Drinking Water and Health,: Volume 4 (1982)

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vll Toxicity of Selectecl Organic Contaminants in Drinking Water The compounds evaluated in this chapter were selected for essentially the same reasons as those enumerated at the beginning of Chapter VI. Chloroform is included because new information has become available since it was evaluated in the first volume of Drinking Water and Health (National Academy of Sciences, 1977~. Dibromochloropropane appears for the first time in this report because of its occurrence in drinking water resulting from its use in agriculture. Toxicological information is required for a large number of solvents that are appearing with greater frequency in wastewater, thereby posing a real or potential hazard for contamination of drinking water. Among them are nitrophenols, nitrobenzene, petroleum products, and polynuclear aromatic hydrocarbons, which are also evaluated in this chapter. Table VII-1 sum- marizes the acture and chronic SNARL's for the compounds reviewed in this chapter. Acetonitrile (CH3CN) Nitriles characteristically contain a cyano group, C-N. Acetonitrile a mononitrile having the formula CH3CN-is also known as methylcyanide, cyanomethane, or ethanenitrile (National Institute for Occupational Safety and Health, 1978a). A colorless liquid, acetonitrile is infinitely soluble in water and has a molecular weight of 41.1. Because of their versatile chemical reactivity, nitrites have many industrial uses, such as in the manufacture of plastics, synthetic fibers, and elastomers, and as a solvent 202

Toxicity of Selected Organic Contaminants in Drinking Water 203 TABLE VII-1 Summation of Acute and Chronic Exposure Levels for Organic Chemicals Reviewed in this Chapter Suggested No-Adverse-Response Level (SNARL), mg/liter, by Exposure Perioda Chemical 24-Hour 7-Day Chronic Nitrobenzene 0.035 0.005 Mononitrophenol 0.29 Dinitrophenol 0.11 b Tr~nitrophenol 4.9 0.2 Benzene 0.25 2,4,6-Tr~chlorophenol 17.5 2.5 aSee text for details on individual compounds. bThis is the average from two calculated SNARL's; see text for details. in the extractive distillation that separates olefins from diolefins, butadiene from butylene, and isoprene from isopentene (Merck Index, 1976; National Institute for Occupational Safety and Health, 1978a; Pozzani et al., 1959~. In 1964 approximately 1,575 metric tons of acetonitrile were used in the United States (National Institute for Occupational Safety and Health, 1978a). The major occupational exposures to nitrites occur primarily by the der- mal and inhalation routes. Depending upon the amount absorbed, nitrites may cause hepatic, renal, cardiovascular, gastrointestinal, and central nervous system disorders, regardless of the route of administration. Although these effects are usually attributed to the metabolic release of cyanide, they may also be partly due to the intact molecule (National In- stitute for Occupational Safety and Health, 1978a). The time-weighted average (TWA) standard for nitrites published by the National Institute for Occupational Safety and Health (NIOSH) is based on reports indicating that certain nitrites are sources of cyanide ions. The TWA for acetonitrile is 20 ppm for up to a 10-hour workshi* in a 40-hour work week (National Institute for Occupational Safety and Health, 1978a). METABOLISM There has been little work concerned with the metabolism and disposition of acetonitrile. In studies by Dequidt and Haguenoer (1972) and Hague- noer and Dequidt (1975a), rats received intraperitoneal injections of

204 DRINKING WATER AND HEALTH acetonitrile in doses ranging from 600 to 2,340 mg/kg. The tissues were analyzed for acetonitrile and both free and combined hydrogen cyanide content. In general, acetonitrile was found to be rather evenly distributed among the various organs, and hydrogen cyanide was found in nearly all organs in varying concentrations. In subsequent studies, Haguenoer and Dequidt (1975b) obtained similar results following the administration of acetonitrile to rats via the inhalation route. A series of studies have shown that humans absorb nitrites through the skin (Sunderman and Kincaid, 1953; Wolfsie, 1960) and through the res- piratory tract (Amdur, 1959; Dalhamn et al., 1968; McKee et al., 1962; Pozzani et al., 19591. After absorption, nitrites may be metabolized to an alpha-cyanohydrin or to inorganic cyanide, which is oxidized to thio- cyanate and excreted in the urine. Nitriles also undergo other types of reactions depending on the moiety to which the cyano group is attached. The cyano group may be converted to a carboxylic acid derivative and am- monia or may be incorporated into cyanocobalamin (National Institute for Occupational Safety and Health? 1978a). HEALTH ASPECTS In general, adverse effects resulting from exposure to the nitrites' in- cluding acetonitrile, occur primarily by the dermal and inhalation routes. Depending on the amount absorbed, nitrites may cause toxic effects in- volving the hepatic, renal, cardiovascular, gastrointestinal, and central nervous systems. These effects may be due in part to the intact molecules, but are also attributed to the metabolic release of cyanide. There are substantial differences among the various nitrites with regard to the amounts necessary to cause poisoning, the durations of exposure' and the time intervals between exposure and manifestation of the adverse effects. These differences are associated with the rate and extent of the release of cyanide ion (Amdur, 1959; National Institute for Occupational Safety and Health, 1978a). Observations in Humans Acute exposure of humans to acetonitrile by inhalation results in head- ache, dizziness, profuse sweating, vomiting, giddiness, hypernea, difficul- ty in breathing, palpitations, irregular pulse, convulsions, loss of con- sciousness, and death, usually resulting from respiratory arrest (Amdur, 1959; Grabois, 1955; National Institute for Occupational Safety and Health, 1978a). Most deaths have followed industrial exposures, and ef

Toxicity of Selected Organic Contaminants in Drinking Water 205 fects have generally been observed anywhere from 3 to 12 hours after ex- posure (National Institute for Occupational Safety and Health, 1978a). This delayed onset of effects has been explained by Amdur (1959) as a slow release of cyanide and its metabolism to thiocyanate. There are ap- parently no reports of toxicity in humans following ingestion of acetonitrile. Observations in Other Species Acute toxic effects observed in animals include labored breathing, anuria, ataxia, cyanosis, coma, and death. Tissue distribution studies indicate that mononitriles are distributed uniformly in the various organs and that cyanide metabolites are found in the spleen, stomach, and skin, smaller amounts being present in the liver, lungs, kidneys, heart, brain, muscle, intestines, and testes (Dequidt and Haguenoer, 1972; Pozzani et al., 1959). Most studies conducted in animals have been concerned with toxicity following inhalation. Studies of orally adminstered acetonitrile have shown LDso's in various species as follows: Sherman rats, 3.8 g/kg (Smyth and Carpenter, 19481; rats, 1.34-6.68 g/kg (Pozzani et al., 1959~; and guinea pigs, 0.14 g/kg (Pozzani et al., 1959~. Kimura et al. (1971) found that the oral toxicity of acetonitrile varied according to the age of the rat. The acute oral LD50's for 14-day-old, young adult, and adult rats were 0.16 g/kg, 3.1 g/kg, and 3.5 g/kg, respectively. Acetonitrile was significantly more toxic in the 14-day-old rat than in the adult. There are no studies dealing with subacute or chronic toxicity of acetonitrile administered orally to animals. Carcinogenicity, Mutagenicity, and Teratogenicity There are no reports indicating any possible carcinogenic, mutagenic, or teratogenic ef- fects of acetonitrile (National Institute for Occupational Safety and Health, 1978a). Acrylonitrile (vinyl cyanide) is suspected of inducing cancer in both animals and humans (National Institute for Occupational Safety and Health, 1978c). CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level {SNARLJ 24-Hour Exposure There are no adequate data from which to calculate a 24-hour or 7-day SNARL.

206 DRINKING WATER AND HEALTH Chronic Exposure There are no adequate data from which to calculate a chronic exposure SNARL. Chloroform (CHCI3) Chloroform was evaluated in the first and third volumes of Drinking Water and Health (National Academy of Sciences, 1977, pp. 713-717; 1980a, pp. 203-204~. The following material, which became available after these volumes were published, updates and, in some instances, reevaluates the information in the earlier reports. Also included are some references that were not assessed id the original report. HEALTH EFFECTS Observations in Humans No new data. Observations in Other Species Hewitt et al. (1979) observed an increased uptake of glutamate-pyruvate transaminase (GPl) and decreased uptake of indicator organic anions and cations in kidney slices from male Swiss-Webster mice exposed to chloroform. Pretreatment of the animals with mirex did not markedly alter either the hepatotoxic or nephrotoxic effects of chloroform. However, pretreatment with kepone potentiated chloroform hepatotoxic- ity and may have increased chloroform-induced kidney damage. The authors concluded that ingestion of kepone may increase the sensitivity of the liver and kidneys to chloroform toxicity. The toxicity of chloroform was more pronounced in males than in fe- males of the following barrier-reared strains of mice: Tif:MAGf, Tif:MF2f, C3H/TifBomf, DBA/JBomf, C57BL/6J/Bomf, and A/JBomf strains. The C3H/TifBomf strain proved to be the most sensitive of the several strains tested (Pericin and Thomann, 1979~. 2,3,7,8-Tetrachloro- dibenzo-p-dioxin did not alter the acute toxicity of chloroform (Hook et al., 1978), but Cornish et al. (1977) showed that the toxicity of chloroform can be markedly potentiated by prior treatment with ethanol or phenobar- bital. Mutagenicity In studies by Agustin and Lim-Sylianco (1978), chloroform showed DNA-damaging and chromosome-breaking (clasto- genic) activity; however, it had no direct effect on base-pair and frame

