-, 96-, and 72~hr LC50 values were 30.5, 25.1, 20.7, 16.9, and 14.6 ppm, respectively (Brenniman et al., 19764. In a similar experiment the xylene 96-hr LC50 for rainbow trout was 10 mg/liter (Folmar, 1976~. Chronic Effects No new data available. l~utagenicity No available data. Carcinogenicity Xylenes have been selected by the National Cancer Institute for carcinogenicity testing (U.S. Environmental Protection Agency, 1978~. Teratogenicity Xylenes have been reported to cross the human placenta (Dowty et al., 1976~.

Toxicity of Selected Drinking Water Contaminants 181 CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level fSNARLJ Twenty-Four-Hour Exposure Little information is available to calcu- late SNARL values for xylenes in drinking water. Moreover, the insolubility of xylenes in water suggests that it is probably not necessary. However, exposure data in humans can be used to construct an impression of how much exposure to xylenes may be tolerated. Sedivek and Flek (1976a) observed no toxic or untoward ejects on humans during or after inhalation of up to 0.4 mg/liter of a 1:1:1 mixture of a-, m-, and p-xylene for an 8-fur period. During the exposure, 64~o of the xylenes was retained. Thus, the SNARL dose in this period may be calculated for a 70-kg man breathing 3.3 me of air containing xylenes at 200 mg/m3. Because these no-adverse-response data are from humans, an uncertainty factor of 10 is applied. It is assumed that during this period 10037 of exposure is from drinking water, which is consumed at the rate of 2 liters/day: 200 mg/m3 X 3.33 m3 X 0.64 (% retention) 10 X 2 liters = 21 mg/liter. Seven-Day Exposure o-Xylene inhalation exposure of rats and dogs for 6 hr/day, 5 days/week for 13 weeks at a concentration of 3,500 mg/m3 in air was tolerated with no adverse effects (Carpenter et al., 1975~. Using these data to calculate a SNARL value for humans inhaling a comparable dose per day, and assuming lOO~o of exposure from water during this period and 2 liters/day consumption: 3,500 mg/m3 X 10 m3/day X 0.64 1,000 x 2 liters/day = 11.2 mg/liter. Chronic Exposure Because there are no data on long-term exposures to xylenes, a chronic SNARL value cannot be calculated. CHEMICALS SELECTED BY THE CHEMISTRY SUBCOMMITTEE Bromine/Bromide/Bromate (Br2, Br~, BrO3-) Elemental bromine occurs naturally as a diatomic molecule, Br2, with a molecular weight of 159.83. It exists as rhombic crystals or dark-red

182 DRINKING WATER AND HEALTH liquid. Its freezing point is -7.3°C; boiling point, +58.73°C; density of liquid, 3.12 at 20°C; density of vapor, 5.5 (air= 1.0~; and vapor pressure, 175 mm Hg at 21°C. The liquid is heavy, yet mobile, volatilizing readily to a dense red vapor with a strong, disagreeable odor resembling chlorine. (The name bromine is derived from the Greek bromos, which means stench.) The liquid and vapor are extremely reactive and corrosive. Bromine can be produced by chlorine displace- ment by electrolysis of inorganic bromides from brines or seawater. It is used as a laboratory chemical and feedstock for the production of a wide variety of inorganic and organic bromides, including ethylene bromide, which is used as an antiknock additive in gasoline. The solubility of bromine in water at 25°C is 33.6 g/liter (Handbook of Chemistry and Physics, 1960; Sax, 1975~. Nearly all of the bromine exists in water as hydrated Br2 molecules; but in saturated solution at 25° C, there is some disproportionation yielding 0.092 g/liter bromide ion (Br~) and 0.11 g/liter hypobromous acid (HOBr). At basic pH, disproportionation yields hypobromite ion (BrO~), which is unstable above 0° C and undergoes further disproportionation to give the stable bromate ion (BrO3 A. In basic solution above 50° C, the reaction 3Br2 + 60H ~ 5Br~ + BrO3~ + 3H2O goes to completion. Disproportionation of bromate to perbromate does not occur. Therefore, bromate ion is quite stable in solution over a wide pH range (Cotton and Wilkinson, 19664. Inorganic bromides and bromates are highly soluble in water (sodium bromide, 795 g/liter at 0°C; sodium bromate, 275 g/liter at 0°C) (Handbook of Chemistry and Physics, 19604. In the past, bromides have been used as drugs for chronic control of epilepsy and as mild sedatives. They still are found in over- the-counter sedative preparations (Goldstein et al., 1974~. The contami- nation of hay by bromide has resulted from breakdown of the fumigant methyl bromide in soil (Knight and Costner, 1977~. Bromates are powerful oxidizing agents capable of causing spontaneous combustion when mixed with oxidizable materials (Sax, 19751. Bromates have also been used in some home-permanent, cold-wave kits as 3% solutions (Masoud et al., 1973) and are added to bread in small amounts as yeast nutrients in bread (Merck Index, 19769. Bromine has been used as a disinfectant in swimming pools, but its use in drinking water disinfection has not been recommended because of its cumulative neurotoxicity. Concentrations of 2 mg/liter are required for disinfection. Taste threshold in humans for bromine in water ranges

Toxicity of Selected Drinking Water Contaminants 183 from 0.17 to 0.23 mg/liter (as bromide) at pH 5 through 9 (Bryan et al., 1973~. METABOLISM Data on elemental bromine and bromate are lacking. The characteristics of bromide absorption and excretion are relatively well known because of the former extensive use of inorganic bromides in prescription and proprietary sedatives. Pharmacokinetically, bromide ion behaves similarly to chloride. It is readily absorbed via the gastrointestinal tract and is distributed in a volume (Vd) somewhat larger than the extracellular space (Vd in humans= 15 liters). The clearance of bromide by the kidneys is approximately 0.9 liter/day in humans, accounting for most of the excretion of that element. From the clearance and distribution volume, the elimination constant may be calculated to be Ke-0.9/15 = 0.06 days, yielding a theoretical biological half-life of 0.693/0.06, or 12 days (Goldstein et al., 1974~. This agrees exactly with the half-life determined by administering 82Br to humans (Soremark, 1960a) and with the result obtained by Gay (1962~. Soremark (1960b) made the interesting discovery that there is a diurnal variation in bromide excretion in humans and that there is no excretion during sleep. Flinn (1941) measured peak blood levels of 500 mg/liter in humans who had been maintained on therapeutic levels of bromide daily for 4 months. One month after cessation of bromide intake, blood levels had dropped to 18.8 mg/liter. Bromide half-life can be shortened to 3 to 4 days by administration of chloride, due to competitive reabsorption in the renal tubule. In rats, half-life values could be varied from 2.5 to 25 days by altering chloride intake from 10 to 144 mg/day (Ranws and Van Logten, 1975~. In dogs that were fed sodium bromide at the rate of 200 to 400 mg/kg/day, the ratio [Br-1 blood / [Halide! blood [Br ] ingested / [Halide] ingested was constant (1.00 to 1.19) (Rosenblum, 1958~. In a 90-day feeding study in rats, Van Logten et al. (1974) observed that plasma bromide increased during the first 3 weeks and then reached a constant plateau. Total halide in plasma remained constant throughout, as did the ratios of bromide in brain: plasma and kidney: plasma (0.2 and 0.6, respectively) (Van Logten et al., 1974~. In this study, the half-life for bromide elimination was 3 to 5 days. Soremark (1960a) reported that the half-life in mice was 1.5 days. In a study by Knight and Costner (1977), cows fed

184 DRINKING WATER AND HEALTH 43 mg/liter bromide in the diet produced milk containing 10.4 to 20.0 mg/liter bromide. In humans, the half-time for reaching steady-state bromide levels at a constant daily intake is 12 days. The level reached will be approximately 16 times the concentration reached on the first day (Goldstein et al., 1974~. HEALTH ASPECTS Observations in Humans Elemental bromine toxicity is due to its extreme chemical reactivity. The undiluted liquid produces severe burns on contact with the skin. The resulting wounds are painful and slow to heal. Bromine vapor is likewise reactive and corrosive, producing severe irritation to the eyes and upper respiratory tract. Pulmonary irritation can also develop in instances of involuntary exposure to concentrations that would not be voluntarily tolerated. The Hygienic Guide Series (1958) indicates that work can be performed comfortably in air having concentrations of 0.98 to 1.95 mg/m3, but not at 3.90 mg/m3. Bromine is a lacrimator at concentrations below 6.5 mg/m3. Concentrations of 65 mg/m3 are not voluntarily tolerated, 390 mg/m3 is dangerous for short- te~ exposure, and 6,500 mg/m3 is rapidly fatal. The threshold limit value (TLV) for Br2 is 0.7 mg/m3, based on irritation to eyes and upper airways (American Conference of Governmental Industrial Hygienists, 19761. Masoud et al. (1973) estimated that the acute fatal dose of inorganic bromate is approximately 57 mg/kg. Ingestion produces a caustic reaction in contacted tissues, accompanied by nausea, vomiting, and epigastric pain. Death is due to acute renal failure, which is caused by a direct nephrotoxic effect of bromate ion. Like chlorates, high doses of bromates also produce methemoglobinemia (Masoud et al., 1973~. Among the inorganic substances bromine, bromates, and bromides, the bromides have been documented most thoroughly with toxicological information because of their past use in therapeutics. The elective plasma concentration of bromide ion for sedation of humans is approximately 960 mg/liter, which corresponds to a maintenance dose of about 17 mg/kg/day. However, priming doses of 40 to 70 mg/kg/day over a period of several days are commonly given, for example, to control epilepsy. Psychotic symptoms and neurological signs occur above approximately 1,600 mg/liter (Goldstein et al., 1974; Green, 1961~. Flinn (1941) reported no significant adverse effects in humans who were maintained on daily doses of inorganic bromide resulting in plasma levels up to 500 mg/liter for 4 months. This corresponds to a

Toxicity of Selected Drinking Water Contaminants 185 maintenance dose of 6.4 mg/kg/day. The most common signs and symptoms of bromide toxicity (bromism) relate to the nervous system, ranging from neuroses and psychoses through severe ataxia. High doses also produce various nonspecific gastrointestinal disturbances, and approximately 25% of the patients will exhibit a greater or lesser degree of skin rash, which Clay be unrelated to dose (Green, 1961; Trump and Hochberg, 1976~. Normal blood levels of bromide range from 1.5 to 50 mg/liter in unexposed individuals. In forensic medicine, blood levels of at least 1,000 mg/liter are regarded as significant when attempting to attribute cause of death to bromism (McBay, 1973~. Observations in Other Species Acute Elects The rapid germicidal action of the halogens is well known. They have found popular use because of their efficacy and relative lack of toxicity to nontarget organisms (see, for example, Parkas, 19474. The oral LD50 of sodium bromide in rats has been reported to be 3,500 mg/kg (Merck Index, 19761. Sodium bromate has an approximate oral LD50 in rabbits of approximately 250 mg/kg (`Toxic Substances List, 19744. Chronic Effects Dogs that were fed sodium bromide at 100 to 400 mg/kg/day for 6 weeks developed signs of gastrointestinal and nervous system toxicity when serum bromide levels reached 1,800 to 4,000 mg/liter (Rosenblum, 19581. In a 90-day feeding study (Van Logten et al., 1974), male and female rats were given doses of 0, 75, 300, 1200, 4,800, and 19,200 ppm of sodium bromide in the diet. Plasma bromide reached a plateau in each dose group within 3 weeks. In the highest-dose male group and the three highest-dose female groups, thyroid weights were significantly increased. Histopathological examination revealed hyperplasia of the thyroid in these groups. Also in the highest-dose male groups there was a significant elevation of adrenal weight and prostate weight relative to body weight. Histopathological examination of the testis revealed decreased spermatogenesis in the highest-dose group. El'piner et al. (1972) reported giving sodium bromide at 0.001 to 5.0 mg/kg/day to rabbits for 6 months. They found that serum ascorbate decreased but that serum aspartate aminotransferase was unaffected except for an increase at the highest dose. Blood glucose was also unaffected. No significant pathology was observed in brain, liver, kidney, spleen, or intestine. Rats that were given 0.01 to 50 mg/kg/day for 6 months showed hair loss and a decrease in serum ascorbate. Protein- bound iodine increased in rats that received 0.01 to 0.1 mg/kg. In rats on

186 DRINKING WATER AND HEALTH 0.1 mg/kg, there was a shortened latent period in which a conditioned reflex responded to a stimulus. Horses that were fed hay containing 6,800 ppm bromide developed severe neurological signs, including weakness, ataxia, and paralysis. Serum levels reached 2,954 mg/liter in intoxicated animals. In normal horse serum, the level was 78 mg/liter (Knight and Costner, 1977~. Mutagenicity No available data. Carcir~ogenicity No available data. Teratogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARLJ Data on the toxicity of elemental bromine and bromate are insufficient or inappropriate for calculation of acute or chronic SNARL values for drinking water. For bromide, the human data obtained from the use of this substance in therapeutics may be used. TwenO~-Four-Hour Exposure An acute priming dose of 40 mg/kg of bromide (as Br~) may be given in therapeutic regimens without ill eject. Assuming 100% intake from water, 2 liters/day intake of water, and 70- kg body weight: 40 mg/kg x 70 kg - 2 liters = 1,400 mg/liter. Acute overdoses of bromide may be treated specifically by administra- tion of excess chloride. Seven-Day Exposure A maintenance dose of 6.4 mg/kg/day yields a serum bromide concentration of 50 mg/liter, which has been found to have no effect over a 4-month period in humans. Using this value, and assuming 100% of the bromide is from water: 6.4 mg/kg x 70 kg = 224 mg/liter. This value is quite close to one-seventh of the 24-hr limit.