Toxicity of Selected Organic Contaminants in Drinking Water 207 shift mutations. Chloroform also caused frame shift mutations in male mice after metabolic activation. Vitamin E treatment decreased the mutagenicity and clastogenicity of chloroform. Clemens et al. (1979) reported genotypic differences in responses to the toxic effects of chloroform that were manifestations of differences in renal rather than hepatic responses or the ability to repair renal damage. In these studies, the authors showed that DBA/2J male mice were more sen- sitive to the 10-day lethal effect of chloroform than were C57BL/6J males, whereas the sensitivity of B6D2F~ (C57BL/6J X DBA/2J) mice was in- termediate. This relative order of sensitivity was preserved following sublethal doses of labelled chloroform with respect to accumulation in subcellular fractions and renal (but not hepatic) dysfunction. Kidneys from mice of the three genotypes were able to repair tubular damage from chloroform. Prior to toxic exposure to chloroform, covalent binding of labeled chloroform to renal microsomes was greater in DBA than in C57BL mice. Pretreatment with phenobarbital enhanced covalent bind- ing by renal microsomes from DBA, but not from C57BL mice. Testoster- one propionate and medroxyprogesterone acetate sensitized the kidneys of both sexes in DBA/2J and C57BL/6J mice. Progesterone and hydrocorti- sone sodium phosphate sensitized the kidneys to chloroform in DBA/2J males but not in DBA/2J females or in C57BL/6J mice of either sex. In the presence of metabolically active mouse liver microsomes and bacteria, chloroform was not activated to mutagenic species (Greim et al., 1977~. It also gave negative results in the Ames test and failed to increase the frequency of sister chromatic exchange (SCE) in fibroblasts of Chinese hamsters (Li et al., 1979~. Results from a study by White et al. (1979) also indicated that chloroform did not increase SCE values. Simmon (1977) reported that bromoform, dibromochloromethane, and bromodichloro- methane are mutagenic whereas chloroform is not mutagenic in Salmo- nella typhimurium strains TA1535 and TA100. Carcinogenicity Reuber (1979) recently reviewed the literature con- cerning the carcinogenicity of 14 organochlorine pesticides in mice. Car- cinomas of the liver were observed most frequently in mice ingesting chloroform, among other organochlorine compounds. When ad- ministered orally in doses of 0.15 mg/kg/day in drinking water, chloroform did not enhance the growth or metastasis of Lewis lung car- cinoma or increase the number of Ehrlich ascites tumor cells in mice in- oculated with tumor cells (Caper and Williams, 1978~. However, when ad- ministered at 15 mg/kg/day in similar experiments it caused increases in pulmonary tumor foci after inoculation with Lewis lung carcinoma cells and in the number of Ehrlich ascites tumor cells. At both doses, it led to

208 DRINKING WATER AND HEALTH more organ invasions by B16 melanoma after inoculation with the tumor cells. When administered by oral intubation, it produced kidney tumors in male Osborne-Mendel rats (Weisburger, 1977~. A series of studies was conducted at the Huntington Research Center in England, using beagle dogs, specific pathogen-free Sprague-Dawley rats, and three stocks (C57BL, CBA, and CSI) of mice, which were given chloroform in a toothpaste base (Heywood et al., 1979; Palmer e! al., 1979; Roe et al., 1979~. The beagle dogs were given the mixture orally in gelatin capsules 6 days/week for 7.5 years, followed by a 20- to 24-week recovery period (Heywood et al., 1979~. Groups of males and females received 0.5 ml/kg/day of the vehicle (toothpaste without chloroform), and eight dogs of each sex remained untreated. The treated groups were composed of eight dogs of both sexes, each receiving chloroform in doses equivalent to 15 and 30 mg/kg/day in the toothpaste vehicle. Another group of the same size received an equivalent amount of toothpaste (0.5 ml/kg/day) without chloroform. At the end of the exposure, a small number of macroscopic and microscopic neoplasms were observed. One dog in each chloroform-treated group had a malignant tumor, but there were no tumors in the livers or kidneys of any dog. Overall, exposure to chloroform in a toothpaste base was not associated with any effects on the incidence of any kind of neoplasia. Groups of 50 cesarean-derived specific pathogen-free male and female Sprague-Dawley rats received either chloroform in doses equivalent to 60 mg/kg/day in a toothpaste base or the vehicle only by gavage 6 days/week for 80 weeks. They were then observed for as long as 15 weeks more (Palmer et al., 19791. Chloroform-treated rats of both sexes survived bet- ter than the controls, although both groups had a high incidence of non- neoplastic respiratory and renal diseases. There were consistent observa- tions of decreases in plasma cholinesterase in female rats, which were shown to be related to activity against butyrylcholine but not to acetyl-,B- methylcholine. Tumors of various sites were observed in 39~o of chloro- form-treated rats of both sexes examined histologically, compared with 38~o of the vehicle controls. There were no treatment-related effects on the incidence of liver or kidney tumors. However, histological observations of malignant mammary tumors were reported in more treated than con- trol rats, but these differences were not statistically different. In another study, mice were given chloroform in a toothpaste base by gavage or in arachis oil in doses up to 60 mg/kg/day, 6 days/week for 8 weeks (Roe et al., 1979~. Control groups were left untreated or given the vehicle only. In general, there were more survivors in the chloroform- treated groups than in the control group. Treatment was not associated with any type of neoplasia. In male, but not female, ICI mice receiving

Toxicity of Selected Organic Contaminants in Drinking Water -209 doses of 60 mg/kg/day, but not in those given 17 mg/kg/day, chloroform in a toothpaste base was associated with an increased incidence of epi- thelial tumors of the kidney. A more pronounced effect of the same kind was observed in mice given chloroform at 60 mg/kg/day in an arachis oil vehicle. This treatment was also associated with a higher incidence and severity of nonneoplastic renal disease. At the dose levels tested, namely 113 and 400 times greater than the average human exposure resulting from the use of toothpaste containing 3.5% chloroform, no adverse effects were seen in the liver and there was no increased incidence of liver tumors, even in the CBA strain with the greatest susceptibility to liver tumor for- mation. At the 17 mg/kg/day level, which is 113 times greater than the average exposure of humans from toothpaste, no excess of renal tumors was observed in males of the particularly susceptible ICI strain. Teratogenicity No new data. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARLJ The following calculations are for noncarcinogenic effects only. 24-Hour Exposure A SNARL of 22 mg/liter was calculated in Drink- ing Water and Health, Volume 2 (National Academy of Sciences, 1980a). Details of the calculation are contained in that volume. 7-Day Exposure This was calculated to be one-seventh (3.2 mg/liter) of the 24-hour SNARL. Chronic Exposure form is a carcinogen in animals. This value cannot be calculated because chloro 1,2-Dibrom~3-chloropropane (C3H5Br2CI) 1,2-Dibromo-3-chloropropane (DBCP) is an amber-brown liquid with a low vapor pressure and solubility (O.l~o w/w), but it is miscible with aliphatic and aromatic hydrocarbons and other solvents and oils. Its molecular weight is 236.36. Since the 1950's, DBCP has been used as a soil fumigant and nemato- cide in emulsifiable concentrates, liquid concentrates. powders, granules, and other formulations. It has been marketed under such trade names as Negmagon, Fumazone, Nemaset, Nematox, and Nemafume.

210 DRINKING WATER AND HEALTH DBCP is produced primarily by the bromination of allyl chloride at room temperature. Technical grade DBCP has been shown to contain up to loo impurities, including epichlorohydrin and allyl chloride. The levels of these compounds and 14 other identified impurities may vary among batches. In contrast to the other impurities' epichlorohydrin is added intentionally as a stabilizer. HEALTH ASPECTS Observations in Humans In August 1977, at the request of the Oil, Chemical, and Atomic Workers Union (OCAW), the National Institute for Occupational Safety and Health (NIOSH) inspected manufacturing facilities in California. In September of that year. an emergency temporal standard for exposure to DBCP was issued by the Occupational Safety and Health Administration (OSHA). A permanent standard for DBCP exposure, which was established on March 17, 1978, limits employee exposure to 1 ppb as an 8-hour time- weighted average concentration. This OSHA standard also prohibited eye and skin contact and provided for monitoring of employee exposure, engineering controls, safe work practices, and various other regulatory re- quirements. A rebuttable presumption against registration (RPAR) and continued registration of pesticide products containing DBCP was issued by the U.S. Environmental Protection Agency (EPA) on September 22, 1977 (U.S. Environmental Protection Agency, 1977~. Both the OSHA standard and the EPA RPAR were based primarily on toxicological findings in animals, suggesting that DBCP causes sterility and is carcinogenic. The NIOSH recommendation for an occupational exposure standard for DBCP was based not only on its carcinogenic and sterilizing potential, but also on its ability to cause diminished renal func- tion, degeneration and cirrhosis of the liver, and mutagenesis in chroni- cally exposed employees. Observations in Other Species Acute Effects Studies by Torkelson et al. ( 1961 ) and by Rakh- matullaev (1971) indicate that the acute oral LDso for DBCP in rats and other animals such as guinea pigs, mice, rabbits, and chickens ranges from 170 to 350 mg/kg. Mice and rabbits appear to be somewhat less sen- sitive than chickens. Effects preceding death following lethal exposures to DBCP included depression, analgesia, skeletal muscle incoordination, and paralysis. DBCP produced slight irritation when applied to the eye,

Toxicity of Selected Organic Contaminants in Drinking Water 211 but corneal irritation was not noted. Dermal application of DBCP produced only transient erythema in rabbits, but both single and repeated dermal exposure produced subcutaneous necrosis with polymorphonuclear leucocyte infiltration. Using a modified Draize technique, Torkelson et al. (1961) reported a dermal LDso of 1,400 mg/kg in rabbits exposed for 24 hours to undiluted DBCP at 60 ppm. Inhalation exposure to vaporized DBCP produced respiratory irritation, apathy, and ataxia. Clouding of the lens and cornea and mortality occurred at higher concentrations. A 1-hour LC50 of 368 ppm was observed in rats, but delayed deaths and kidney pathology were noted following inhalation exposures to concentra- tions as low as 50 ppm. Subchronic Effects Torkelson et al. (1961) fed diets containing DBCP concentrations of 0, 5, 20, 150, 450, and 1,350 mg/kg to male and female rats for 90 days. Female rats exhibited increased kidney weights at )() _ , . , ~ . . . . . . . . . 111~/~g alla a~ ~llg~ler aoses, retarded welgnt gain at doses of 150 mg/kg and higher, and increased liver weight at 450 mg/kg. There were no Growth effects in male rats fed diets oontninina lo then An m~r/lrn ~, ~^-~ 4 · ~ ~A A-459 / ~,5~ ~ ~ ~ ~ ~ _ ~ ~ ~ . · · . . .. . ~ ulstologlca1 effects were minimal at all dosage levels in both sexes. In contrast to the diet exposure studies, repeated inhalation exposure produced poor growth, increased susceptibility to secondary infection, and both gross and histological changes in the testes of male rats at the lowest exposure rate tested (50 exposures of 7 hours daily, 5 days a week to DBCP at 5 ppm). Hicher concentrations of DBCP produced mortality _ ~ _ ~ ,,, . · · . ^. . and similar findings were observed in monkeys, guinea pigs, and rabbits. Intramuscular injection of testosterone, cortisone, and ACTH (adrenocor- ticotropic hormone) failed to protect male rats against the testicular ef- fects of exposure to DBCP (Torkelson et al., 19611. Mutagenicity DBCP was shown to be directly mutagenic to Salmo- nella typhimurium TA1530 and Escherichia cold Poll A by Rosenkrantz (1975~. Prival et al. (1977) reported that the compound is a direct weak mutagen in Salmonella TA1535. Blum and Ames (1977) reported mutagenic effects with DBCP in Salmonella TA100 with metabolic activa- tion. The observation of Vogel and Chandler (1974) that 1,2-dibromo- propane was also mutagenic in Drosophila could have been due to the ability of vicinal 1,2-dibromides to rearrange in solution to fo'-',, reactive bromonium ion. The substitution of a chlorine atom on the third carbon of propane would probably not alter this property significantly. Thus, DBCP would be expected to be mutagenic on the basis of its 1,2-dibro- mide configuration. However, Biles et al. (1978), in a more recent reeval- uation of the mutagenicity of DBCP, suggest that most, if not all, of the

212 DRINKING WATER AND HEALTH mutagenic effects attributed to the technical grade of DBCP used in previous mutagenesis assays is due to epichlorohydrin or to other highly mutagenic contaminants in the technical product. Using Salmonella typhimurium TA1535 with and without S-9 activation, these investigators demonstrated mutagenic effects (increased revertants at doses ranging from 0 to 1,600 ~g/plate) with technical DBCP, pure epichlorohydrin, and distillates from the technical product. Pure (redistilled) DBCP did not produce mutagenic effects even at high dosage levels. On the basis of the mutagenesis assay data, these authors calculated that the mutagenic ef- fect of technical DBCP can be attributed almost entirely to the presence of the stabilizer epichlorohydrin in the technical product. They also observed that the use of S-9 activation in the mutagenesis assay eliminated the mutagenic activity of epichlorohydrin and produced a mutagenic effect from pure DBCP, suggesting the formation of a mutagenic metabolite. Allyl chloride, which is also an impurity in technical DBCP, was found to be mutagenic in these studies, but the mutagenic potency of this contami- nant is less than that of either epichlorohydrin or technical DBCP. Carcinogenicity The EPA RPAR Final Position Document for DBCP (U.S. Environmental Protection Agency, 1978) describes four studies that attribute carcinogenic effects to DBCP in laboratory animals. The first of these is the National Cancer Institute (NCI) (1978) bioassay of DBCP for possible carcinogenicity. Partial results of this study have also been published by Weisburger (1977), and preliminary observations on the car- cinogenicity of DBCP were reported by Olson et al. (1973~. Two additional studies, which were sponsored by Dow Chemical, were conducted at the Hazelton Laboratories. The fourth study was a skin bioassay study con- ducted by Van Duuren at the New York Medical Center. An earlier report by Van Duuren (1977) describes the carcinogenicity of epichlorohydrin, allyl chloride, and other DBCP-related halohydrocarbons. When the NCI carcinogenicity bioassay was initiated in 1972, DBCP was only one of several halohydrocarbons under test. It was administered by gavage 5 days a week to male and female Osborne-Mendel rats at dosage levels of 12 and 24 mg/kg for 14 weeks' after which the dosages were increased to 15 and 30 mg/kg for a period of 73 or 64 weeks, respec- tively. Similar studies with male and female B6C3F~ mice were also con- ducted at higher concentrations of DBCP. The major conclusions of this study were that, "In rats and mice of both sexes, statistically significant incidences of squamous-cell carcinomas of the forestomach occurred in each dosed group with a positive association between dose level and tumour incidence." The NCI also concluded that DBCP is carcinogenic to the mammary gland of female rats. Toxic nephropathy was observed in

Toxicity of Selected Organic Contaminants in Drinking Water 213 the treated rats and mice, but testicular damage was not prominent. Ini- tial reports from the two Dow/Hazelton studies and the Van Duuren study also indicated an increased incidence of "possible neoplasms" or "raised areas or nodules" in the forestomach of male and female rats. Squamous cell carcinoma of the forestomach of female Swiss mice was also observed in the Van Duuren study. All of these studies have been vigorously criticized for various defects in protocol, e.g., multiple agents on test in same room, inappropriateness of the control groups, study deficiencies, etc., but the conclusion of the EPA's final position document (U.S. Environmental Protection Agency, 1978) is that the carcinogenicity trigger for DBCP had not been rebutted. However, the question of whether "pure" DBCP is a carcinogen requires further study in view of the demonstration by Biles et al. (1978) that some of the adverse effects of technical DBCP can be attributed to its epichlorohydrin content. Van Duuren (1977) has also demonstrated that allyl chloride, a contaminant of technical DBCP formulations, may be carcinogenic, which could result from the presence of the known car- cinogens epichlorohydrin and glycidaldehyde in the metabolic pathway of allyl chloride. In previous NIOSH documents relating to epichlorohydrin (National Institute for Occupational Safety and Health, 1976, 1978b), evidence has suggested that epichlorohydrin produces carcinogenic effects in both laboratory animals and in humans. The 1978 NIOSH document also reports mutagenic effects (chromosomal aberrations) in humans, and the 1976 NIOSH document described the ability of epichlorohydrin to in- duce sterility in rats. Teratogenicity No data available. REPRODUCTIVE AND FERTILITY STUDIES A conventional three-generation reproduction study in mammals does not appear to have been conducted with DBCP. However, several reports describe the adverse effects on the testes and the production of sterility in DBCP-exposed humans. Torkelson et al. (1961) reported that testicular atrophy occurred in rats repeatedly exposed to DBCP vapors at 20 ppm and higher and that there was a decreased testes weight in rats exposed to DBCP vapors at levels as low as 5 ppm. Similar effects have been reported in a series of Russian papers (Faidysh, 1973; Faidysh and Avkhimenko, 1974; Faidysh et al., 1970; Rakhmatullaev, 1971; Reznik and Sprinchan, 1975~. Prolongation of the estrous cycle was also noted by Reznik and Sprinchan (1975~. Rao e! al. (1979) and Burek et al. (1979) reported inhalation studies in

214 DRINKING WATER AND HEALTH which male rats and rabbits were exposed to DBCP at levels of 0.1 and 1 ppm for 14 weeks or 10 ppm for 10 weeks. The results of these studies in- dicated that rabbits are somewhat more susceptible to the adverse testicular effects of inhaled DBCP and that the no-effect level for DBCP by this route is 0.1 ppm. Sterility in a group of pesticide formulation workers exposed to DBCP was initially observed in December 1977 (Whorton et al., 19771. These in- vestigators described azoospermia or oligospermia in 14 of 25 nonvasec- tomized men and an increased serum level of follicle-stimulating hormone (FSH) and luteinizing hormone (LH) in these individuals. Additional results of these and related studies have been reported subsequently by the same group of investigators (Biava et al., 1978; Marshall et al., 1977~. In a recent 1-year followup report on 21 of the original subjects, Whorton and Milby (1980) indicated that recovery occurred in most of the oligospermic individuals but not in the azoospermic men. Several additional epidemiological studies of DBCP-exposed workers have been initiated by the EPA, and preliminary reports from these studies are generally consistent with the Whorton study. Similar observa- tions have also been reported by Glass et al. (1979) and Potashnik et al. (19781. Two major questions remain unanswered: first, does repeated exposure to DBCP, even at low levels from which apparent recovery could occur, produce cumulative damage in the male reproductive tract and, eventu- ally, a nonrecoverable azoospermia? Second, are the adverse effects of DBCP on the testes due to DBCP itself or to a contaminant such as epichlorohydrin, which is known to cause sterility as well as mutagenic and oncogenic effects (National Institute for Occupational Safety and Health, 1978b)? CONCLUSIONS AND RECOMMENDATIONS In view of the possibility that epichlorohydrin, allyl chloride, or one of the other contaminants of technical DBCP may be responsible for the adverse toxicological effects observed in both animals and in humans exposed to DBCP, it is premature to recommend a suggested no-adverse-response level (SNARL) for pure DBCP. Dimethy~formamide [HCON(CH3~2] Dimethylformamide (DMF) has a variety of applications in industrial pro- cesses, primarily as a solvent for liquids, gases, Orlon, and similar polyacrylic fibers, and in the formulation of organic compounds. It also is