Toxicity of Selected Drinking Water Contaminants 187 Chronic Exposure The most appropriate data for chronic exposure appear to be the concentrations of bromide found in the serum of normal individuals (1.5-50 mg/liter). The midrange figure, 26 mg/liter, corresponds to a daily intake of 26 mg/liter x 0.06 x 15 liter = 23.4 ma, or a daily dose of 0.334 mg/kg. Assuming a 20% intake from water in a chronic situation: 0.334 mg/kg x 70 kg x 0 20 = 2.3 mg/liter. This value corresponds favorably with the 2 mg/liter required for disinfection. In 1969, the World Health Organization recommended an acceptable daily intake (ADI) of 1.0 mg/kg for total inorganic bromide intake. Using this figure and assuming a 20970 intake from water, one can calculate a permissible bromide concentration of 7 mg/liter. The Food and Drug Administration (U.S. Department of Health, Education, and Welfare, 1975) lists tolerances for inorganic bromide in various food- stuffs as a range from 12 to 400 ppm and for fermented malt beverages as 25 ppm. The Russian investigators cited above (El'piner et al., 1972) recommended a maximum bromide concentration of 0.2 mg/liter in drinking water. Their toxic end point was based on presumed neurologi- cal effects in rats as measured by changes in response to a conditional stimulus. The significance of these results is difficult to assess. Bromodichloromethane (CHCI2Br) Although bromodichloromethane is generally considered to be insoluble in water, Dowty et al. (1975) have identified it as a component of New Orleans drinking water. They reported the highest concentration in raw water (source unidentified) to be 11 ,ug/liter and in finished water, 116 ,ug/liter. METABOLISM Smith et al. (1977b) reported that a single oral dose of 20 mg/kg (as l4C) in rats is cleared very rapidly. Only 32 SO is recovered from the gastrointestinal tract and the carcass after 3 hr. and 41% after 6 hr. Most of the compound was recovered from the stomach. Fat contained more than any other tissue. Less than 1% appeared in the urine. The balance of the dose is probably exhaled either as the parent compound or as a metabolite. In monkeys, Smith et al. (1977b) observed that the half-life of a similar dose was 4 to 6 hr and that the peak concentration in blood varied from 1 to 2 mg/liter (as TIC).

188 DRINKING WATER AND HEALTH HEALTH ASPECTS Observations in Humans There are no data on the toxicity of bromo- dichloromethane in humans. Dowty et al. (1975) identified trace amounts in pooled plasma samples from eight persons who drank New Orleans drinking water. Observations ir' Other Species A cute Effects Bowman et al. (1978) reported that the oral LDso values for adult female and male Swiss ICR mice are 900 (range, 811-999) and 450 (range, 326 621) mg/kg, respectively. The difference between males and females is statistically significant (P < 0.05~. Deaths occurred over a period of 1 to 6 days. Administration of 500 mg/kg produced sedation and anesthesia that lasted approximately 4 hr. Postmortem examination indicated fatty infiltration of the liver, pale kidneys, and hemorrhaging of the adrenals. The investigators observed no gross changes in other tissues. Chronic Effects No available data. Mutagenicity Bromodichloromethane was mutagenic in the Sal- monella typhimurium (TA-100 without S-9 mix) assay (Simmon et al., 1978~. Carcinogenicity Intraperitoneal injections given over 8 weeks in a cumulative dose of 2,400 mg/kg did not produce pulmonary tumors in strain A/St mice, which were examined 24 weeks after the first injection under conditions in which bromoform was found to be positive (achiest et al., 1977~. CONCLUSIONS AND RECOMMENDATIONS Since the only acute oral exposure data fall within the LDso range for male mice, the suggested 24-hr no-adverse-response level (SNARL) cannot be calculated. In view of the lack of data on the sublethal and chronic oral toxicity of bromodichloromethane, SNARL values for 7-day and lifetime exposures cannot be calculated.

Toxicity of Selected Drinking Water Contaminants 189 Bromoform (CHBr3) Bromoform was evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 695~. The following material became available since that 1977 publication. In some instances, the new information necessitated reevaluation of the information in the earlier report. Also included are some earlier references that were not assessed in the original report. HEALTH ASPECTS Observations in Species Other Than Humans The oral LD50 values for adult female and male Swiss ICR mice are 1,550 (range, 1,165-2,062) mg/kg and 1,400 (range, 1,205-1,595) mg/kg, respectively. Deaths occurred 1 to 9 days following exposure (Bowman et al., 1978~. Mutagenicity Simmon et al. (1978) reported that bromoform was mutagenic in the Salmonella typhimurium TA-100 microsome assay under conditions in which chloroform was negative. Carcinogenicity Theiss et al. (1977) demonstrated that strain A/St male mice given 23 intraperitoneal injections (3/week) of 48 mg/kg bromoform (a cumulative dose of 1,100 mg/kg) developed a statistically significant increase in the number of pulmonary tumors when examined 24 weeks after the first injection. There was a 75% survival rate in the bromoform-treated group compared with a survival rate of 94% in the control group. Mice that were similarly treated with 100 mg/kg (a cumulative dose of 2,400 mg/kg) did not have an increase in the number of pulmonary tumors. Moreover, all other compounds tested, including chloroform, were negative in this particular bioassay. The data are difficult to interpret because of the lack of dose-response in this study. CONCLUSIONS AND RECOMMENDATIONS In view of the lack of data on sublethal oral toxicity of bromoform, suggested no-adverse-response-level (SNARL) values for 24-hr and 7- day exposures cannot be calculated. Because the chronic exposure data are considered to be equivocal, they are inadequate for calculating a chronic SNARL.

190 DRINKING WATER AND HEALTH Catechol [C6H4~0H)21 Catechol is used industrially as a polymerization inhibitor, antioxidant, component in electrosensitive copy paper, as a photographic developing agent, as an oxidation base in certain hair dye preparations, in the manufacture of rubber, and in the synthesis of pharmaceuticals and pesticides. Catechol also occurs in natural products, such as onions, crude beet sugar, wood tar, and in coal and cigarette smoke (Raff and Ettling, 1966~. In 1970, the consumption of catechol in the United States was approximately 545,000 kg (Anonymous, 1973~. The acute toxicity of catechol has been compared to that of phenol (Deichmann, 1971; Deichmann and Keplinger, 1963~. Recent reviews concerning the toxicity of catechol have also considered the similarity of this compound's response to those produced by closely related agents such as resorcinol and hydroquinone (Flickinger, 1976; International Agency for Research on Cancer, 19771. METABOLISM Catechol is readily absorbed from the gastrointestinal tract as well as through the skin (Flickinger, 1976~. It oxidizes in viva to form the corresponding quinone (Deichmann and Keplinger, 1963) which may be more reactive than the parent compound. Semiquinone radicals may be formed which, in turn, produce superoxide anion and hydroperoxides (Timbrell and Mitchell, 1977~. Catechol and its conjugation products, as well as hexuronic, sulfuric, and other acids, are found in urine (Deichmann and Keplinger, 1963~. Glucuronide and sulfate conjugates have also been detected in the urine of chickens and dogs following an injection of [3H]-catechol into the renal artery (Rennick and Quebbe- mann, 19701. In humans and rats the biological half-life of catechol is comparable to that of phenol (Hirosawa et al., 1976~. HEALTH ASPECTS Observations in Humans Occupational exposures to catechol resulting in serious health erects have not been reported. However, Hirosawa et al. (1976) observed that exposure of workers to a combination of catechol and phenol resulted in irritant effects on the upper respiratory tract. The average exposures were 1.8 ppb in air for catechol and 55.6 ppb for phenol. The investigators reported excursion peaks up to 70 ppb and 260 ppb for catechol and phenol, respectively. The observed effects

Toxicity of Selected Drinking Water Contaminants 191 that resulted from the combination of catechol and phenol could not be attributed to catechol alone (Hirosawa et al., 1976~. Deichmann and Keplinger (1963) reported eczematous dermatitis in humans following skin contact with catechol. They also observed central nervous system convulsions in humans following inhalation exposure to catechol. Observations in Other Species Acute Effects The acute oral LDso of catechol ranges from 0.1 to 0.3 g/kg body weight in various animal species (Deichmann and Keplinger, 1963; Flickinger, 1976~. The lethal intravenous dose for dogs is approximately 0.04 g/kg (Deichmann and Keplinger, 1963~. Flickinger (1976) observed that the single dose LD50 for catechol, when applied to the skin of albino rabbits for 24 hr. is approximately 0.8 g/kg. Catechol is also an eye irritant and a primary irritant of the skin in rabbits (Flickinger, 19761. The acute toxic effects in rats that had inhaled catechol-water aerosols during single l- or 8-fur periods were limited to irritation of extremities and tremors on the first day following exposure to concentrations between 2,000 and 2,800 mg/m3 in chamber air (Flickinger, 1976~. Flickinger observed no effects following 8 hr of exposure to 1,500 mg/m3. Angel and Rogers (1972) anesthetized rats to prevent the reflex withdrawal of their hind legs. When these rats were exposed to 0.38 mmol/kg catechol, they responded to a strong pinch by developing myoclonic convulsions. Chronic Effects Other than the effects noted below for catechol carcinogenicity, no chronic effects are known. Mutagenicity There are no data on the mutagenicity of catechol in bacterial systems (Dean, 19781. However, spindle effects, anaphase fragments, and a variety of metaphase aberrations have been observed in chromosomes of onions (Allium cepa). Carcinoger~icity Catechol demonstrates potent cocarcinogenicity when applied to the skin of mice 3 times a week in combination with benzota~pyrene (Van Duuren and Goldschmidt, 1976; Van Duuren et al., 1973). Van Duuren and Goldschmidt (1976) gave mice 40 ,~`g of catechol simultaneously with 5 fig of benzot~x~pyrene. Catechol more than doubled the number of mice that developed papillomas and

192 DRINKING WATER AND HEALTH carcinomas as compared with the number of control mice that received Only benzota~pyrene. When they tested for tumor-promoting activity of catechol, no tumors developed (Van Duuren and Goldschmidt, 1976~. Cholesterol pellets containing 20C%o catechol, which were implanted into the bladders of mice, may have increased the incidence of bladder carcinomas after 25 weeks (Boyland et al., 1964~. Teratogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS Suggested No-A dverse-Resporlse Level fSNARLJ Twenty-Four-Hour Exposure The data on the acute oral toxicity of catechol are insufficient to allow for the calculation of reliable 24-hr, 7- day, or chronic SNARL values. However, inhalation by rats of 1,500 mg/m3 of a catechol-water aerosol was tolerated without adverse effect (Flickinger, 1976~. If these data are used to calculate a SNARL for acute catechol exposure, the following calculation may be made for a 7~kg human who consumes 2 liters of water in 24 hr. assuming that water contributed 100% of exposure during this period and an uncertainty factor of 1,000: 1,500 mg/m3 X ]0 m /dlay X 0.3 (% retention) = 2.2 mg/liter. No comparable additional data are available to provide for the calculation of 7-day or chronic SNARL values. Other than the cocarcinogenic potential of catechol, no serious toxicity has been reported. Given the long-term experience with catechol in industry, the toxic properties are minimal in comparison with many other agents. High acute doses of catechol may produce effects on the central nervous system (convulsions), and contact with the skin may produce dermatitis. Catechol and its metabolites are readily excreted in urine, and bioaccumulation is not expected. In view of the dearth of data concerning the potential chronic effects of catechol, further experimentation on long-term oral toxicity is recommended. Adequate teratogenicity and mutagenicity studies are also required. In the absence of essential information, no chronic SNARL values for catechol in drinking water can be recommended.