Toxicity of Selected Organic Contaminants in Drinking Water 215 used as a solvent in Spandex fiber-lengthening reactions, for polyacry- lonitrile, vinyl chloride polymers, epoxy polymers, cellulose-derived polymers, urea-formaldehyde polymers, polyamides, and paint strippers, and in the production of acrylic fibers (Zaebst and Robinson, 1977~. Its solvent properties are also useful in pigments and dyes and in gasoline anti-icing additives. DMF is a colorless liquid that is miscible with water and with most organic solvents. It is only slightly soluble in petroleum ether. It has an unpleasant, fishy odor, a vapor density of 3.7 mm mercury at 25°C, and a freezing point of-61°C (Deichmann and Gerarde, 19691. Inhalation exposures to DME are moderately hazardous to human health. The compound is moderately irritating to the skin and is a definite hazard when absorbed through the skin (Patty, 19631. There does not appear to have been injury in humans who have inhaled DMF, at least at low con- centrations (50 ppm) (Patty 19631. According to Gosselin e! al. (1976), the probable lethal dose in humans for DMF ranges from 500 mg to 5 g/kg. The current threshold limit value (TLV) is 10 ppm. IETAB OLISM During the 24 hours following oral administration of DMF to rats at 4 to 400 mg/kg, nonmetabolized DMF was detected in the blood at concentra- tions proportional to the dose given. Following DMF doses less than 200 mg/kg, the urinary concentration of N-methylformamide, the primal metabolite of DME, was greater than that of the parent compound. After DMF doses higher than 200 mg/kg, the reverse was observed. Excretion of DMF and N-methylformamide was similar following administration of DMF at 200 mg/kg. Essentially nontoxic doses are associated with greater excretion of N-methylformamide than DMF, whereas toxic doses are associated with higher concentrations of DMF than N-methylformamide in urine (Sanotskii et al., 1978~. Krivanek et al. (1978) exposed eight healthy male humans to DME vapor at a concentration of 8.79 + 0.33 ppm for 6 hours daily for 5 con- secutive days. All urine voided by the subjects was collected from the beginning of the first exposure to the 24th hour after the end of the last ex- posure. Each sample was analyzed for N-methylformamide. The in- vestigators observed that N-methylformamide was rapidly eliminated from the body, concentrations in urine peaking within a few hours after the end of each exposure period. Very little N-methylformamide was found in a sample examined 24 hours after exposure, and none of the compound was detected in a sample examined 48 hours after exposure. There was no increased excretion of N-methylformamide in the urine

216 DRINKING WATER AND HEALTH following repetitive exposure. The mean for the 7-hour (end of exposure) sample was 4.7 ,ug/ml, or 436.8 fig total. Lower and upper one-sided 95% tolerance limits for 95% of a population were 1.2 ,ug/ml (367 ,ug total) and 13.9 ,ug/ml (1,625 ,ug total). The coefficient of variation (CV) for the micrograms of N-methylformamide per milliliter was approximately 25 times more variable than the CV for total micrograms (Krivanek et al., 19781. Maxfield et al. (1975) reported a study of workers occupationally exposed to DMF. Urine samples were collected at the beginning and at the end of the workshift. Measurable amounts of DMF metabolites appeared in the urine after a single exposure to a TLV dose (10 ppm) of DMF. The metabolites were present within 2 to 3 hours after an exposure, but generally maximum concentrations were not obtained for 6 to 12 hours. Thereafter, the concentration declined and generally reached an undetect- able level 24 hours after the exposure ended. DMF was absorbed through the skin and lungs when men were exposed to TLV concentrations of the vapor. A series of studies dealing with the metabolism of DMF in rats, dogs, and humans was conducted by Kimmerle and Eben (1975a,b) and Eden and Kimmerle (1976~. In acute and subacute inhalation studies, rats and dogs were exposed to DMF. Dogs were also subjected to subchronic in- halation tests. The rats were subjected to DMF concentrations of 21, 146, or 2,005 ppm during a 3-hour trial. During a 6-hour exposure, they were exposed to DMF at 29 or 170 ppm. When repeated exposures were ad- ministered to rats, the average concentration was 350 ppm. In a 6-hour test, dogs were exposed to 20, 32, 143, or 172 ppm. In a 5-day test (6 hours/day), they were exposed to 23 or 59 ppm. In a 4-week test (6 hours/day), the concentration was 20 ppm. In addition to the known metabolite N-methylformamide, another metabolite, formadine, was found in the urine of rats and dogs. The elimination rates for N-methyl- formamide and formadine were slower in dogs than in rats. An accumula- tion of N-methylformamide was detected in the blood and urine of dogs subjected to repeated inhalation exposures to DMF at 59 ppm. In rats ex- posed to substantially higher subacute concentrations of DMF (350 ppm), this phenomenon was not observed. In the 20-ppm repeat study, the dogs in the subacute and subchronic groups showed no sign of accumulating metabolitets). The results of the liver and kidney function tests in dogs ex- posed to DMF at 20 ppm for 4 weeks were also normal (Kimmerle and Eben, 1975a). In an acute inhalation test, Kimmerle and Eben (1975b) exposed four persons to DMF at concentrations of approximately 26 or 87 ppm. Another four persons were exposed daily to a concentration of approximately 21

Toxicity of Selected Organic Contaminants in Drinking Water 217 ppm for 4 hours daily on 5 consecutive days. The behavior of the DMF and its metabolites (N-methylformamide and formadine) in blood and urine was examined. DMF was no longer detectable in the blood a few hours after exposure. It was detectable in the urine only after acute ex- posure to 87 ppm. The N-methylformamide levels in the blood and liver increased continuously or remained for several hours at the level recorded immediately after inhalation. N-Methylformamide was detected in the urine 4 hours after exposure to DMP. Most of the substance was elimi- nated within 24 hours. Elimination of forrnadine, however, was delayed. After workers were exposed repeatedly to the maximum allowable concen- tration in the working area, there was no accumulation of N-methylforrna- mide in the blood or urine. The investigators suggested that determination of N-methylformamide in a 24-hour urine sample be used as a routine monitoring procedure for employees exposed to DMF. Concentrations above 50 ,ug in a 24-hour urine sample indicated that exposure exceeded the maximum allowable concentration of 10 ppm (Kimmerle and Eben, 1975b). Delayed catabolism of DMF was observed in rats and dogs after oral administration of ethanol prior to inhalation exposure to DMF. This was also true for humans. During repeated exposure and daily pretreatment with ethanol, the metabolism of DMF was inhibited. The metabolism of ethanol was also influenced by the presence of DMF (Eben and Kim- merle, 19761. DMF was absorbed through the skin and lungs when workmen were ex- posed voluntarily to its vapor (Maxfield et al., 1975~. Application of DMF; to the skin also resulted in absorption. After these exposures, N-methyl- formamide appeared in the urine in small but measurable concentrations. The estimated dose recovered ranged from 0.5% to lo of the adminis- tered DMP. Individual variation in the metabolitefs) excreted was marked, probably resulting from difference in the actual amount absorbed and such other factors as the rate of metabolism and of renal clearance. Muravieva et al. (1975) have also shown that free DMF appears in the blood and urine regardless of the route of administration (i.e., via the gas- trointestinal tract, by inhalation, or by contact with the skin). Oral administration of ethanol to rats at 2.0 g/kg before inhalation ex- posures to DMF at 87, 104, or 209 ppm for 2 hours only or for 2 hours daily for S days was investigated by Eben and Kimmerle (1976~. They also ex- amined the effects of DMF in dogs subjected to 2-hour inhalation ex- posures to 210-240 ppm. In both experiments there was a delayed effect on the demethylation of DMF. A dose of ethanol at 2 g/kg before inhala- tion did not influence DME metabolism in rats. The metabolism of N-methylformamide was inhibited by ethanol.

218 DRINKING WATER AND HEALTH In humans exposed to SO to 80 ppm DMF for 2 hours prior to oral ad- ministration of 19 g/person ethanol, the blood level of DMF did not in- crease. However, N-methylformamide concentrations were slightly lower and were not detected in the blood immediately after exposure. Excretion of formamide in the urine was increased by ethanol when measured 24 hours after the start of DMF administration by inhalation and was detected for up to 32 hours. The metabolism of ethanol was also influenced by the presence of DMF (Eben and Kimmerle, 19761. HEALTH ASPECTS Observations in Humans Krivanek et al. (1978) exposed eight healthy men to DMF vapor at a con- centration of 8.79 + 0.33 ppm for 6 hours daily for 5 consecutive days. All urine voided by the subjects was collected from the beginning of the first exposure to 24th hour after the end of the last exposure. Each sample was analyzed for N-methylformamide. The investigators observed that N-methylformamide was rapidly eliminated from the body via the urine and that the concentrations peaked within a few hours following the end of each exposure. Very little N-methylformamide was found in the sample collected 24 hours after the exposure ended, and none was found 48 hours after. There was no increase in N-methylformamide in the urine following repetitive exposures. The mean value for the 7-hour (end of exposure) sample was 4.7 ~g/ml, or 786.8 ,ug total. Lower and upper one-sided 95% tolerance limits for 95% of the population were 1.2 ,ug/ml (367 ,ug total) and 13.9 ,ug/ml (1,625 ,ug total) (Krivanek et al., 19781. Lyle and his colleagues (1979) reported facial flushing and other symp- toms in 19 of 102 men who worked with DMF. The highest concentration of DMF measured in the air of the workplace was 200 ppm. Twenty-six of the 34 episodes occurred after the workers had consumed alcoholic drinks. The metabolite N-methylformamide was detected in the urine of these workers on 45 occasions, the highest concentration recorded at 75 ,ul/liter. The DMF-ethanol reaction is possibly attributable to the inhibi- tion of acetaldehyde metabolism, probably by N-methylformamide. Stamova et al. (1976) examined workers occupied in the production of polyacrylonitrile fibers near Burgas, Bulgaria. There was a tendency toward elevation of morbidity due to neuroses and diseases of the peri- pheral and autonomic nervous system, stomach and duodenum, and skin. There was a neurasthenic syndrome in 41.5% of the patients as well as skin changes such as toxic dermatitis, itching dermatosis, and chronic eczema. No definite correlation between exposure to DMF and these out

Toxicity of Selected Organic Contaminants in Drinking Water 219 breaks was established. However, Chary (1974) presented evidence to sup- port the presence of acute pancreatitis in two patients exposed to DMP. Symptoms included upper abdominal pain, nausea, vomiting, and ery- thema of exposed parts. Potter (1973) reported that an accidental dermal and respiratory exposure to DMF by a single patient produced severe ab- dominal pain, hypotension, leukocytosis, and hepatic damage. Pain began 62 hours after exposure to DMF and was associated with positive results obtained from a Watson-Schwartz test for urine porphobilinogen. A 34-year old, healthy maintenance fitter spent 4 hours repairing a blocked pipe under a DMF reaction vessel. Since the worker noticed an unusual smell during the job, the concentration of DMF in the at- mosphere was measured and found to be 30 ppm. During the next several days, the worker developed dyspnea, tightness of the chest, and a generalized red blotchiness of the skin after alcohol consumption (shivers, 1978~. When humans were exposed by inhalation to 10 ppm DMF vapor for 6 hours daily for 5 consecutive days, absorbed DM17 could be correlated with urinary N-methylformamide (Krivanek et al., 1977~. The rate of urinary excretion increased rapidly during the exposure period, peaked within 3 hours after the end of exposure, and was nearly zero by the begin- ning of the next exposure period. The best index of DMF exposure was the total amount of urinary N-methylformamide excreted within 24 hours. The concentrations excreted were related to the duration and level of ex- posure and the subject's physical activity. Es'kova-Soskovets (1973) observed eight volunteers exposed to DMF and styrene for 60 days. The combined effect of DMF and styrene absorbed through the skin resulted in inhibited activity of the hematopoietic system, e.g., reduced content of mature formed elements in the peripheral blood and shifts in the protein fraction of the blood plasma, indicating an un- favorable effect. Exposure of workers to DMF along with ac~rlonitrile and methyl meth- acrylate led to occasional eczema, toxic dermatitis, and vertigo, but the incidence of skin lesions was low (Bainova, 1975~. Delayed skin sensitivity and allergenic dermatitis were observed in 11 of the 28 workers exposed (Bainova, 19751. Observations in Other Species Acute Effects The acute oral LDso for DMF in rats is 2,800 mg/kg, while its percutaneous LDso in rats and rabbits is approximately 5,000 mg/kg (Deichmann and Gerarde, 19691. The oral LD50 in mice is 1,122 mg/kg. The acute LDSo in gerbils ranges from 3,000 to 4,000 mg/kg after

220 DRINKING WATER AND HEALTH a single intraperitoneal or subcutaneous injection or after stomach intuba- tion. The threshold limit value for DMF in air is 10 ppm (American Con- ference of Governmental Industrial Hygienists, 1971, pp. 3, 90, 71~. In rats, acute inhalation of DMF resulted in hemorrhage and edema of the lungs, hemorrhage and degeneration of the liver, and other, less severe alterations in the kidneys and heart. The greatest damage was found in the liver (Cruz and Corpino, 1978~. Subchronic and Chronic Effects Although the combined action of 5 mg/m3 inhaled DMF and a 30% aqueous solution of DMF cutaneously applied was negative, 10 mg/m3 inhaled DMF and a 60% aqueous solu- tion of DMF disturbed phagocytosis and decreased glycogen content in neutrophils of rats in a chronic test (Medyankin, 1975~. Cruz and Corpino (1978) reported myocardial changes in rats after exposure to DMF vapor for 0.5 hour daily for 30 days followed by sacrifice immediately or 60 days after the end of the treatment. The effects were related to the concentra- tion and length of exposure, neither of which were reported. Chronic LDso's for DMF in female Mongolian gerbils ranged from 3,800 to 4,000 mg/kg within 3 to 6 days after consuming DMF in drinking water at 34,000 to 66,000 mg/liter and ranged from 90,000 to 100,000 mg/kg within 80 to 200 days after consuming DMF in drinking water at 10,000 to 17,000 mg/liter (Llewellyn et al., 1974~. Mutagenicity DMF was negative in the Ames Salmonella typhi- murium assay both with or without metabolic activation (McCann et al., 1975~. Thus, it is recommended as a carrier solvent for compounds being tested in the Ames system that are not water soluble. DMF was also negative in a transplacental host-mediated hamster cell culture system (Quarles et al., 1979~. Williams and Laspia (1979) were unable to induce DNA repair in a hepatocyte primary culture/DNA repair system with DMF. DMF treatment of a cell culture established from a transplantable murine rhabdomyosarcoma induced morphologic differentiation and caused a marked reduction in the tumorigenicity of the sarcoma cells. Fourteen of 17 CE/J mice receiving injections of inducer-treated cells did not develop tumors after 6 months, whereas all 21 mice receiving inocula of untreated sarcoma cells died of the disease between the 11th and 31st day. The drug-treated cells did not grow in soft agar. The untreated tumor cells grew in the semisolid medium, showed a reduced serum requirement, and had a higher saturation density than did the drug-treated cells. Thus, the reduction in tumorigenicity of DMF-treated cells correlates with cer

Toxicity of Selected Organic Contaminants in Drinking Water 221 fain in-vitro growth properties that are more characteristic of normal, mesenchymally derived cells than of sarcoma cells (Dexter, 1977~. Borenfreund and colleagues (1975) reported that the combined action of DMF and 5-bromodeoxyuridine reversed the random, irregular growth pattern of cultured tumor cells to a regular pattern typical of nonmalig- nant cells. The effect was eliminated by removal of the DMP, which altered the components of the surface membrane glycoprotein whereas 5-bromodeoxyuridine did not. The morphologic alterations induced-by 5-bromodeoxyuridine were reversed by excess thymidine, while those in- duced by DMF were not. Furthermore, DME appeared to act on the cell membrane, whereas 5-bromodeoxyuridine may interact directly with the genetic material. Reproductive Effects and Teratogenicity When applied to the skin of pregnant rabbits during organogenesis, DMF induced slight embryo mor- tality but had no apparent teratogenic effects (Stula and Krauss, 19771. In another study, Filimonov et al. (1974) observed that DMF produced no ef- fect on the function of the maternal heart during acute asphyxia when DMF was inhaled by rats in concentrations of 400 mg/m3 for 5 to 7 hours at various stages of gestation or during the entire gestation period. Sheveleva et al. (1979) reported that inhalation of DMF at 10.7 mg/m3 for 4 hours daily for 20 days increased embryonic mortality in pregnant rats, disturbed the estrous cycle in nonpregnant rats, and inhibited urinary elimination of hippuric acid after loading with sodium benzoate. No mutagenic effects or alterations in the functioning of the testes were reported. Scheufler (1976), however, reported that DMF was teratogenic in the rat, but only after repeated administration. Scheufler and Freye (1975) observed that DMF was embryotoxic and teratogenic in pregnant mice following intraperitoneal injections for 10 or 14 days during gesta- tion. Four percent of the pups were deformed, the major defects occurring in the occipital area of the osseous skull. Kimmerle and Machemer (1975) reported that fetal development was not influenced by exposure of pregnant rats to DMF at approximately 18 ppm; however, fetuses taken by cesarean section from dams exposed to 172 ppm weighed significantly less than those from control dams. Skeletal development of these fetuses was normal, and all other reproductive parameters were within the normal range for this strain of rat. The authors concluded that inhalation of DME in concentrations up to ap- proximately 10 times more than the maximum allowable concentration was not teratogenic in rats.