Toxicity of Selected Drinking Water Contaminants 193 Chlorine Dioxide, Chlorite, Chlorate, Chioramines (CI02, CIO2-, CIO3-, NH2CI) The chlorine oxides and the oxygen acids of chlorine and their salts are widely used as oxidizing agents in explosives and for bleaching and sterilizing. The following compounds are of interest as constituents of drinking water as a result of disinfection treatments that are alternatives to or combined with the use of free chlorine: chlorine dioxide; chloric acid, chlorate ion; chlorous acid, chlorite ion; hypochlorous acid, hypochlorite ion; monochloramine; and dichloramine. As disinfecting agents, two classes may be distinguished in order of relative activity (Symons et al., 1977~: (1) powerful oxidants (chlonne dioxide > hypochlorous acid) and (2) weaker compounds (hypochlorite ion > dichloramine > monochloramine). Chlorite is a reduction end product from reactions of chlorine dioxide with materials in water. Chlorite may also persist in water as an excess reagent when chlorine dioxide is generated in the aqueous solution. Alternatives to chlorine disinfection, which are under consideration in order to reduce concentrations of trihalomethane, include chlorine dioxide, ozone plus chlorine, which is followed after a time interval by ammonia to reduce trihalomethane accumulation by the formation of chloramines, and less-than-adequate chlorine, which is followed by chlorine dioxide or chloramines to maintain a residual (Symons et al., 1977~. Chlorine Dioxide Chlorine dioxide is a greenish-yellow gas with an irritating odor, which decomposes readily and with explosive force to chlorine and oxygen. Mixtures of chlorine dioxide in air are potentially explosive at >logo, and react violently with organic materials at even lower concentrations (Hailer and Northgraves, 1955; Paulet and Desbrousses, 19701. For these reasons, it is usually manufactured at the point of use rather than transported. It is soluble in water to 2.9 g/liter. Although volatile and photosensitive. it is quite stable and solutions in pure water can be maintained for months in closed containers (White, 19721. It is used primarily as a bleach for textiles and for pulp in the manufacture of paper. This capacity as an oxidant also led to its application in drinking water treatment to improve quality of taste and odor at Niagara Falls. New York, in 1944 (Symons et al., 19771. It is now used in drinking water treatment for control of phenols. for oxidation of iron and manganese, and for final disinfection prior to distribution. In Europe. it is used more

194 DRINKING WATER AND HEALTH widely, but at lower levels than in the United States (Symons et al., 1977~. Chlorine dioxide for water treatment is generated using sodium chlorite by either of two methods: 1. Chlorine Cl2 + H2O ~ HOCI + HCl HOCI + HC1 + 2NaClO2 ~ 2CIO2 + 2NaC1 + H2O 2. Sodium hypochlorite NaOCI ~ HC1~ NaC1 + HOC1 HOC1 + HC1 + 2NaClO2 ~ 2C102 + 2NaC1 + H2O Chlorine dioxide is formed from chlorite at low pH. Inattention to input ratios or failure to maintain a pH < 3.5 can result in dosing the water with chlorite (White, 1972) The chemistry of chlorine dioxide in water treatment is not well understood since a balance sheet determination of the various ions of chlorine is complicated and since the nature and amount of by-products and end products Carl vary with the amount of chlorine and the amount and type of organic material present (Miltner, l9 /7~. Chlorate is generated by the oxidation of chlorine dioxide in the presence of hypochlorous acid and is produced during the generation of chlorine dioxide (Miltner, 1977~: lIOC1 ~ C1O2 ~ C1O3 + HC1. Ch~k:~ine dioxide does not cause formation of trihalom~ethanes~ does not react with ammonias and does not cause formation of chloramines (Stevens en al, 1976~. It can be forced to disproportionate to chlorate and chlorite but only by raising pH to 11 or 12 (White. 1972~. This is not believed to be an important reaction in the water undergoing treatment (Benarde et al., 1965~:

Toxicity of Selected Drinking Water Contaminants 195 2CIO2 + 2NaOH ~ NaClO3 + NaClO2 + H2O. It is generally accepted that the predominant reaction product of chlorine dioxide in water treatment is chlorite and that chlorate and other ions are produced in minor amounts (White, 1972~. An approxi- mately 50% conversion of chlorine dioxide to chlorite was reported by Miltner (1977) and Noack (1978) who used water containing natural humic acids. Oxidation of manganese and of iron with chlorine dioxide to remove them from drinking water using chlorine dioxide also yields chlorites as the reduction product. HEALTH ASPECTS Observations in Humans Proposed limits of use of chlorine dioxide were based primarily upon assessment of the hazards of residual chlorite. Concerned with possible in-vivo methemoglobin production by chlorite, Musil et al. (1964) recommended that no chlorite reach the distribution point. That recommendation is endorsed by the Norwegian Health Authority (Symons et al., 1977~. In the USSR, Fridlyand and Kagan (1971) established a threshold concentration for chlorine dioxide of 0.45 to 0.40 mg/liter, based upon organoleptic properties in water. Van Haaren (cited in Atkinson and Palin, 1973) noted that in West Germany, the maximum applied dose is limited to 0.3 mg/liter. Others practices suggest a residual of 0.5 mg/liter leaving the treatment plant (Atkinson and Palin, 1973~. The U.S. Environmental Protection Agency has proposed that the amount used in the treatment process not exceed 1 mg/liter of water. Observations in Other Species Chronic Effects A 90-day study examined effects of 20 ppm (2- mg/kg/day) and 200 ppm (9 mg/kg/day) of chlorine dioxide in the drinking water of African green monkeys. Minimal local irritation of the oral mucosa was observed in the highest exposure group of monkeys, which had a lower water consumption than the controls, or in the lowest exposure group. Preliminary reports state that no measurable toxicity to the hematopoietic system or to other systems was observed (U.S. Environmental Protection Agency, 1977b,c). In a 2-year study, rats exposed to chlorine dioxide in their drinking water showed no adverse ejects after consuming I.1 mg/kg/day, but those drinking 11 mg/kg/day showed a higher mortality by the end of the study (Haag, 19491.

196 DRINKING WATER AND HEALTH Mutagenicity No available data. Teratogenicity No available data. Carcinogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS Suggested No-A dverse-Resporlse Level (SNARLJ Chronic Exposure Details of the African green monkey experiment were not available at the time of this writing, but the preliminary report stated that maintenance for 90 days on 200 ppm, 9 mg/kg/day, gave no adverse effect other than irritation of oral mucosa. For rats, the highest no-adverse-effect level in the 2-year study of chronic toxicity was 1.1 mg/kg/day. Calculating a SNARL for a 70-kg human, assuming the uncertainty factor = 100 and that lOO~o of daily exposure during this period is from 2 liters of drinking water: 1.1 mg/kg/day x 70 kg 100 x 2 liters = 0.38 mg/liter. As a disinfectant, the residual level of this oxidant depends upon the reducing capacity of organic matter in the water. Although a better understanding of chlorine dioxide and its products in drinking water is necessary, the typical result under desired conditions would appear to be a minimal residual of chlorine dioxide, with approximately 50~o of the initial oxide persisting as chlorite ion. The current recommendation that no more than 1 mg/liter of chlorine dioxide be added during treatment of drinking water is consistent with the toxicity data available for chlorine dioxide. Short-Term Exposure to Higher Levels of Chlorine Dioxide Exposure to appreciably higher levels would be self-limiting in almost all cases, since chlorine dioxide is typically not transported but is generated at the point of use, is volatile in water, and, apparently, has a taste and odor threshold of 0.4 to 0.45 mg/liter, which is only slightly above the calculated SNARL for chronic exposure.

Toxicity of Selected Drinking Water Contaminants 197 Chlorite HEALTH ASPECTS Observations in Humans No available data. Obserratior~s in Other Species Acute Effects Most studies of acute toxicity have used commercial sodium chlorite at 78% to 83% in aqueous solution. Oral LDso values reported for rats were 182 mg/l~g (Sperling, 1959) and 140 mg/kg (Musil et al., 1964~. Musil et al. (1964) found that the pattern of death of rats in the acute toxicity study was characterized by slowed breathing, fee- bleness, and death without spasms. On autopsy, the lungs were filled with hemorrhages,.the heart was filled with blood, and the liver and kidneys were normal. In seven rats, which were treated per os with 320 mg/kg (over twofold the LD50) and then killed and exsanguinated within 10 min of treatment, an average of 57.3% of their total hemoglobin was in the form of methemoglobin (Musil et al., 1964~. Chronic Effects Oxidative damage of red blood cells is abundantly evident from studies using lower exposure levels. The cat is recognized as being highly susceptible to the induction of methemoglobin formation by aromatic amines and by nitrite (Spicer, 1950~. This effect of chlorite on hemoglobin was examined in a series of in-vitro and in-vivo studies by Heffernan and Guion (1978a,b). Washed red blood cells of cats and of other mammals, including humans, became oxidized to contain methe- moglobin in a dose-dependent response to chlorite. The most sensitive adverse response from chlorite in drinking water was a 20~o decrease in the half-life of red blood cells of cats that had received 3 mg/kg/day for 30 days. Mutagenicity No relevant data were available. Tetratogenicity No available data. Carcinogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARL)

198 DRINKING WATER AND HEALTH Twenty-Four-Hour Exposure No acute toxicity data other than an LD50 were available. Seven-Day Exposure The generation of chlorite is believed to occur in approximately 50% yield following treatment of water with chlorine dioxide. With inefficient utilization of sodium chlorite in the initial generation of chlorine dioxide, a chlorine dioxide treatment that is intended to be 1 mg/liter could result in chlorite concentrations of nearly 1 mg/liter. The most sensitive adverse response reported for chlorite in drinking water was a decrease in half-life of red blood cells in the cat, for which 0.6 mg/kg/day was the highest no-adverse-e~ect dose. For a 70-kg human, assuming an uncertainty factor of 100 and lOO~o of the exposure coming from drinking 2 liters of water daily during this period, a calculated no-adverse-e~ect dose is: 0.6 mg/kg x 70 kg = 0.21 mg/liter Implications of this eject of chlorite and this calculated level are discussed in the Recommendations section below. Chlorate Chlorate ion is a strong oxidizing agent. It can be formed readily from hypochlorite ion in weakly alkaline or acid conditions by the reaction: 3 HC1O ~ 2 HC1 + HCIO3. As sodium chlorate, it has been used widely as a nonselective herbicide and in leather, paper, and textile processing. Potassium chlorate was recommended for ulcerative stomatitis in humans for over 100 years, but it was dropped from the National Formular~r in 1960 because of its lack of efficacy and potential toxicity. METABOLISM Chlorate ion per os in aqueous solution is known to be rapidly eliminated in the urine. In a study of seven female dogs, Ross (1925) gave 500 mg/kg doses of chlorate in 500 ml of water. Excretion was at its highest rate in the second 2-fur period. From 55% to 70~o was excreted in the first

Toxicity of Selected Drinking Water Contaminants 199 6 hr. By 24 to 48 hr. 76% to 99% had been excreted unchanged in the urine. The amount of chlorate in the blood peaked at 2 hr. It ranged from 5 to 81 mg/100 ml in five dogs, and it decreased to little or none by 24 hr. None of the dogs showed adverse effects. HEALTH ASPECTS Obserralions in Humans There are clinical descriptions of poisoning from accidental or suicidal ingestion of sodium chlorate weed killer, overdose of potassium chlorate, or consumption of potassium or sodium chlorate when the respective chloride was intended (Lee et al., 1970~. Common poisonous effects are the rapid oxidative destruction of red blood cells, possibly followed or preceded by increased me/hemoglobin, and eventual cyanosis and progressive kidney failure. From these cases, the lethal dose in adults is estimated to be 20 to 35 g for sodium chlorate (Jackson et al., 1961) and 5 to 30 g for potassium chlorate (Rosenblatt et al., 1975~. For a 70-kg human, the oral lethal dose for these salts is 71 to 500 mg/kg. Human blood containing 0.25% potassium chlorate and standing at room temperature showed no evidence of methemoglobin after 8 hr and showed complete change to methemoglobin with hemolysis in 24 hr. Blood with the same chlorate concentration was converted to methemo- globin with hemolysis if left standing 5 hr at 39°C. Methemoglobin formation was also accelerated if blood cells were first hemolyzed in distilled water or if the blood was stored saturated with carbon dioxide (Richardson, 1937~. The taste threshold for chlorates, 20 ppm, is the recommended permissible level for reservoir waters in the USSR (Rosenblatt et al., 1975). Observations in Other Species Acute Elects Lipschitz (1932) studied acute toxicity of sodium chlorate in water given gastrically to cats. Single doses of 1.35 g/kg or greater administered in 60- to 80-ml solutions resulted in typical poisoning symptoms (cyanosis, dyspnea, circulatory depression, me- themoglobinemia and death) in 5 hr and in less time with increasing dose. Doses of 450 mg/kg or less were without effect. Chronic Elects Richardson (1937) gave seven cats intramuscular injections of potassium or sodium chlorate in 5% or logo solution. Daily

200 DRINKING WATER AND H"LTH doses of 0.05 to 0.25 g/kg were administered for 25 to 32 days. None of these cats had demonstrable me/hemoglobin. But, upon histological examination, wildcats receiving greater than 0.05 g/kg showed fibrosis and atrophy of distal renal tubules. Rats maintained for 6 months on oral doses of magnesium chlorate at 3 g/kg/day produced 61 offspring which were examined 1 day after birth (Rosenblatt et al., 1975~. The progeny showed pulmonary and cerebral edema and focal hemorrhage in lungs. Mutagenicity No available data. Teratogenicity No available data. Carcinogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS Under usual circumstances of chlorine dioxide treatment of drinking water, chlorate is expected to be a minor end product, and the risk of exposure to chlorite in the drinking water would be expected to be greater than the risk that is related to the chlorate generated as an end product from chlorine dioxide treatment. There are insufficient data concerning the eject of dilution on the toxic ejects of high doses of chlorate. The lowest published lethal concentration (LDLo) of concentrated solutions in humans is 71 mg/kg. Chloramines Chloramines are generated by the combination of ammonia and hypochlorous acid. The reaction: HOC1 + NH3 ~ NH2CI + H2O yields monochloramine. Further substitution generates dichloramine and nitrogen bichloride. For water treatment, monochloramine (plus some dichloramine) is generated on site by combining chlorine to ammonia at a weight ratio of 3: 1 (White, 1972~. Nitrogen bichloride is avoided since its presence may contribute to taste and odor problems in the water. The presence of chloramines in drinking water is a result of (a) chlorination of water with reaction of the natural free ammonia; (b) addition of ammonia followed by chlorine to decrease the production of trichloro