222 DRINKING WATER AND HEALTH Carcinogenicity No available data. CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level {SNARL) 24-Hour Exposure There are no adequate data from which to cal- culate this SNARL. 7-Day Exposure There are no adequate data from which to calculate this SNARL. Chronic Exposure This SNARL cannot be calculated since there is a lack of adequate chronic exposure data. Nitrobenzene (C6H5NO2) This compound was reviewed previously by the Criteria and Standards Division, Office of Water Planning and Standards, U.S. Environmental Protection Agency (1979c). The committee reviewed and discussed that document for accuracy and completeness. In this section, the committee summarizes the pertinent data in that document and provides additional material where necessary. Ninety-seven percent of nitrobenzene is used to reduce aniline, which has wide application in the manufacture of dyes, rubber, and medicines. Nitrobenzene is also an intermediate in the production of explosives and organic chemicals and is used as an industrial solvent, a combustible pro- pellent, a paint solvent, and an ingredient of perfumes and shoe and metal polishes (Dorigan and Hushon, 1976~. It is produced by the nitration of benzene with nitric and sulfuric acids. Annual production of Nitrobenzene in the United States ranges from 90,000 to 315,000 metric tons (Dorigan and Hushon, 1976; Lu and Met- calf, 1975~. The greatest loss of Nitrobenzene during production (estimated at 3,600 metric tons annually) occurs in the effluent wash (Dorigan and Hushon, 1976~. The compound may also form spontane- ously in the atmosphere from the photochemical reaction of benzene with oxides of nitrogen (Dorigan and Hushon' 1976~. Nitrobenzene, also referred to as nitrobenzol, oil of mirbane, or oil of bitter almond, is a pale yellow, oily liquid with an almond-like odor. It has a melting point of 6°C, a boiling point of 210-211°C, and at 25°C its vapor pressure is 0.340 mm of mercury (Windholz et al., 19761. It is slightly soluble in water, 0.1 per 100 parts of water (1,000 mg/liter) at

Toxicity of Selected Organic Contaminants in Drinking Water 223 20°C (Kirk and Othmer, 1967~. Its odor is detectable in water at concen- trations as low as 30 ,ug/liter (Austere et al., 1975~. The threshold limit value (TLV) of nitrobenzene in the air of industrial plants is set at 5 mg/m3 (1 ppm) (American Conference of Governmental Industrial Hygienists, 1980~. METABOLISM Nitrobenzene is a very lipid-soluble compound, having an oil to water par- tition coefficient of 800. In a study conducted in rats, the ratio of the con- centration of nitrobenzene in adipose tissue to that in the blood of internal organs and muscles was approximately 10:1 1 hour after an intravenous dose (Piotrowski, 1977~. In rabbits receiving an oral dose of 0.25 ml of nitrobenzene, 50~o of the compound had accumulated unchanged in tis- sues within 2 days after the intubation (Dorigan and Hushon, 1976~. In humans, the rate of nitrobenzene turnover is sufficiently slow to result in its accumulation in tissues under conditions of daily exposure (Piotrowski, 1967~. Nitrobenzene is metabolized via two main pathways: reduction to aniline followed by hydroxylation to aminophenols and direct hydroxyla- tion of nitrobenzene to form nitrophenols. Further reduction of nitro- phenols to aminophenols may also occur (Piotrowski. 1977~. Reduction of nitrobenzene to aniline occurs via the unstable intermediates nitro- sobenzene and phenyl hydroxylamine, both of which are toxic and have pronounced ability to form me/hemoglobin. The reductions occur in the cytoplasmic and endoplasmic reticulum regions of the cell catalyzed by the nitroreductase enzyme system (routs and Brodie, 1957~. The resulting aniline is then acetylated to acetanilide or is hydroxylated and excreted as aminophenol. Metabolism of nitrobenzene to nitroso and hydroxylamino derivatives by microsomal nitroreductase apparently plays an important role in methemoglobin formation, since treatment of rats with inhibitors such as SKF-525A causes a significant reduction in methemoglobin levels after dosing with nitrobenzene (Kaplan and Khanna, 1975~. However, in- duction of microsomal enzymes by phenobarbital and similar inducers decreases nitrobenzene-induced methemoglobinemia, presumably as a result of greater rates of nitrobenzene metabolism and excretion (Kaplan et al., 19741. Reddy et al. (1976) showed that the gut flora of rats also con- tributed to the reduction of nitrobenzene and the subsequent formation of me/hemoglobin. Robinson et al. (1951a,b) studied the metabolism of t4C-nitrobenzene in the rabbit. Approximately 55~o of the dose was excreted as metabolites in the urine during the 2 days after dosing 20% in the form of nitro com

224 DRINKING WATER AND HEALTH pounds and 35~o as amino compounds. The nitro compounds found in urine were nitrobenzene, o-nitrophenol, and 4-nitrocatechol in very small amounts, and m- and p-nitrophenol in relatively large amounts. Amino compounds included the major (355to ~ urinary metabolite p-aminophenol as well as o- and m-aminophenol. All of the phenols were excreted as either glucuronide or sulfate conjugates. Approximately loo of the dose was ex- pired from the rabbits as SCOT. Parke ~ 1956) investigated the fate of ~4C-nitrobenzene in rabbits. Radioactivity was distributed as follows: loo in respiratory carbon diox- ide, 60~o in the urine, Do in the feces, and between limo and 20~o in the tissues. Unchanged nitrobenzene was eliminated by expiration (0.5~o) and in the urine ~ < 0.1 To ). Major urinary metabolites included p-aminophenol (alto), m-aminophenol (Woo), o-aminophenol (ado), aniline (0.3~o), o-nitrophenol (O.l~o), m-nitrophenol (95r0), p-nitrophenol (ado), 4-nitrocatechol (0.7~o), nitroquinol (O.l~o), and p-nitrophenol mercapturic acid (0.3~o). HEALTH ASPECTS Observations in Humans Nitrobenzene is readily absorbed by contact with the skin, inhalation of the vapor, or by ingestion. There are reports of nitrobenzene poisoning resulting from ingestion of synthetic almond oil in baked products and toothache relievers applied to the gums and from its contamination of alcoholic drinks and food (Nabarro, 1948~. Leader (1932) reported a case of nitrobenzene poisoning in a child who was given `'oil of almonds" for relief of a cold. Acute nitrobenzene poison- ing has also resulted from ingestion of denatured alcohol (Donovan, 1920; Wirtschafter and Wolpaw, 1944) and from its use as an abortifacient (von Oettingen, 1941~. Nitrobenzene is also readily absorbed through the lungs, which can re- tain up to 80~o of the chemical (Piotrowski, 1967; Salmowa et al., 1963~. Poisoning has occurred after inhalation of a solution containing the com- pound, which was sprayed on a child's mattress to exterminate bedbugs (Nabarro, 1948; Stevenson and Forbes, 1942), and after inhalation of nitrobenzene used as a scent in perfume and soap (Dorigan and Hushon, 1976~. Chronic and acute poisoning from exposure to nitrobenzene fumes in production plants are well documented (Browning, 1950; Dorigan and Hushon, 1976; Hamilton, 1919; Zeligs, 1929~. In an 8-hour workday, a worker exposed to the TLV for nitrobenzene of 5 mg/m3 (1 ppm) could absorb 24 mg through the lungs (Piotrowski, 1967~. Since nitrobenzene is

Toxicity of Selected Organic Contaminants in Drinking Water 225 also absorbed through the skin, industrial poisoning cannot be attributed to inhalation alone. Nitrobenzene is highly fat-soluble and can be absorbed through the human skin at rates as high as 2 mg/cm2/hour (Dorigan and Hushon, 1976~. The medical literature contains many reports of poisoning from ab- sorption of nitrobenzene in shoe dyes and laundry marking ink. These reports were common during the 19th century and the first half of this century. Poisoning following the wearing of newly dyed wet shoes were reported as early as 1900 (Levin, 1927~. The poisoning can result from nitrobenzene or aniline, both of which were used in shoe dyes and cause the same toxic symptoms. There have been reports of shoe dye poisoning in an army camp (Levin, 1927), in children who were given freshly dyed shoes (Graves, 1928; Levin, 1927; Zeitoun, 1959), and in adults. Gener- ally, the affected people were brought to the physician's attention because of dizziness, bluish lips and nails (cyanosis), headache, and sometimes coma. All of these effects were due to methemoglobin formation from the absorbed nitrobenzene and/or aniline. Cyanosis and poisoning of newborn who came into contact with diapers or pads containing marking ink were also very common. This generally oc- cu~Ted when the diapers or pads were freshly stamped by the hospital laundry (Etteldorf, 1951; MacMath and Apley, 1954; Ramsay and Har- vey, 1959; Zeligs, 1929~. The toxicity was often severe in premature in- fants who were in an incubator and surrounded by fumes as well as the dye in the cloth (Etteldorf, 1951~. General absorption of nitrobenzene is the cause of many of the chronic and acute toxic effects observed in nitrobenzene workers. The amount of cutaneous absorption is the function of the ambient concentration, the amount of clothing worn, and the relative humidity, which increases ab- sorption as it becomes higher (Dorigan and Hushon, 1976~. A worker ex- posed to the TLV of 5 mg/m3 could absorb up to a total of 33 mg/day, ap- proximately 9 mg of which is absorbed cutaneously (Piotrowski. 1967~. Pacseri and Magos (1958), who measured ambient nitrobenzene in in- dustr~al plants, found levels up to 8 times higher than the current TLV. There is a latent period of 1 to 4 hours before signs and symptoms of nitrobenzene poisoning appear. The chemical affects the central nervous system, producing fatigue, headache, vertigo, vomiting, and general weakness. In some cases, severe depression, unconsciousness, and coma also result (Browning, 1950; Hamilton, 1919; Pacseri and Magos, 1958; Piotrowski, 1967, 1977~. Nitrobenzene is a powerful methemoglobin former, cyanosis appearing when methemoglobin levels reach 15~o. Blood methemoglobin levels from nitrobenzene have ranged from 0.6 g/lOO ml in industrial chronic exposures to l O g/ l OO ml in acute poisoning (Myslak