Toxicity of Selected Drinking Water Contaminants 201 phenol and other substances with objectionable taste and odor; (c) chlorination followed after a time by ammonia; and (d) generation of chloramine in finished water by the addition of chlorine and ammonia. Chloramines are termed "combined chlorine." In cases (c) and (d) they are generated to provide a residual disinfectant during distribution of drinking water. Use of combined chlorine (NH2C1) rather than "free chlorine" (HOCl) limits the further production of trihalomethanes in the water distribution system (Hubbs et al., 1978; Symons et al., 1977~. According to White (1972), chloramines are most widely used in water treatment for the purpose of maintaining a chlorine residual, although they have also been used for primary disinfection. Chloramines are weak disinfectants in comparison to ozone, chlorine dioxide, and chlorine (Symons et al., 19779. Some of the disinfectant capacity of chloramines may be due to the hypochlorous acid that is formed at equilibrium by the reversible synthetic reaction (White, 1972~. HEALTH ASPECTS Observations irz Humans Eaton et al. (1973) demonstrated that chlora- mine in dialyzing fluid causes acute hemolytic anemia, which is characterized by oxidation of red blood cells to form Heinz bodies and increased concentration of methemoglobin in uremic patients undergo- ing hemodialysis. In the characterization study, red blood cells (RBC) showed a dose-dependent increase in relative methemoglobin content upon incubation in water containing chloramines. The RBC also showed an inhibition of hexose monophosphate metabolism (HMP shunt). Gubar and Kozlova (1966) proposed a permissible residual concentra- tion of chloramine in drinking water to be 0.5 to 1.0 mg/liter, on the basis of organoleptic properties. Mutagenicity In studies by Shih and Lederberg (1976) monochlora- mine was a weak mutagen for converting a tryptophan auxotroph trpC to tryptophan independence in addition to being a potent bactericidal agent for Bacillus subtilis. Wlodkowski and Rosenkranz (1975) showed that sodium hypochlorite was a weak base exchange mutagen in Salmonella typhimurium strains TA- 1530, TA- 1536, and TA- 1535. Whether these bactericidal or mutagenic effects are due to a common ionic or nonionic form of these test agents or of their reaction products in the culture media was not investigated. (Chloramines of water, mono-, and dichloramine are not to be confused with "chloramine" which is nitrogen mustard, 2,2'-dichloro-N-methyl diethylamine.)

202 DRINKING WATER AND H"LTH Carcinogenicity No available data. Teratogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARL) There are insufficient data for an estimate of a SNARL in humans for either acute or chronic exposure. Chloramino Acids In addition to reacting with natural ammonia in the water, hypochlorous acid reacts with amino acids and other amines to form organic haloamines. Oxidative deamination of organic amino compounds is also expected to occur as a result of chlorination (Helz et al., 1978~. HEALTH ASPECTS No available data for humans or other species. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SHARLJ There are insufficient data for an estimate of a SNARL in humans for either acute or chronic exposure. Recommendations Additional studies are needed on the chemistry of chlorine dioxide and chloramines in reactions with organic material of drinking water. Maintenance of the integrity of the red blood cell and of its hemoglobin is promoted by several enzyme systems in the cell. The tendency of either chlorite, chlorate, or chloramines to cause oxidation of hemoglobin to Heinz bodies or to methemoglobin would be opposed by the reductive capacity of glutathione and by increased production of NADPH from glucose utilization via the hexose monophosphate shunt. The first step in this latter pathway requires glucose-6-phosphate dehydrogenase (G-6-PDH). Humans deficient in or possessing variant forms of this enzyme are known to be at greater risk to drug-induced methemoglobinemia.

Toxicity of Selected Drinking Water Contaminants 203 Considerable data are available on the sensitivity of the very young to oxidative damage of red blood cell constituents and on the genetic types and distribution of glucose-6-phosphate dehydrogenase deficiency among humans of different racial and geographic origin (Goldstein et al., 1974~. Other genetic traits, such as idiopathic methemoglobinemia, are also known as is the possibility of synergistic responses from various drugs leading to oxidative changes in the red blood cells. Studies of the susceptibility of these human populations to damage of the hemato- poietic system by chlorate and, particularly, by chlorite and chloramines are needed. Chloroform (CHCI3) This compound was evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 7131. The following material, which became available after that 1977 publication, updates and, in some instances, reevaluates the information in the earlier report. Also included are some references that were not assessed in the original report. HEALTH ASPECTS Observations in Humans Oral doses of 30 ml (44.6 g) and 100 ml (148.3 g) produce severe nonfatal poisonings in humans. Ingestion of 200 ml (296.6 g) was fatal to human adults (Van Oettingen, 19641. Dowty et al. ( 1975 ~ identified chloroform in pooled plasma samples from eight persons who drank New Orleans drinking water. Observations in Other Species Bowman et al. (1978) reported that the oral LD50's for adult female and male Swiss ICR mice are 1,400 mg/kg (range, 1, 12~1,680) and 1,120 mg/kg (range, 789-1,590) of body weight, respectively. Death occurred 1 to 9 days following exposure. Hill (1977) provided some evidence that genetic factors can influence the sensitivity of mice to the lethal ejects of chloroform. However, there were no genotypic differences in the hepatotoxic effects although renal toxicity appeared to be related to gene the. - 0 or Chronic Elects Miklashevskii et al. (1966) exposed male guinea pigs and male albino rats to 0.4 mg/kg oral doses and other groups of the same two species to 2% of their respective oral LD50's, which amounted to 35 mg/kg for guinea pigs and 125 mg/kg for rats. The experiment ran for 5 months. Daily administration was implied but not specified.

204 DRINKING WATER AND H"LTH Neither species exposed to 0.4 mg/kg doses showed any changes in conditioned reflexes, autonomic or cardiac activity, blood protein ratios, catalase concentrations, or phagocytic capacity. In the guinea pigs given 35 mg/kg doses, the blood albumin-globulin ratio decreased and the blood catalase activity decreased by the second month of exposure. Guinea pigs that died at this exposure had fatty infiltration, necrosis, cirrhosis of the liver parenchyma, lipoid degeneration, and proliferation of interstitial cells in the myocardium. The rats that received the 125 mg/kg doses had an impaired ability to develop new conditioned reflexes during the fourth and fifth months. Mutagenicity Chloroform was negative in the Salmonella typhimuri- um TA-100 microsome assay under conditions in which other trihalo- methanes were positive (Simmon et al., 1978~. Carcinogenicity Male strain A/St mice given 24 intraperitoneal injections of 200 mg/kg over 8 weeks (a cumulative dose of 4,800 mg/kg) did not develop pulmonary tumors under conditions in which bromo- form was positive (Theiss et al., 1977~. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARLJ The following calcula- tions are for noncarcinogenic effects only. Twenty-Four-Hour Exposure Human data indicate that the lowest dose producing toxic symptoms is 30 ml (44.5 g). Assuming a 1,000-fold uncertainty factor, a 2 liters/day water intake, and that lOO57o of the exposure is from drinking water, the SNARL is calculated as follows: 44.5 gx 1 1,000 x 2 liters = 22 mg/liter. Sever-Day Exposure Assuming that repeated daily intake of chloro- form produces cumulative effects, the SNARL value for 7-day exposures should be one-seventh of the 24-hr exposure limit. This is 3.2 mg/liter. Lifetime Exposure This value cannot be calculated because chloro ~ . . . . Iorm IS a carcinogen in anlma s.

Toxicity of Selected Drinking Water Contaminants 205 Dibromochloromethane (CHBr2CI) Although it is generally considered to be insoluble in water, dibromo- chloromethane has been identified as a component of drinking water in New Orleans (Dowty et al., 1975~. HEALTH ASPECI S Observations in Humans There are no data on the toxicity of dibromo- chloromethane in humans. Trace amounts have been identified in pooled plasma samples that were taken from eight persons who drank water from New Orleans supplies (Dowty et al., 1975~. Observations in Other Species Acute Effects The oral LD50's for female and male Swiss ICR mice are 1,200 (945-1524) mg/kg and 800 (667-960) mg/kg, respectively. The difference between the two groups is statistically significant (P < 0.05~. Deaths occurred from 1 to 5 days after exposure. Oral administration of 500 mg/kg produced sedation and anesthesia that lasted approximately 4 hr. Upon postmortem examination, Bowman et al. (1978) found fatty infiltration of the liver, pale kidneys, and hemorrhaging of the adrenals, but they observed no gross changes in other tissues. Chronic Elects No available data. Mutagenicity Dibromochloromethane is mutagenic in the Salmonella typhimurium TA-100 strain assay (Simmon e' al., 1978~. Teratogenicity No available data. Carcinogenicity No available data. CONCLUSIONS AND COMMENDATIONS Suggested No-Adverse-Response Level (SNARL) Twenty-Four-Hour Exposure Animal data indicate that 500 mg/kg produces sedation and anesthesia Using this dose and an uncertainty factor of 1,000, and assuming a 70-kg body weight and a daily water consumption of 2 liters/day, the following SNARL can be calculated:

206 DRINKING WATER AND H"LTH 500 mg/kg x 70 kg 1,000 x 2 liters = 18 mg/liter. In view -of the lack of data on the sublethal oral toxicity of dibromochloromethane, SNARL values for 7-day and lifetime exposures cannot be calculated. 2,4-Dichlorophenol (C6H3OHOl2) This compound was evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 725~. The following material, which became available after that 1977 publication, updates, and, in some instances, reevaluates the information in the earlier report. Also included are some references that were not assessed in the original report. HEALTH ASPECTS Acute oral LD50 values for 2,4-dichlorophenol have been reported for rats at 580 mg/kg (Christensen et al., 1976), the male rat at 2,830 mg/kg (Vernot et al., 1977), and the mouse at 1,600 mg/kg (Christensen et al., 1976) and 1,630 mg/kg (Vernot et al., 19774. Acute intraperitoneal and subcutaneous LD50 values for the rat have been reported to be 430 and 1,730 mg/kg (Christensen, 19764. Extensive toxicity studies on 2,4-dichlorophenol have been undertaken by Kobayashi et al. (1972~. Acute oral LD50 values were 1,600 mg/kg in ICR (Institute for Cancer Research) mice and approximately 4,000 mg/kg in Sprague-Dawley rats. The investigators found no appreciable differences between males and females. Moreover, they observed no remarkable symptoms of intoxication except for depression of motor activity. In a 6-month feeding study in which male mice were fed diets containing 2,4-dichlorophenol ad libitum (O. 45, 100, and 230 mg/kg/day), Kobayashi et al. (1972) observed no adverse changes in growth rate, hematology, serum SGOT and SGPT, or behavior up to the maximum dosage level of 230 mg/kg/day. They did find slight abnormalities in liver histopathology in animals receiving the highest dosage. They regarded 100 mg/kg/day as the maximum no-effect level. The authors concluded that 2,4-dichlorophenol was a relatively safe substance.

Toxicity of Selected Drinking Water Contaminants 207 CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARL) Chronic Exposure Little information is available on the toxicity of 2,4-dichlorophenol. In the absence of other, more definitive data, lifetime no-effect levels may be estimated from the study of Kobayashi et al. (1972~. The highest no-adverse-effect dose in this study with mice was 100 mg/kg/day. Using an uncertainty factor of 1,000, the following calculations may be made, assuming that 20% of this compound will be due to drinking water and that a 70-kg human consumes 2 liters of water per day: 100 mg/kg/daY 2 li70 kg x 02 = 0.7 mg/liter. This yields a SNARL in drinking water of 0.7 mg/liter for a lifetime exposure to 2,4-dichlorophenol. This value should be considered as tentative pending completion of longer term animal feeding experiments. There are insufficient data for estimating a 1- or 7-day SNARL. Long-term chronic toxicity and carcinogenicity studies are needed for 2,4-dichlorophenol in at least two species along with studies on the mutagenic, teratogenic, and carcinogenic properties of the compound in order to verify the current SNARL for drinking water. Glyoxylic Acid, Glyoxal, and Methyl Glyoxal (OHCCOOH, OHCCHO, CH3COCHO) Glyoxylic acid, glyoxal, and methyl glyoxal are of interest as potential contaminants of drinking water because of their apparent, or possible, formation from a variety of organic precursors during ozonization of drinking water. There is no information indicating potential point sources for contamination of drinking water supplies by these sub- stances. Therefore, the potential for accidental contamination at high concentrations is extremely remote. METABOLISM The mammalian metabolism of glyoxylic acid has received considerable attention, largely in relation to the formation of oxalic acid from endogenous or exogenous precursors. The 24-hr urinary excretion of glyoxylate by 12 normal human adults averaged 3.9 mg (2.2~.0 mg).

208 DRINKING WATER AND H"LTH Since studies of the metabolic fate of infused glyoxylate had indicated that only a small proportion was excreted unchanged, Hockaday et al. (1964) estimated that humans normally synthesized from 150 to 600 mg of glyoxylate daily. The source of endogenous glyoxylate is uncertain, but oxidative deamination or transamination of glycine was thought to be the major source (Hockaday et al., 1964; Williams and Smith, 1968~. Exogenous precursors of glyoxylate in mammals include ethylene glycol. McChesney et al. (1972) summarized the central role of glyoxylate in the metabolism of ethylene glycol, which is converted first to glycol aldehyde, then to glycolic acid and, finally, to glyoxylic acid. These investigators observed that female rhesus monkeys excreted largely unchanged glyoxylate following a large oral dose of 500 mg/kg, but after a single dose of 60 mg/kg, oxalate was the major urinary metabolite (14% of the dose). In addition to direct oxidation of glyoxylic acid to oxalic acid, at least seven other metabolic pathways of glyoxylate have been identified in mammalian systems (Williams and Smith, 19681. These include trans- aminations with L-glutamate, L-ornithine, or other amino acids to yield glycine and the corresponding keto acids; reduction to glycolate; reaction with c~-ketoglutarate or with pyruvate to form a-hydroxy-,6- ketoadipate or 2-keto-4-hydroxyglutarate; and FMN (flavin mono- nucleotide) catalyzed condensation with acetyl-CoA (coenzyme A) to form carbon dioxide and formyl-S-CoA. Frederick et al. (1963) gave normal patients a 1 ,umol dose of i4C- labeled glyoxylic acid intravenously. The subjects excreted approximate- ly 15% of the administered ~4C as i4CO2 during the first 3 hr after administration. Patients with primary hyperoxaluria, a genetic disorder of oxalate metabolism, excreted only 2% to 5% of the dose as carbon dioxide. However, the hyperoxaluric patients excreted 25% of the administered dose as oxalate and 15% as glycolate in the first 24-hr urine sample, in contrast to approximately 12% and 4137O, respectively, in control subjects. No information on metabolism of glyoxal or methyl glyoxal was available for review. HEALTH ASPECTS Observations in Humans Other than the metabolic studies cited above, no data concerning effects of glyoxylic acid in humans were available for review. It appears from the report of Frederick et al. (1963) that a single

Toxicity of Selected Drinking Water Contaminants 209 dose of at least 1 ,umol of glyoxylic acid can be tolerated without serious nJury. Glyoxal was rated as a strong sensitizer when applied as a logo solution to human skin in tests to evaluate contact sensitizers (Kligman, 1966). There were no data concerning effects of methyl glyoxal on human health. Observations in Other Species Acute Effects Bove (1966) conducted limited studies of the acute toxicity of glyoxylic acid in rats in connection with studies of the toxicity of ethylene glycol. Rats that were given oral doses of glyoxylate at 3 or 6 g/kg body weight all died within 1 hr. They tolerated 1 g/kg for at least 24 hr prior to sacrifice; however, all of the rats had renal tubular oxalosis. In these experiments, glyoxylate was more acutely toxic than ethylene glycol, glycoaldehyde, or glycolic acid. Richardson ( 1973) confirmed that glyoxylate had greater toxicity than ethylene glycol and glycolate and found that partial hepatectomy increased the susceptibility of rats to oral 0.5-g doses of sodium glyoxylate. Three of 15 intact rats died; 6 of 15 died in a group that had one-third hepatectomies 24 hr previously; and 11 of 15 died in a group that had two-thirds hepatecto- mies. The author concluded that the acute toxicity of glyoxylate was not due to the oxalate that was formed but, rather, to an eject of glyoxylate itself. However, this conclusion is tenuous because hepatectomy also resulted in increased excretion of urinary oxalate. Laborit et al. ( 1971 ) found that 200 mg/kg of glyoxylate given intraperitoneally was an Logo in rats. Mice were less susceptible with an approximate intrapentoneal LD50 of 325 mg/kg. Glyoxylate was twice as toxic as glycolaldehyde and approximately 5 times as toxic as glycolate for mice. Deaths occurred rapidly and did not appear to involve either a direct mechanism of the central nervous system or to depend on the formation of oxalate. McChesney et al. (1972) reported that an oral dose of 500 mg/kg of glyoxylic acid '`appeared to be well tolerated" by female rhesus monkeys. In-vitro and ~n-vivo studies have demonstrated that glyoxylic acid inhibits oxidative metabolism, and it seems probable that a direct metabolic effect of glyoxylate is responsible for its acute toxicity (Kleinzeller, 1943; Lamothe et al., 197 1~. Smyth e' al. (1962) reported that a solution of 29% glyoxal was acutely toxic when fed to rats in doses of approximately 7.46 ml/kg body weight.

210 DRINKING WATER AN-D H"LTH Dermal LD50 was greater than 20 ml/kg for this solution which was mildly rated or moderately irritating to rabbit skin and eyes. Chronic Effects There were no data concerning the effects of chronic exposure to glyoxylic acid, glyoxal, or methyl glyoxal. Carcinogenicity No available data. Mutagenicity No available data. Teratogerlicity No available data. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARLJ The only available data concern acute (LDso) exposures and are inadequate for estimates of a SNARL for either acute or chronic exposure. The chemistry of ozonization of water should be reviewed and further research conducted to determine if the treatment of drinking water with ozone is likely to result in the formation of glyoxylate or related compounds that would deliver a dose to humans at approximately 100 mg/day. If so, extensive research on the chronic effects of this substance, especially in relation to hyperoxaluric conditions, should be conducted. The 100 mg/day dose is suggested as a guideline because this dose appears to be the approximate lower limit of daily endogenous production of glyoxylate (Hockaday et al., 1964~. Hexachlorobenzene (C6Cl6) This compound was evaluated in Drinking Water and Health (National Academy of Sciences, 1977, p. 6671. The following material, which became available after that 1977 publication, updates and, in some instances, reevaluates information in the earlier report. Also included are some references that were not assessed in the original report. METABOLISM Absorption of hexachlorobenzene (HCB) is extremely poor when administered to rats in aqueous suspension but is quite good (average 80%) when the chemical is given in an oil solution (Koss, 1976~. Peak

Toxicity of Selected Drinking Water Contaminants 211 tissue radioactivities were reached 2 to 5 days following administration of i4C-labeled HCB. A urinary metabolite, 2,4,5-trichlorophenol, was identified in the urine of rats that had been fed HCB (Renner and Schuster, 1977~. HCB is degraded in the rat primarily to pentachlorophenol and then to 2,3,4,5- tetrachlorophenol (Engst et al., 1976~. Minor metabolites of HCB in the rat include pentachlorobenzene, 2,3,4,6-tetrachlorophenol, 2,3,4tri- chlorophenol, and 2,4,6-trichlorophenol. HEALTH ASPECTS Observations in Humans During 1955-1959, an outbreak of human poisoning occurred in Turkey as a result of the consumption of HCB- treated wheat (Cam and Nigogosyan, 1963; DeMatteis et al., 1961; Schmid, 1960~. Some deaths resulted, but the major syndrome was cutaneous porphyria with skin lesions, porphyrinuria, and photosensi- tization. The estimated dosage was approximately 50 to 200 mg/day (0.71 to 2.9 mg/kg/day) for long periods before toxic manifestations became apparent (Cam and Nigogosyan, 1963; Schmid, 1960~. Peters (1976) described the Turkish poisoning incident and the successful use of ethylenediaminetetraacetic acid (EDTA) in the treat- ment of poisoned individuals. Observations in Other Species Subchronic and Chronic Effects HCB was administered daily by Savage to adult female rhesus monkeys for 60 days at doses ranging between 8 and 128 mg/kg (Iatropoulos et al., 1976~. Cloudy swelling and centrilobular hepatocellular hypertrophy were seen in the liver of the HCB-treated monkeys including the one animal that was dosed at 8 mg/kg/day. In a 90-day study, male pigs received 0.05, 0.5, 5.0, and 50 mg/kg/day of 100% pure HCB mixed in a diet (den Tonkelaar et al., 1978~. Animals that consumed the highest dosage level died during the experiment. Increased excretion of coproporphyrin was observed in the 5.0 and 0.5 mg/kg/day groups, and induction of liver microsomal enzymes and characteristic histopathological changes in the liver occurred at 5.0 mg/kg/day. The investigators determined the no-e~ect level to be 0.05 mg/kg/day under the conditions of this experiment. HCB treatment results in a suppression in the immune response of mice that are fed a diet containing 167 ppm HCB for 6 weeks (Loose et

212 DRINKING WATER AND HEALTH al., 1977) and in rats fed a diet with 250 ppm HCB for 8 to 10 weeks (Zprin and Fowler, 1977~. Groups of six male and six female beagle dogs were given daily doses of 1,000, 100, 10, and 1 mg of 99.9% pure HCB per dog for 12 months (Gralla et al., 1977~. Mortality, anorexia, and weight loss occurred at the highest and, to a lesser degree, at the next highest dosage level (approximately equivalent to 108 to 176 mg/kg/day and 10.1 to 16.5 mg/kg/day, respectively). A dose-related neutrophilia appeared in animals receiving the two highest dosages. While the dogs were free of porphyria, nodular hyperplasia of gastric lymphoid tissues was found in all treated dogs including those given 1 mayday (equivalent to 0.01 to 0.1 5 mg/kg/day). Livers of male and female Sprague-Dawley rats that had been fed diets containing 1, 5, 10, and 25 ppm pure HCB for 3 to 12 months were examined by electron microscopy (Mollenhauer et al., 1975~. There were significant changes in the cellular regions, which contained smooth endoplasmic reticulum as early as 3 months at dietary levels of HCB as low as 5 ppm (0.25 mg/kg/day). Mitochondria became elongated or swollen under similar conditions. In a second report, Mollenhauer et al. (1976) observed a unique degeneration of lipid vesicles into an autophagic vacuole or storage vesicle in rats that had been fed 5 ppm HCB (0.25 mg/kg/day) for 3 to 12 months. Rats receiving diets containing 50 ppm HCB for 2 weeks incurred increases in hepatic microsomal mixed-function oxidase activities (den Tonkelaar and van Esch, 1974), but no inductive effect was seen at 20 ppm (1.0 mg/kg/day). Food deprivation stimulated the induction of microsomal enzyme activity in male and female rats that had consumed 40 ppm HCB (2.0 mg/kg/day) and also resulted in higher tissue HCB concentrations (Villeneuve et al., 19771. Stonard and Greig (1976) fed female rats diets containing 0.01% HCB (mean daily intake of 1.89 mg/kg/day) for 90 days. There was a mixed pattern of hepatic microsomal enzyme induction, with some characteris- tics of both the 3-methylcholanthrene and phenobarbital type of inducers. They observed similar but less dramatic effects after a 14-day feeding at the same level Daily oral administration of 10 mg/kg HCB to adult male rats for 14 days caused significant increases in hepatic cytochrome P-450 levels (Carlson and Tardiff, 19761. HCB-induced porphyria and accumulation of uropo~phyrinogen III in chick cells require the presence of endogenous iron (Sinclair and Granick, 1974) and are also markedly enhanced in siderotic rats (Louw et al., 19771. HCB treatment has no effect on erythrocyte porphyrin content but causes an increase in the porphyrin content of the kidney

Toxicity of Selected Drinking Water Contaminants 213 and spleen and a marked increase in the liver (San Martin de Viale et al., 1977~. Rajamanickam et al. (1972) reported that female rats fed 0.2% HCB in their diet for several days developed a large increase in hepatic cytochrome P-450 levels, under conditions where total heme synthesis and b-aminolevulinic acid (ALA) synthetase activity were unchanged. After a week or more of the HCB diet, the rats exhibited a twofold increase in ALA synthetase activity and severe porphyna. Experimental HCB porphyria in rats is characterized by an increased excretion of porphyrins belonging essentially to isomeric type III and a massive accumulation of uroporphyrinogen III and 7-COOH porphyrin in the kidney, spleen, and liver (San Martin de Viale et al., 1970; 1977~. The high excretion of uropo~phyrinogen III and its accumulation in the liver and other tissues of HCB-dosed animals are apparently the combined result of an increase in S-aminolevulinic acid synthetase activity and inhibition of the enzyme uropo~phyrinogen III decarboxylase (Elder et al., 1976; Louw et al., 1977~. Carcinogenicity A recent report by Cabral et al. (1977) indicates that HCB is carcinogenic in the Syrian golden hamster. Animals were fed lifetime diets containing 50, 100, and 200 ppm 99.5% pure HCB ad libitum. After 70 weeks there was a significantly higher dose-related incidence of tumors in the treated animals, particularly in males. The percentage of male control animals with tumors was 7.5% while the incidence in males receiving the 50 ppm (4 mg/kg/day), 100 ppm (8 mg/kg/day), and 200 ppm (16 mg/kg/day) doses was 60.0~o, 90.0~o, and 98.2%, respectively. Major tumor types included alveolar adenomas of the thyroid, hepatomas, liver hemangioendotheliomas, and adrenal tumors. Another, yet to be confirmed study in mice, in which only lung neoplasia was observed, failed to show any HCB-induced pulmona~y neoplasia (Theiss et al., 19771. Strain A/S + male mice were given intraperitoneal injections of HCB at 8, 20, and 40 mg/kg 3 times a week for a total of 24 injections. Twenty-four weeks after the first injection, the animals were sacrificed. The number of pulmonary adenomas in the treated mice was not significantly different from that of the controls. Carcinogenic Risk Estimate In a recent study by Cabral et al. (1977), Syrian golden hamsters were fed ad libitum for life with diets containing 50 to 200 ppm of 99.5% pure HCB. There was a significantly higher dose- related incidence of alveolar adenomas of the thyroid, hepatomas, and adrenal tumors in the HCB-exposed animals when compared with the