226 DRINKING WATER AND HEALTH et al., 1971; Pacseri and Magos, 19581. Normal methemoglobin levels are approximately 0.5 g/100 ml. Chronic exposure to nitrobenzene may lead to spleen and liver damage, jaundice, liver impairment, methemoglo- binemia, sulfhemoglobinemia, dark-colored urine, anemia. the presence of Heinz bodies in erythrocytes. and hemolytic icterus. Reported fatal doses of nitrobenzene in humans have varied widely- from less than 1 ml to more than 500 ml (Dorigan and Hushon, 1976; von Oettingen. 1941; Wirtschafter and Wolpaw, 1944~. A minimum lethal dose of 4.0 ml has been reported by von Oettingen (1941~. For a 70-kg adult, this would approximate 69 mg/kg. Metabolic transformation and excretion of nitrobenzene and its metab- olites in humans are slower by an order of magnitude than they are in rats and rabbits. Measurements of nitrobenzene concentrations in the blood of an acutely exposed person indicate that the compound is cumulative, re- maining in the human body for a prolonged period (Piotrowski, 1967, 1977~. There have been similar observations of persistence of the two ma- jor urinary metabolites' p-aminophenol and p-nitrophenol, in patients treated for acute or subacute poisoning. Because of the slow rate of nitrobenzene metabolism in humans, the concentration of p-nitrophenol in urine increases for approximately 4 days during exposure, eventually reaching a value 2.5 times that found during the first day. The half-life for the urinary excretion of p-nitrophenol from humans after a single dose was approximately 60 hours (Salmowa et al., 19631; 84 hours was observ- ed in a female who attempted suicide (Myslak et al., 1971~. The amount of nitrobenzene to which an individual has been exposed can therefore be estimated from the level of total (free and conjugated) p-nitrophenol in urine (Ikeda and Kita, 1964; Piotrowski, 19771. The urinary metabolites in humans account for only 20% or 30% of the nitrobenzene dose. The fate of the remaining nitrobenzene and its metabolites is not known (Piotrowski, 1977~. Ikeda and Kita (1964) measured the urinary excretion of p-nitrophenol and p-aminophenol in a patient admitted to a hospital with toxic symp- toms resulting from a 17-month chronic industrial exposure to nitro- benzene. The rate of excretion of these two metabolites was similar and paralleled the level of methemoglobin in blood. Observations irl Other Species Acute Effects The acute oral LDso for nitrobenzene in the rat is 640 mg/kg, while the LDso in rats for intraperitoneal and percutaneous doses are 640 and 2,100 mg/kg, respectively (Fairchild, 1977b). The lowest

Toxicity of Selected Organic Contaminants in Drinking Water 227 published lethal oral doses in the dog, cat, and rabbit were 750, 2,000, and 700 mg/kg, respectively. Effects were loss of reflexes, methemoglo- binemia, tremors, paralysis of hind legs, and labored respiration. Subchronic and Chronic Effects Levin (1927) demonstrated in-vivo production of methemoglobin by nitrobenzene in dogs, cats, and rats, but not in guinea pigs or rabbits. Reddy et al. (1976) reported a delay in methemoglobin formation in germ-free rats by nitrobenzene and postu- lated that the gut flora of the rats was responsible for the in-vivo reduction and for the me/hemoglobin-forming capacity of nitrobenzene. Yamada (1958) did a chronic toxicity study in rabbits that received a subcutaneous dose of 0.8 mg/kg nitrobenzene per day for 3 months. He found a decrease in erythrocyte number and hemoglobin content early in the exposure, but these values increased during the 3 months and did not return to the normal levels. Urinary excretion of detoxification products was variable in the early stages of the exposure, but then all the detoxifica- tion reactions (reduction, hydroxylation, and assimilation) were de- pressed. As a result of these observations, Yamada divided the response of rabbits to nitrobenzene into three stages: initial response, resistance, and exhaustion. Daily subcutaneous exposure of rats to nitrobenzene at doses of 50 mg/kg for 1 month had no effect on nitroreductase or aniline hydroxylase activity or length of hexobarbital-induced sleeping time, but the antipyretic effect of phenacetin was decreased (Wisniewska-Knypl et al., 1975). Mutagenicity Chiu et al. (1978) tested nitrobenzene using the Ames assay with Salmonella typhimurium strains TA98 and TA100. They found the compound not to be mutagenic. Since other nitrobenzene derivatives demonstrated mutagenicity in in-vitro assays, however, additional testing of the parent compound may still need to be done. Carcinogenicity No available data. Nitrobenzene is currently undergo- ing testing for carcinogenicity in the National Toxicology Program. Reproduction and Teratogenicity Kazanina (1968a,b) administered nitrobenzene in doses of 125 mg/kg/day subcutaneously to pregnant rats during preimplantation and placentation. Delay of embryogenesis, altera- tion of normal placentation, and abnormalities in the fetus were observed. Gross morphologic defects were seen in 4 of the 30 fetuses examined. There were changes in the tissues of the chorion and placenta of pregnant women whose work involved rubber catalysis with n*robenzene. Men l

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Toxicity of Selected Organic Contaminants in Drinking Water 229 strual disturbances after chronic exposure to nitrobenzene have also been reported (Dorigan and Hushon, 1976~. CONCLUSIONS AND RECOMMENDATIONS The toxicity of nitrobenzene is great, and there is a pressing need for ade- quate chronic toxicity data, from both short-term studies and longer-term studies in several animal species. Similarly, the compound has not been adequately tested for carcinogenicity, mutagenicity, reproductive effects, or teratogenicity. Current data on human toxicity and acute and subacute toxicity of nitrobenzene are summarized in Table VII-2. Suggested No-Adverse-Response Level {SNARLJ 24-Hour Exposure The minimal lethal dose for humans appears to be 69 mg/kg (von Oettingen, 19411. Since the water solubility of nitro- benzene is approximately 1,000 mg/liter, humans could theoretically con- sume a lethal dose in an acute spill situation. However, the minimal detectable odor level in water is quite low (30 ,ug/liter). The minimal ef- fect level for humans is reported to range between 0.010 and 0.026 mg/kg (von Oettingen, 1941~. Using this value, a safety factor of 10, and assum- ing that 2 liters of drinking water daily provide the only source during this period to a 70-kg human, one may calculate the 24-hour SNARL as: 0.010 mg/kg X 70 kg = 0.035 mg/1iter. 2 liters X 10 7-Day Exposure No data are available for calculation. Using the acute 24-hour SNARL of 0.035 mg/liter and dividing by 7 days, one may derive the SNARL as: 0.035 m/liter 7 0.005 mg/liter. Chronic Exposure This SNARL cannot be calculated due to a lack of adequate chronic exposure data. Nitrophenols (C6H5NO3OH) These compounds were reviewed previously by the Criteria and Standards Division, Office of Water Planning and Standards, U.S. Environmental

230 DRINKING WATER AND HEALTH Protection Agency (1979d). The committee has reviewed and discussed that document for accuracy and completeness. The following paragraphs summarize those findings and are augmented by additional information deemed necessary by the committee. There are three isomeric forms of mononitrophenol, namely 2-nitro- phenol, 3-nitrophenol, and 4-nitrophenol. 2-Nitrophenol and 4-nitro- phenol are synthesized commercially by hydrolysis of the appropriate chloronitrobenzene isomer with aqueous sodium hydroxide at elevated temperatures (Howard et al., 19761. Production of 4-nitrophenol is achieved through diazotization and hydrolysis of 3-nitroaniline (Mat- suguma, 1967~. The mononitrophenol isomers are used in the United States primarily as intermediates for the production of dyes, pigments, pharmaceuticals, rubber, lumber preservatives, photographic chemicals, and pesticides (U.S. International Trade Commission, 19761. The mono- nitrophenols may also be produced by a microbial or photodegradation of pesticides, which contain those moieties. Approximately 4,500 to 6,750 metric tons of 4-nitrophenol are produced annually (Howard et al., 1976~. Although production figures for 4-nitrophenol are not available, Hoecker et al. (1977) estimate that production is less than 450 metric tons annually. 4-Nitrophenol is probably the most important of the mononitrophenols in terms of quantities used and potential environmental contamination. Approximately 15,750 metric tons of 4-nitrophenol were used in 1976 (Chemical Marketing Reporter, 19761. Most (trio) of the 4-nitrophenol produced is used in the manufacture of ethyl and methyl parathions, but the compound has itself been used as a fungicide. Possible sources of human exposure to 4-nitrophenol can also result from the microbial or photodegradation of the parathions or from in-vivo metabolism following ingestion of parathion or other similar organophosphate insecticides. The physical and chemical properties of mononitrophenols are sum- marized in Table VII-3. Threshold concentrations for the mononitrophenols have been deter- mined by Makhinya (1964~. For odor, taste, and color, these concentra- tions were reported as follows: 2-nitrophenol, 3.83, 8.6, and 0.6 mg/liter; 3-nitrophenol, 389, 164.5, and 26.3 mg/liter; and 4-nitrophenol, 58.3, 43.4, and 0.24 mg/liter, respectively. Six isomeric forms of dinitrophenols are possible, but 2,4-dinitrophenol is the most important commercially. Approximate annual usage of 2,4-di- nitrophenol is estimated at 450 metric tons (Howard et al., 1976~. It is used primarily as a chemical intermediate for the production of sulfur dyes, azo dyes, photochemicals, pesticides, wood preservatives, and ex- plosives (Matsuguma, - 1967; Springer et al., 1977a,b). 2,4-Dinitrophenol is synthesized commercially by the hydrolysis of 2,4-dinitro-1-chloro

Toxicity of Selected Organic Contaminants in Drinking Water 231 TABLE VII-3 Properties of Mononitrophenolsa Property 2-Nitrophenol 3-Nitrophenol 4-N*rophenol Melting point Boiling point Density Water solubility at 40°C Vapor pressure at 49.3°C K b a 44-45°C 214-216°C 1.485 g/cm3 3.3 g/liter 1 mm Hg 97°C 194°C 1.485 g/cm3 30.28 g/liter 1 13-1 14°C 279°C 1.479 g/cm3 32.8 g/liter 7.5 x 10-8 5.3 X 10-9 7 X 10-8 aData from Stephen and Stephen, 1963, and Windholz et al., 1976. bDissociation constant. benzene with sodium hydroxide at 95° to 100°C (Matsuguma, 19671. Pro- duction figures and usage data for the remaining five dinitropheno] isomers are not available, but it is reasonable to assume that usage of these compounds in the United States is veer limited. The physical and chemical properties of the dinitrophenol isomers are summarized in Table VII-4. Six isomeric forms of trinitrophenol are possible, but significant com- mercial usage of these compounds is apparently limited to 2,4,6-trinitro- phenol (picric acid). Picric acid is used as an explosive, germicide, fungicide, and tanning agent, in electroplating, and in the pharmaceu- tical and chemical industries. Annual production and the extent to which picric acid is applied in any of these areas is not known. The properties of 2,4,6-trinitrophenol are summarized in Table VII-S. TABLE VII-4 Properties of Dinitrophenol Isomersa Melting Water Point, Kab X 10-5, Solubility, Density, Isomer °C 25°C g/liter g/cm3 2,3-Dinitrophenol 2,4-Dinitrophenol 2,5-Dinitrophenol 2,6-Dinitrophenol 3,4-Dinitrophenol 3,5-Dinitrophenol 144 1.3 144-1 15 10 104 0~7 27 4.2 21 63.5  134 122-123 2.2 0.79 0.68 0.42 2.3 1.672 1.6 1.702 1.681 1.683 a From Stephen and Stephen, 1963, and Windholz et al., 1976. bDissociation constant.