214 DRINKING WATER AND H"LTH controls. The available dose-response data were used to make statistical estimates of both the lifetime risk and an upper 95% confidence bound on the lifetime risk at the low dose level. These estimates are of lifetime human risks and have been corrected for species conversion on a dose/surface area basis. The risk estimates are expressed as a probability of cancer after a lifetime consumption of 1 liter/day of water containing 1 ,ug/liter of the compound. For example, a risk of 1 x 10-6 implies a lifetime probability of 2 x 10-5 of cancer if 2 liters/day were consumed; and if the concentration of the carcinogen was 10 ,ug/liter during a lifetime exposure, this compound would be expected to produce one excess case of cancer for every 50,000 persons exposed. Using 220 million as the population of the United States, there would be 4,400 excess deaths from cancer, or 62.8 per year. For hexachlorobenzene at 1 ,ug/liter, the estimated lifetime risk for humans is 1.9 x 10-5. The upper 95% confidence estimate is 2.9 x 10-5. Both of these estimates are the averaged risks which have been calculated from data on the male and female hamsters. CONCLUSIONS AND RECOMMENDATIONS The acute toxicity of HCB is relatively low, but subchronic or chronic exposure of laboratory animals or humans to the compound induces a variety of toxic effects, including the development of severe porphyria in females and, to a lesser degree, in males. Recently? Cabral et al. (1977) obtained convincing evidence that HCB is carcinogenic in the Syrian golden hamster. Lifetime feeding of diets containing as little as 50 ppm of 99.5% pure HCB (4 mg/kg/day) resulted in a 60.0% incidence of tumors as compared to 7.5% in the control animals. Since the compound appears to be a carcinogen, lifetime exposure concentrations were not estimated. Suggested No-Adverse-Response Level (SNARLJ Seven-Day Exposure Data that are suitable for estimating a 1- or 7- day SNARL for HCB are very limited. However, some estimations may be made. Well et al. (1969) have given a formula for predicting minirnal- effect dosage levels for a 1-week exposure from the results of a 90-day feeding study. Mollenhauer et al. (1975, 1976) observed abnormal liver ultrastructure in rats fed 5 ppm of HCB in the diet (0.25 mg/kg/day) for 3 to 12 months. Multiplying this value by 3.0 (Well et al., 1969), a 7-day minimal-effect dosage level of 0.75 mg/kg/day may be obtained. Using an uncertainty factor of 1,000, and assuming that all of the HCB is

Toxicity of Selected Drinking Water Contaminants 215 ingested via drinking water during this period and that a 70-kg human consumes 2 liters of water per day, the following calculation can be made to yield a SNARL in drinking water of 0.03 mg/liter for a 7-day exposure to HCB: 0.75 mg/kg/day x 70 kg 1,000 x 2 liters = 0.03 mg/liter. The data are insufficient to make any realistic estimate based on a 1- day exposure to HCB. Recently, Cabral et al. (1977) obtained convincing evidence that HCB is carcinogenic in the Syrian golden hamster. Lifetime feeding of diets containing as little as 50 ppm of 99.5 pure HCB (4 mg/kg/day) resulted in a 60.0% incidence of tumors as compared to 7.57 in the control animals. Since the compound appears to be a carcinogen, lifetime exposure concentrations were not estimated. Additional studies are needed to confirm the carcinogenicity of HCB and also to determine whether the compound is mutagenic. lodine/lodide/lodate (12, 1, 1O3 ~ Elemental iodine occurs in the diatomic state, I2, and has a molecular weight of 253.8. In solid form, it has a metallic luster, forming blue-black scales or plates. Iodine sublimes readily to a deep blue vapor, which becomes violet when mixed with air. The density of the solid is 4.93 at 20°C; its melting point is + 113.5°C; its boiling point is + 184.4°C; and its vapor pressure is 0.31 mm of mercury at 25°C. At 25°C and 1 atm, saturated air contains 408 ppm, or 4,243 mg/m3 (1 ppm in vapor = 0.0104 mg/liter, or 10.4 mg/m3) Its solubility in water at 25°C is 0.34 g/liter. The solubility of elemental iodine is greatly enhanced in the presence of iodine salts due to the reaction: I + I- I The iodine and triiodide ions are colorless (Hygienic Guide Series, 1965~. The element is prepared by reduction of iodates and periodates that occur with natural deposits of saltpeter ~aNO31. Iodide ion is oxidized by other halogens. e.g.. bromine: 2I- + Br2 ~ I2 + 2Br~. This reaction involves nucleophilic displacement of bromide (Br~) by iodide (I-), with intermediate formation of the interhalogen compound.

216 DRINKING WATER AND HEALTH iodine bromide (IBr). In a similar manner, thiosulfate is oxidized to tetrathionate by iodine by stepwise nucleophilic displacement of iodide by thiosulfate: 2S2O32- + I2 ~ S4O62 ~ 2I . Thiosulfate can be administered as an antidote against the irritant properties of elemental iodine. The solution chemistry of iodine is analogous to bromine, undergoing basic hydrolysis to the hypohalous acid, which is followed by disproportionation to the halate and halide Ions: HOD + I2~HOI + I SI2 + 60H- ~ IO3 + 5I + 3H2O Periodates (IO4-) are powerful oxidizing agents (Gould, 1962~. Chemical uses of elemental iodine or iodine compounds include pharmaceuticals, antiseptics, photographic materials, catalysts, analyti- cal reagents, and agents in the purification of zirconium and hafnium (Hygienic Guide Series, 19651. Iodine was first used as an antiseptic by a French surgeon in 1839. It is still widely used for this purpose because of its proven efficacy, economy, and low toxicity in tissue. It is bactericidal, virucidal, and amebicidal at dilutions of 5 x 10-2 to 5 x 10-3 g/liter. Its germicidal properties are present in either aqueous or alcoholic solution (Esplin, 19701. Iodine is an essential trace element, which is required for synthesis of thyroid hormone. Estimates for the iodine requirement of adult humans may range from 80 to 150 payday, which are similar to those for iodide (National Academy of Sciences, 19741. Iodine deficiency results in goiter, a compensatory hyperplasia of the thyroid. In most areas of the world, the proportion of total intake of iodine from water is negligible or very low. The source of most of the intake is food. However, iodide levels in water often serve as an indicator of high or low iodide intake in a region and correlate inversely with high or low prevalence of goiter (Under- wood, 19774. Iodide can be used therapeutically to treat hyperthyroidism and goiter. The primary source of dietary iodide is seafood, which may contain from 200 to 1,000 ,ug/kg of iodide. Kelp and other seaweeds may contain as much as 0.1% to 0.2~o iodide by weight. In Japan, where seaweed is regularly consumed as a national delicacy, goiter is virtually unknown.

Toxicity of Selected Drinking Water Contaminants 217 To provide the approximate daily requirement of 100,ug/day, table salt in the United States is supplemented with 100 ,ug of potassium iodide per 1 g of sodium chloride (Astwood, 1970~. In addition, bread made with iodine-containing dough conditioners may also supply substantial dietary iodine (Kidd et al., 1974~. A recent study by Oddie et al. (1970) determined geographic distribution of iodide intake in the United States by measuring thyroid i3iI uptake in approximately 30,000 euthyroid subjects throughout the country. Iodide intake is inversely correlated with the uptake of radioiodine in the thyroid. Regions of lowest iodide intake coincided approximately with regions previously (192~1940) having a high prevalence of endemic goiter. The present range of intake was found to be 240 to 740 ~g/day, considerably above current estimates of the iodide requirement. Major sources are thought to be iodized salt and iodine compounds in bread. Underwood (1977) cites an upward trend in iodine consumption in the United States, mainly due to cumulative addition from adventitious sources including iodates in bread, iodized salt vitamin preparations, iodine-containing medications and antiseptics, and coloring agents. Iodoform antiseptics, which are used in various stages of milk production, processing, or transportation, have increased iodide levels in milk from 13 to 23 ~g/liter to l 13 to 346 ,ug/liter. Iodine has been used to disinfect both drinking water supplies (Black et al., 1968; Morgan and Karpen, 1953) and swimming pools (Byrd et al., 1963) with excellent results. A panel determined the taste threshold of iodine to be 0.147 to 0.204 mg/liter. This was expressed as I-, but it includes all chemical forms that are present from pH 5 to 9. The iodide residual for effective disinfection in either swimming pools or drinking water is approximately 1.0 mg/liter (Bryan et al., 19731. Data in Drinking Water and Health (National Academy of Sciences, 1977) indicate that bromination and iodination are probably efficacious methods of water disinfection, but there are many unknowns surround- ing toxicity and other considerations. The same document contains a brief discussion of radioactive isotopes of iodine in the section on Radioactivity in Drinking Water. Radioiodine as a potential hazard will not be discussed in this report. METABOLISM Iodine is converted rapidly to iodide ion and is absorbed efficiently as such throughout the gastrointestinal tract (Welt and Blythe, 19704. Small amounts of the element may be absorbed through the skin. Iodine vapor reaching the lung is converted to iodide and is absorbed (`Hygienic Guide

218 DRINKING WATER AND HEALTH Series, 1965~. The volume (V) of distribution for iodide ion is larger than the total extracellular space (Welt and Blythe, 1970~. Compartmental analysis of iodide distribution in four normal men gave V = 18.2 liters, which was adjusted to the 70-kg mass (Hickey and Brownell, 1954~. Iodide accumulates in the thyroid to an extraordinary degree. Total iodine content of a healthy human adult is 15 to 20 ma. Between 7097O and 8097O of this is found in the thyroid. The normal thyroid weighs 15 to 25 g, which is 0.03% of the body weight. Much of the total iodine is found in iodoamino acids, which are incorporated into a storage protein, thyroglobulin. Even so, the free iodide ion concentration in the thyroid is normally about 50 times the plasma concentration. Normal plasma iodide occurs in the range of 0.8 to 6.0 ,ug/liter. Saliva iodide is proportional to plasma iodide from physiological levels up to about 1,OOO,ug/liter. Muscle iodide concentration is less than 0.001 times the thyroid concentration; however, due to the large total mass of muscle, the next largest proportion of total iodide is contained in this tissue. The eye contains a relatively high concentration, particularly in the orbitary fat and orbicular muscle (Underwood, 1977; Welt and Blythe, 1970~. Excretion of iodide is largely renal. It appears to be filtered and partially reabsorbed, largely by passive diffusion (Welt and Blythe, 1970~. In four human subjects with average body weights (92 kg) and serum iodide (4.1 ~g/liter), urinary excretion of iodide was 221 ,ug/day (Hickey and Brownell, 1954~. Rats excreted 505'o of an administered iodide dose in 6 to 7 hr. Metabolic equilibrium of retained iodide is attained in 4 days. The half-life of this second phase is stable at about 9.5 days, representing a turnover of organic iodine (Underwood, 1977~. Iodide is secreted in milk. Cow's mills normally contains 20 to 70 g/liter. This can be increased to 510 to 1,070 ~g/liter by feeding potassium iodide at 100 mg/day (Underwood, 1977~. HEALTH ASPECTS Observations in Humans Direct toxicity of iodine is due to its irritant properties. The Hygienic Guide Series (1965) recommended a maximum allowable concentration for iodine vapor of 0.1 ppm (1.04 mg/m3) based on human experience and analogy to chlorine. This value may be too high to prevent eye irritation. Short exposure tolerance studies in humans indicate that 0.1 ppm (1.04 mg/m3) is tolerable and that 0.3 ppm (3.12 mg/m3) is not (time limit not stated). In another study, four humans tolerated 0.57 ppm (5.93 mg/m3) of iodine vapor for 5 min without eye irritation, but all suffered eye irritation within 2 min at 1.63

Toxicity of Selected Drinking Water Contaminants 219 ppm (17.0 mg/m31. The atmospheric level that is immediately hazardous to human life is not known. Pulmonary edema is produced rapidly following exposure to high concentrations. In humans and animals, cough and excessive tearing are characteristic signs. Skin irritation is possible. Saturated vapor produces brown staining and loss of the corneal epithelium upon exposure of the human eye. Complete healing occurs in 2 to 3 days. The fatal ingested dose of I2 in humans is estimated to be 2 to 3 g (29 to 43 mg/kg), but recovery has been reported after ingestion of 10 g (`Hygienic Guide Series, 1965~. For tincture of iodine (2% I2 + 257O Nat in alcohol), the fatal dose is 30 to 250 ml (9 to 70 mg/kg of I2) Acute toxicity is due largely to irritation of the gastrointestinal tract. Little I2 is absorbed, although I- is absorbed. I2 is bound to starches and proteins in the gastrointestinal tract. Severe irritation produces fluid loss and consequent shock. Rarely are individuals hypersensitive to iodine. But those that do may react markedly, even to skin application, exhibiting skin eruptions and fever (Esplin, 1970~. Acute poisoning in humans from an initial dose of iodide is rare. Oral doses of approximately 3.3 mg/kg iodide (as KI) are commonly given every few hours as an expectorant without untoward effects, with the exception of occasional hypersensitivity with angioneurotic edema, which sometimes progresses to swelling of the larynx, resulting in suffocation (Welt and Blythe, 19701. Occasionally, other toxic responses have been seen within hours after administration of iodine. These include fever, arthralgia, lymphoadenopathy, eosinophilia, and, more rarely, multiple petechiae of the skin and mucous membranes (Bianco et al., 19711. In a report of seven case histories of adverse reactions to iodides, most of the reactions occurred in response to simple inorganic iodides, but some were observed after exposure to organic compounds that release iodide. The most frequent reaction was swelling of the salivary glands ("iodide mumps"), although a reaction resembling severe serum sickness was also observed, as were erythema and a pounding headache. Underwood (1977) and Cheftel and Truffert (1971) pointed out that iodide toxicity is relatively low and that there is a large safety margin due to prompt excretion of excess iodide. However, they noted the occasional hypersensitivity reaction. Even so, signs of toxicity generally disappear upon removal of exposure to excess iodide. Chronic iodide poisoning (iodism) is more common than acute reactions. Toxicity to iodide occurs eventually in all individuals, when they are given a sufficiently high dose (Welt and Blythe, 19701. Symptoms often resemble a "sinus cold'' and might also include acneform skin lesions and irritation of the gastrointestinal tract.

220 DRINKING WATER AND H"LTH Symptoms disappear spontaneously after administration of iodide is stopped. Bianco et al. (1971) described a 54-year-old man who reacted adversely to a vitamin preparation containing 0.15 mg of iodide (as KI) for 10 days. Symptoms of chronic iodism were present, including a marked swelling of the salivary glands. The patient returned to normal after 8 days. Other signs and symptoms that have been cited in reports of chronic iodism are a metallic or brassy taste, gingivitis, burning sensation of oral mucosa, increased salivation, coryza, sneezing, irritation of conjunctive' edema of eyelids, and severe headache. The mechanism of the syndrome is unknown, and treatment is mainly supportive and empirical, removal of iodide being the principal remedy (Bianco et al., 1971). Underwood (1977) suggested that chronic intake of 2,000 ,ug/day of iodide should be regarded as excessive and potentially harmful, although daily intake in some regions of Japan may reach 50,000 to 80,000 ,ug without apparent ill effect. Special segments of the population may be at risk to increasing iodide intake due to adventitious sources. For example, in Tasmania, where potassium iodide is added to bread and iodoform antiseptics are used in the production of milk, an increased incidence of thyrotoxicosis was observed in 40- to 80-year-old women who had preexisting goiter (Underwood, 1977~. Iodination has been proposed as a disinfection procedure in lieu of chlorination. To investigate this possibility, the chronic toxicity of iodide was evaluated by adding iodide to chlorinated water at a military field station in the tropics. The dose was 12 mg/man/day for the first 16 weeks. It was then increased to 19.2 mg/man/day for the following 10 weeks. Men were given periodic physical examinations and routine checks of body weight, skin, heart, thyroid, renal function, and hematological indices (Morgan and Karpen, 1953~. No ill effects were observed over a period of months. Black et al. (1968) studied the effectiveness of iodine in disinfection of drinking supplies and swimming pools. They evaluated the toxicity in humans at levels used in disinfection. Two water systems, which serve three prisons in Lowell, Florida, were used in the study. The total inmate population was composed of 750 men, women, and 13- to 16-year-old girls when the study began in October 1963. For 31 of the 43 months of the study, 1 ppm of I2 was added to the water at the women's prison and girls' school. This level was added to the water in the men's prison for 41 of the 43 months. For 2 months, water was supplemented with 4 ppm of iodide ion. Then, in the women's prison and girls' school, total iodide was reduced in gradual increments to 0.3 ppm for 10 months in order to evaluate bactericidal effectiveness. General health and thyroid function

Toxicity of Selected Drinking Water Contaminants 221 of a test group of 133 inmates were evaluated twice before iodine exposure began, 4 times during the first 10 months of exposure, and a fifth time after 37 months. At iodine doses of 1 ppm, radioiodine uptake by the thyroid was depressed, but protein-bound iodine changed relatively little until iodide was increased to 5 ppm. Serum thyroxine levels remained unchanged throughout the exposure. No adverse effects or hypersensitivity reactions were detected. By 37 months, the test group had declined to 29 of the original 133, but the results were the same. Serum iodide increased from a normal range of 10 to 20 ~g/liter to 60 ,llg/liter as measured at 1 month and at 37 months. In an ancillary study of nonprison personnel, who swam in a pool containing water at 5 ppm iodide, no changes in thyroid function were observed. Byrd et al. (1963) evaluated eye irritation, measured protein-bound iodine and urine iodide, and enumerated bacteria count in water after a single exposure and after 1 week and 1 month of repeated daily exposure of members of a college swim team and swim class, who swam in pools containing iodine residuals of 0.4 ppm. Disinfection efficacy was rated as good, and eye irritation was notably absent. There were no significant changes in protein-bound iodine or urinary iodide. Observations in Other Species Acute Elects The Hygienic Guide Series (1965) reported a limited study in which four rats survived exposure to 370 ppm (3,848 mg/m3) of iodine vapor. Dogs that had been exposed to high concentrations of iodine vapor rapidly developed pulmonary edema. The approximate LD50 for potassium iodide given intravenously to rats is 285 mg/kg; sodium iodide has an approximate intravenous LD50 in the rat of 1,300 mg/kg; and sodium iodate has an approximate intravenous LD50 in the rabbit of 75 mg/kg (Toxic Substances List, 1974~. Bock and Wright (1964) presented evidence suggesting that resistance to acute lethality from sodium iodide may be subject to control by a single dominant gene. They found considerable strain difference in 14- day LDso values which were obtained following intraperitoneal injection of sodium iodide in mice. Fourteen strains of mice were used, giving LD50 values ranging from 430 to 2.030 mg/kg. Clifton and Makous ( 1973) reported a disturbing acute erect of sodium iodate. A single intravenous dose of approximately 30 mg/kg in rabbits caused eventual retinal degeneration and blindness. This effect of iodate is rather specific. It is attributable to effects on the retinal epithelium occurring within 1 hr of dosing (Clifton and Makous, 19731.

222 DRINKING WATER AND H"LTH Chronic Elects Underwood (1977) identified four stages of response to excess levels of iodide intake, all pertaining to thyroid function: at low levels, a temporary increase in the absolute uptake of iodine by the thyroid and incorporation into organic iodine compounds; inhibition of release of iodine from thyrotoxic thyroid; and inhibition of organic iodide formation, leading to iodide goiter (Wolff-Chaikoff effect); and, at very high levels, saturation of action transport for iodide and, usually, intervention of acute toxic effects of iodide. There is considerable species difference in susceptibility to toxic ejects of iodide. Also, there appears to be a high tolerance to normal nutritional levels of iodide, pointing to a large margin of safety. Domestic livestock are frequently given dietary supplements of iodine as prophylaxis or treatment for various diseases. McCauley et al. (1972) reported four field cases in which organic iodine supplements seemed to interfere with the ability of cattle to cope with infectious or noninfectious insult under situations of added stress (e.g., transportation, weather, high milk production, or parturition). In these situations, cattle on ethylene- diamine dihydroiodide, which is used to prevent foot rot, developed coughing and increased bronchial, nasal, and lacrimal secretion. Serum iodide levels were 1,030 to 2,750 ~g/liter. The normal range in cattle is 60 to 80,ug/liter. Newton et al. (1974) fed 0 to 200 ppm iodide to Holstein bull calves as calcium iodate for approximately 100 days. The normal dietary require- ment of cattle for iodide is about 0.1 ppm in the feed. Animals fed 100 to 200 ppm developed coughing and profuse nasal discharge. Eject on thyroid weights was variable. All levels of iodate in feed increased serum iodide concentrations. At 200 ppm there was decreased hemoglobin and serum calcium and increased serum butanol-extractable iodine. In two of the trials, in which iodide was fed at 25 to 100 ppm, all levels increased adrenal weights. Above 50 ppm there was decreased body weight gain and feed intake. Minimum toxic dose was thought to be near 50 ppm, although a portion of the animals had adverse signs at lower levels. At 25 ppm, the daily dose of iodide would have been 1.42 mg/kg at the start of feeding and 0.54 mg/kg at the end (due to body weight gain), based on figures given for feed consumption (Newton et al., 1974~. In similar experiments, Newton and Clawson (1974) found swine to be more resistant to iodine in the diet than were cattle. An iodide concentration of 0.2 ppm is sufficient to meet the dietary requirement for this element in swine. In these experiments, one group of swine received 0, 10, 20, 40, and 80 ppm (as I-) in the form of calcium iodate. A second group of swine received 25, 50, 100, 200, 400, 800, and 1,600 ppm in the same manner. Body weight, feed intake, thyroid weight, liver iron, total

Toxicity of Selected Drinking Water Contaminants 223 serum iodine, butanol-extractable iodine, and hemoglobin were mea- sured. All levels of iodine supplementation increased serum iodine levels and thyroid weights, but no differences in performance were measured at or below 400 ppm in the feed. At 400 and 1,600 ppm, liver iron was significantly decreased. Growth rate, feed intake, and hemoglobin were decreased at 800 ppm. Dietary or intramuscular iron supplements offset the effects of iodide on growth rate and feed intake, but had no effect on serum iodide level. For the first 28 days of an 84-day exposure to 0 to 80 ppm, there was salivary gland swelling, which then subsided. This swelling was not noted in high-dose experiments. At 800 and 1,600 ppm, there was encrustation around the eyes due to increased lacrimal secretion. One pig in each of the 800- and 1,600-ppm groups developed corneal opacity and hyperplasia, which was confined to the cornea. For overt signs of toxicity that did not reverse spontaneously on continued feeding, the minimum toxic dose was considered to be 400 ppm (corresponding to about 11 mg/kg iodide at the end of the feeding period) (Newton and Clawson, 1974). Braverman and Ingbar (1963) reported that rats that received 0.01% iodide as sodium iodide ire drinking water for 5 days and then 500,ug parenterally 3 hr before sacrifice adapted to the acute inhibitory effect of iodide on incorporation of iodide into organic iodine compounds. The adaptation appears to be at the level of active transport of iodide into the thyroid (Braverman and Ingbar, 1963~. Cantin ( 1967) demonstrated that erythema in rats produced by repeated daily doses of various inorganic iodide salts (0.125-1.00 mmol/kg) could be prevented by prior chronic treatment (20 days at 0.5 mmol/kg). An injection of Evans blue dye showed increased permeabili- ty of skin capillaries that was associated with erythema. Cantin also observed that damage to mast cells was dose-related. Ammerman et al. (1964) fed adult female rats 0 to 2,500 ppm iodide from 0 to 35 days post partum. Some were allowed to mate, and the survival of their offspring was noted. Others were killed at 17 to 19 days of gestation or 24 to 48 hr post partum. The ovulation rate, implantation rate, fetal development, and histology of mammary tissue were exam- ined. Offspring showed increasing mortality with increasing dose. Most deaths of the young rats occurred within 24 hr. Rats surviving 48 to 72 hr post partum usually survived until weaning. Dead pups had virtually no milk in their stomachs. Secretion of milk in mammary glands of mothers was absent or markedly diminished. No other significant effects were noted. In an ancillary experiment, Ammerman and coworkers fed iodide at 2,500 ppm to male rats for 200 days. This resulted in no impairment in reproductive performance.

224 DRINKING WATER AND H"LTH In the same study, they found no adverse ejects on reproduction in males or females other than pup survival due to decreased lactation in the mothers, which resulted from high dietary levels of iodide. Mutagenicity No available data. Carcirzogenicity No available data. Teratogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS The toxicity of iodine and iodide appears to be relatively low, although hypersensitivity does occur. Levels of approximately 100 ,ug/day must be consumed to maintain suitable steady-state concentrations of this essential element. Because of the essentiality of iodine, Pechenkina (1964) has recommended that levels of iodine in drinking water be no less than 0.002 ppm. Dawson ( 1974) has recommended 10 mg/liter of iodide as a safe level for potable water, but this would result in a daily intake of at least 20 mg from water alone, which is excessively high. Since hypersensitivity apparently does not occur at levels of iodide that provide the dietary requirement, similar levels in water should be acceptable. Data for calculating a SNARL value for elemental iodine or for iodate are lacking although there are some data for total iodine in drinking water. Care should be taken when substituting iodates for iodides in dietary supplements in view of the specific retinal toxicity of iodate. This latter aspect should be investigated further. Suggested No-Adverse-Response Level (SNARLJ For iodide, estimates of acceptable levels can be based on clinical experience with iodides and the above-mentioned test exposures of humans to iodide in drinking water. Twenty-Four-Hour Exposure Using the typical dose of 3.3 mg/kg iodide (as an expectorant), and assuming lOO~o intake from 2 liters/day of drinking water by a 70-kg human during this period: 3.3 mg/kg x 70 kg 2 liters = 115.5 mg/liter.

Toxicity of Selected Drinking Water Contaminants 225 Seven-Day Exposure Since a dosing regimen of iodide expectorant could be expected to continue over several days, it would not be unreasonable to use the single dose for the 24-hr exposure and simply divide it over a 7-day period: I 15.5 mg/liter = 16 5 mg/1iter. Over 7 days at this level, some hypersensitivity reactions may be encountered. Chronic Exposure Human studies involving 16 weeks of exposure to approximately 0.17 mg/kg/day iodide in drinking water followed by 10 additional weeks at 0.27 mg/kg/day resulted in no ill effects. This SNARL assumes a 2G% intake from drinking water: 0.17 mg/kg x 70 kg x 0.20 2 liters = 1.19 mg/liter. This calculation corresponds to the finding of no ill effects in the prison and girls' school study' which used iodine disinfection at a level of 1 ppm I2 for 31 months. Nonanal [CH3(CH2~7CHO] Nonanal? which has a molecular weight of 142.24, is found in at least 20 essential oils' such as rose' citrus, and pine. In the United States, less than 4 550 kg/year is used as an ingredient in various fragrances (Opdyke. 1973~. It is approved by the Food and Drug Administration for use in food (2 1 CFR 1 2 1.11 64) as a result of a GRAS (generally recognized as safe) review in 1965. The Council of Europe recommends an ADI of 1.0 mg/kg. The Food and Agricultural Organization/World Health Organization (1969) recommends an ADI of 0.1 mg/kg. M ETABOLISM No data are available on this aldehyde specifically, but the C-8 and C-10 are readily oxidized by animals to fatty acids and then to carbon dioxide and water.

226 DRINKING WATER AND H"LTH HEALTH ASPECTS Observations in Humans Maximization tests in 25 volunteers using 1% concentration resulted in no sensitization (Kligman, 1966~. Observations in Other Species Acute Effects Acute oral LD50 in rats and dermal LD50 in rabbits were >5.0 g/kg (Shelanski, 1971~. No symptomatology was given. Chronic Effects No available data. Mutagenicity No available data. Carcinogenicity No available data. Teratogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS The risk of acute hazard would appear to be very low in view of the high (>5 g/kg) oral and topical LDso in animal tests. These very high acute LD50 results suggest that the ADI established by the Council of Europe (I mg/kg) is providing a high, though not quantifiable, safety factor. No estimate of acute or chronic risk can be made due to lack of data. Octanal [CH3(CH2~6CHO] Octanal, which has a molecular weight of 128.22, is found in essential oils such as citrus oils. In the United States less than 4,550 kg/year is used as an ingredient in various fragrances. It is approved by the Food and Drug Administration for use in food (21 CFR 121.1164) as a result of a GRAS (generally recognized as safe) review in 1965. The Council of Europe recommends an ADI of 1 mg/kg. METABOLISM According to Williams (1959) octanal is readily oxidized in animals to the corresponding fatty acid which is then normally oxidized to carbon dioxide and water.

Toxicity of Selected Drinking Water Contaminants 227 HEALTH ASPECTS Observations in Humans A standard repeat-insult patch test in 40 subjects, using 0.25~o concentration in alcohol, resulted in no sensitiza- tion (Majors, 1972~. Observations in Other Species Acute Effects Smyth et al. (1962) fouled that the single dose oral LDso in rats is 5.63 ml/kg; the single skin penetration LD50 in rabbits is 6.35 ml/kg; 8-fur inhalation of"concentrated" vapor by rats results in no deaths; and skin and eye irritations in rabbits are mild. They reported no symptomatology. Penetration through intact mouse skin was reported by Meyer et al. (1959~. Chronic Effects No available data. Mutagenicity No available data. Carcinogenicity No available data. Tera~ogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS The acute data in animals, the one topical study in humans, and the brief, although apparently straightforward, metabolic conversion of octanol seem to characterize a compound of low hazard if orally ingested. To consume the Council of Europe's recommended ADI of 1 mg/kg would require a concentration of 35 mg/liter of water (assuming consumption of 2 liters/day in a 70-kg human). An ADI of even this magnitude would still appear to allow for a large safety factor based on the acute oral data in rats. The lack of subacute or chronic toxicity data in humans or laboratory animals precludes assessment of acute or chronic risk. Resorcinol [C6H4~0H)21 Resorcinol is used industrially in the production of dyes, plasticizers, textiles, resins, pharmaceuticals, and adhesives for wood, plastics, and

228 DRINKING WATER AND H"LTH rubber products (Flickinger, 1976~. In 1974, the U.S. production of resorcinol was 16 million kg (U.S. International Trade Commission, 1976), 4,500 kg was imported, and 3.6 million kg was exported. Typically, industrially produced resorcinol has a purity greater than 99.5% (Flickinger, 1976~. The acute toxicity of resorcinol has been compared to that of phenol and catechol (Deichmann, 1971; Deichmann and Keplinger, 1963~. Recent reviews concerning the toxicity of resorcinol have also considered the similarity of its response to that produced by closely related agents such as hydroquinone and catechol (Flickinger, 1976; International Agency for Research on Cancer, 1977~. The threshold limit value for occupational exposures to resorcinol, based on a time-weighted average value (8-hr workday, 40-hr workweek), was set in 1976 as 45 mg/m3 (10 ppm) (American Conference of Governmental Industrial Hygienists, 1976~. METABOLISM Resorcinol is readily absorbed from the gastrointestinal tract. However, compared to catechol or phenol, resorcinol is less readily absorbed through the skin of rabbits (Flickinger, 1976~. Resorcinol and its conjugation products with hexuronic, sulfuric, and other acids are found in urine (Deichmann and Keplinger, 1963~. HEALTH ASPECTS Observations in Humans The application of resorcinol to skin may cause redness, itching, dermatitis, edema, or corrosion (Deichmann and Keplinger, 19631. Additionally, ingestion may produce hypothermia, hypotension, decreased respiratory rate, tremors, and hemoglobinuria (Deichmann and Keplinger, 1963~. Observation of production and maintenance workers exposed to resorcinol (analytical surveys showed airborne concentrations of resor- cinol up to 9.6 ppm) revealed no lost time at work and no compensation claims due to occupational disease (Flickinger, 1976~. DoPico et al. (1975) reported inflammatory diseases involving the upper and lower respiratory tract of workers in the synthetic rubber tire industry who had been exposed to a thermosetting resin containing resorcinol and methylene aminoacetonitrile. However, they did not establish the relationship between occupational disease and resorcinol.

Toxicity of Selected Drinking Water Contaminants 229 Observations in Other Species Acute Elects The acute oral LD50 of resorcinol ranges from 0.37 to 0.98 g/kg of body weight in various animal species (Deichmann and Keplinger, 1963; Flickinger, 1976~. The lethal intravenous dose for dogs is approximately 0.7 to 1.0 g/kg (Deichmann and Keplinger, 1963~. The single dose LD50 for resorcinol applied to the skin of albino rabbits for 24 hr is 3.36 g/kg (Flickinger, 1976~. Resorcinol is also an eye irritant but not a primary irritant of the skin in rabbits (Flickinger, 1976~. Acute toxic effects in rats that were exposed to single 1-fur or 8-fur inhalation exposures to resorcinol-water aerosols were not detected. In rats exposed to resorcinol at 2,000 to 7,800 mg/m3 for 8 hr. there was no effect on weight gain after 14 days. No gross lesions attributable to inhalation of the aerosols were noted at autopsy (Flickinger, 1976~. Flickinger ( 1976) also reported no evidence of gross respiratory damage in rats or guinea pigs whose throats had been sprayed 3 times a day for 2 weeks with a 197O water solution of resorcinol. Histopathological examination of lungs showed slight tracheobronchitis, but the incidence was the same in controls and resorcinol-treated animals (Flickinger, 19761. No evidence of toxic ejects was detected in rats, rabbits, or guinea pigs that were sacrificed periodically over several months following inhalation exposure to water aerosols of 34 mg/m3 resorcinol 6 hr per day for 2 weeks (Flickinger, 1976~. Chronic Elects No available data other than that noted below under . . . carc~nogen~c~ty. Mutagenicity No mutations were induced by resorcinol in Salmonella typhimurium in the presence or absence of rat liver activating systems (Ames et al., 1975; Dean, 1978; McCann et al., 1975a). Furthermore, resorcinol had no eject on the incorporation of tritiated thymidine into testicular DNA of mice (Seller, 1977~. However, Dean (1978) showed that resorcinol induced chromosome breakage in plant cells. Carcinogerlicity Resorcinol was not cocarcinogenic in mice in combi- nation with benzofa~pyrene (Van Duuren and Goldschmidt, 19761. Moreover, no tumor-promoting activity of resorcinol has been detected (Van Duuren and Goldschmidt, 1976~. The repeated application of resorcinol to the skin of female mice for their lifetime produced no statistically significant increases in tumor incidence (Stenback and Shubik, 19741. In addition, application of 0.02 ml of a 5570, 10%, or 5057O solution of resorcinol to the dorsal skin of

230 DRINKING WATER AND H"LTH rabbits produced no chemically related toxicity or tumongenesis following lifetime observation (Stenback, 1977~. Hair dye formulations containing resorcinol as components also failed to produce evidence of toxicity, aplastic anemia, or carcinogenicity when tested for 18 months by topical application to mice (Burnett e' al., 1975, 1976, 1977-781. Teratogenicity Topical applications of hair dye formulations contain- ing resorcinol to pregnant rats on gestational days 1, 4, 7, 10, 13, 16, and 19 failed to elicit any teratogenic or embryotoxic effects (Burnett et al., 1976~. Teratological studies using resorcinol alone have not been reported. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response-Level (SNARLJ Twenty-Four-Hour Exposure The data on the acute oral toxicity of resorcinol are inadequate; however, no-adverse-response dosages have been determined for resorcinol in rats following exposure to resorcinol in air at 7,800 mg/m3 for 8 hr (Flickinger, 1976~. Assuming that this dose represents the largest minimal-effect dose, and an uncertainty factor of 1,000, the following 24-hr SNARL value may be calculated for a 7~kg human consuming 2 liters of water daily with lOO~o of exposure coming from water during this period. Both of the acute SNARL's are based on data derived from inhalation studies and a 30% retention. 7,800 mg/m3 x 10 m3/dlay x 0.3 (% retention) 1,000 x 2 liters = 11.7 mg/liter Seven-Day Exposure Inhalation of water aerosols of resorcinol (34 mg/m3) 6 hr/day for 2 weeks was tolerated without adverse effects by rats, rabbits, and guinea pigs (Flickinger, 19761. Using these data, an uncertainty factor of 100, and the same assumptions as above, the following calculation may be made for a 7-day SNARL: 34 mg/m3 x 10 m3/day x 0.3 (% retention) 100 x 2 liters = 0.5 mg/llter. This value may be far lower than necessary since lifetime topical application of resorcinol to the skin of female mice produced no carcinogenicity or toxicity (Stenback and Shubik, 1974~. However, there are no data allowing for the calculation of a chronic SNARL based upon oral resorcinol exposure.

Toxicity of Selected Drinking Water Contaminants 231 The acute toxic ejects of resorcinol are similar to those of catechol, hydroquinone, and phenol. However, resorcinol is less toxic following dermal exposure than either catechol or phenol. No long-term toxic effects due to resorcinol have been observed in humans or animals. Short-term mutagenesis tests do not indicate any mutagenic properties for resorcinol. Furthermore, topical applications of resorcinol to the skin of animals over a lifetime have not indicated any increases in tumor . . inch pence. However, studies to determine chronic effects of resorcinol following long-term oral or inhalation exposure are required to assess fully the potential toxicity of resorcinol. Adequate teratogenicity and mutagenici ty studies are also required. In the absence of essential inflation, no limit values for resorcinol in drinking water can be recommended. REFERENCES Abdallah, A., and M. Saif. 1962. Tracer studies with antimony-124 in man. Pp. 287-309 (discussion, pp. 317-325) in G. E. W. Wolstenholme, and M. O'Connor, eds., Ciba Foundation Symposium on Bilhaniasis, Cairo. Little, Brown & Co., Boston. Adams, E. M., H. C. Spencer, V. K. Rowe, and D. D. Irish. 1950. Vapor toxicity of 1,1,1- trichloroethane (methylehloroform) determined by experiments on laboratory animals. AMA Arch. Ind. Hyg. Oeeup. Med. 1:225-236. Adams, E. M., H. C. Spencer, V. K. Rowe, D. D. MeCollister, and D. D. Irish. 1951. Vapor toxicity of triehloroethylene determined by experiments on laboratory animals. AMA Arch. Ind. Hyg. Oeeup. Med. 4:469~81. Adir, J., D. A. Blake, and G. M. hIergner. 1975. Pharmaeokineties of fluorocarbon 11 and 12 in dogs and humans. J. Clin. Pharmaeol. 15: 76~770. Aksoy, M., K. Dineol, S. Erdem, and G. Dincol. 1972. Acute leukemia due to chronic exposure to benzene. Am. J. Med. 52:16(}166 Aksoy, M., S. Erdem, and G. Dineol. 1974a. Leukemia in shoe-workers exposed chronically to benzene. Blood 44:837-841. Aksoy, M., S. Erdem, K. Dineol, T. Hepyuksel, and G. Dincol. 1974b. Chronic exposure to benzene as a possible contributory etiologie factor in Hodgkin's disease. Blut 28:293- 298. Aksoy, HI., S. Erdem, and G. Dincol. 1976. Types of leukemia in chronic benzene poisoning. A study in thirty-four patients. Aeta Haematol. 55:65-72. Albro, P. W., and L. Fishbein. 1972. Intestinal absorption of polychlorinated biphenyls in rats. Bull. Environ. Contam. Toxieol. 8:26-31. Alexander, P., M. Fox, K. A. Stacey, and L. F. Smith. 1952. Reactivity of radiomimetic compounds. 1. Cross-linking of proteins. Bioehem. J. 52:177-1B4. Allen, J. R. 1975. Response of the nonhuman primate to polyehlorinated biphenyl exposure. Fed. Proe. Fed. Am. Soe. Exp. Biol. 34:1675-1679. Allen, J. R., and O. H. Norbaek. 1973. Polyehlorinated biphenyl- and triphenyl-indueed gastric mucosal hyperplasia in pnmates. Seienee 179:498~99.

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