232 DRINKING WATER AND HEALTH TABLE VII-5 Properties of 2,4,6-Trinitrophenol (Picric Acidja Property Value Melting point Boiling point Vapor pressure Density Water solubility at 25°C 122-123°C Sublimes, explodes at 300°C 1 mm Hg at 195°C 1.763 g/cm3 1.25 g/liter . a From Stephen and Stephen, 1963, and Windholz et al., 1976. Knowles et al. (1975) have demonstrated the production of a large number of monophenols, including 2-nitrophenol, in a model system simulating gastric digestion of smoked bacon. Under these conditions, in which nitrosation is favored, their results indicate that nitrosation of phenols in smoked bacon may occur in the stomach with resultant forma- tion of 2-nitrophenol. 4-Nitrophenol may be produced in the atmosphere through the photo- chemical reaction between benzene and nitrogen monoxide. Nojima et al. (1975) irradiated a combination of benzene vapor and nitrogen monoxide gas, which resulted in the production of nitrobenzene, 2-nitrophenol, 4-nitrophenol, 2,4-dinitrophenol, and 2,4,6-dinitrophenol. The authors suggested that these nitro compounds may have affected seriously stricken victims of photochemical smog in Japan, whose symptoms character- istically included headache, breathing difficulties, vomiting, elevated body temperature, and numbness in the extremities. METABOLISM Mononitrophenols Mononitrophenols are readily absorbed by the gastrointestinal tract and rapidly excreted, primarily in the urine. Lawford et al. (1954) showed that elimination of 4-nitrophenol by the monkey following oral and intra- peritoneal doses of 20 mg/kg was complete within S hours. Excretion by mice, rats, rabbits, and guinea pigs is also rapid. Most doses were com- pletely eliminated from the blood within 2 hours a*er administration. Rates at which 4-nitrophenol disappeared from the blood decreased in the following descending order: mouse, rabbit, guinea pig, rat, and monkey. Metabolism of the mononitrophenols occurs via one of three mecha

Toxicity of Selected Organic Contaminants in Drinking Water 233 nisms, primarily by conjugation, resulting in the formation of either glucuronide or sulfate conjugates. Other mechanisms include reduction to amino compounds or oxidation to diphenols. Sulfate and glucuronide conjugation processes are major detoxification mechanisms in many species, including mammals (Quebbemann and Anders, 1973~. 4-Nitrophenol is used as a preferred substrate for analysis of UDP- glucuronyl transferase (Aitio, 1973; Litterst et al., 1975), which is local- ized primarily in the microsomal fraction of the liver of most species Sulfate conjugation of 4-nitrophenol also occurs, but is decreased during pregnancy in rabbits (Pulkkinen, 1966b) and increases with age in the rat, guinea pig, and humans (Pulkkinen, 1966a). The relative rates of glucuronide versus sulfate conjugation of 4-nitrophenol may depend on the levels of phenol present (Moldeus et al., 1976~. Robinson et al. (1951a) studied the metabolic detoxification of the mononitrophenol isomers in rabbits. These workers showed that at doses of 0.2 to 0.3 g/kg, conjugation in vivo with glucuronic and sulfuric acids was almost complete, and only small amounts (<logo) of the unchanged free phenols were excreted. With all three of the mononitrophenol isomers, the major conjugation product was nitrophenyl glucuronide, which accounted for approximately 70~o of the dose. The corresponding sulfate conjugates were also excreted. Reduction of the nitrophenols oc- curred to a small extent, and there was a small amount of additional ox- idation. Thus, 2-nitrophenol yielded traces of nitroquinol; 3-nitrophenol yielded nitroquinol and 4-nitrocatechol; and 4-nitrophenol yielded 4-nitrocatechol. The following distributions of metabolites in the rabbit were noted by Robinson et al. (1951a,b): 2-nitrophenol gave 82~o nitro compounds and 3~o amino compounds, 71% of the dose excreted as glucuronide con- jugates and 11% as ethereal sulfates; 3-nitrophenol resulted in 74% nitro compounds and 10% amino compounds in the urine, present as glucuro- nides (78370) and ethereal sulfates (19%~; and 4-nitrophenol was excreted in the urine, 87~o as nitro compounds and 14~o as amino compounds, 65% of which occurred as glucuronic conjugates and 16~o as ethereal sulfates. Dinitrophenols Dinitrophenol isomers are readily absorbed from the gastrointestinal tract and through the skin and lungs (Gehring and Buerge, 1969b; Haney, 1959; von Oettingen, 1949~. The dinitrophenol isomers are not stored to any significant extent in the tissues of humans or laboratory animals following absorption, but are readily excreted, primarily via the urine. Gehring and Buerge (1969b) have studied the elimination of 2,4-dinitro

234 DRINKING WATER AND HEALTH phenol from the serum of ducklings, mature rabbits, and immature rab- bits following intraperitoneal administration of the compound. Serum levels of 2,4-dinitrophenol in mature rabbits declined to less than 1% of the original maximum value within 7 hours. The rate of elimination is sig- nificantly reduced in immature rabbits. Lawford et al. (1954) also studied the phat~acokinetics of the other dinitrophenol isomers and showed that elimination from the blood of mice, rabbits, guinea pigs, rats, and monkeys was complete within 30 hours. Harvey (1959) determined the half-lives for all six dinitrophenol isomers from the blood of mice and rats following a single large dose given intraperitoneally (Table VII-6. Examination of the urine of a man fatally poisoned by 2,4-dinitro- phenol showed that it contained 2-amino-4-nitrophenol, 4-amino-2-nitro- phenol, and diaminophenol (Perkins, 19191. Williams (1959) stated that 2,4-dinitrophenol is excreted by mammals in the following forms: partially unchanged, partially conjugated with glucuronic acid, and reduced to 2-amino-4-nitrophenol, 2-nitro-4-aminophenol, and probably 2,4-di- aminophenol. Rats orally dosed with 2,4-dinitrophenol excreted both free dinitrophenol (78%) and 2-amino-4-nitrophenol (kilo) (Senczuk et al., 1971~. Parker (1952) examined the enzymatic reduction of 2,4-dinitro- phenol by rat liver homogenates and found 4-amino-2-nitrophenol to be the major metabolite (90%) along with 2-amino-4-nitrophenol (lO(7o) and trace amounts of 2,4-diaminophenol. In contrast, Eiseman et al. (1974) reported 2-amino-4-nitrophenol as a major metabolite (75% of total amine) in the same system. Considerably smaller amounts of 4-amino-2 TABLE VII-6 Rates at Which Dinitrophenol Isomers Are Eliminated from the Blood of Mice and Rats Following a Single Large Intraperitoneal Dosea Half-time for Elimination, min IsomerMiceRats _ 2,3-Dinitrophenol2.712.5 2,4-Dinitrophenol54.0225 2,5-Dinitrophenol3.313.0 2,6-Dinitrophenol238210 3,4-Dinitrophenol3.511.5 3,5-Dinitrophenol2.72.1 a From Harvey, 1959. Doses ranged between 20 to 180 mg/kg for mice and 20 to 90 mg/kg for rats.

Toxicity of Selected Organic Contaminants in Drinking Water 235 nitrophenol (23%) was found following enzymatic reduction by rat liver homogenates. These investigators also detected only trace quantities of diaminophenol. All six dinitrophenol isomers are uncouplers of oxidative phosphoryla- tion. 2,4-Dinitrophenol is considered to be the classic uncoupler and, hence, is widely used by biochemists. The relative potency of the 6-nitro- phenols toward the uncoupling of rat liver mitochondria was found to be in the following declining order: 3,5- > 2,4- > 2,6- = 3,4- > 2,4- = 2,5-dinitrophenol (Burke and Whitehouse, 19671. Since the order of in- vivo toxicities of the dinitrophenol isomers shown in Table VII-7 differs somewhat from that of the relative uncoupling potency reported by Burke and Whitehouse (1967), there are apparently differential rates of absorp- tion and/or the metabolism of the various isomers in vivo. Both 2,4- and 3,5-dinitrophenol inhibit porcine heart maleate dehydrogenase in vitro (Wedding et al., 1967), but at concentrations 10 to 100 times greater than those causing uncoupling. Tr~nitrophenols Autopsy examination of dogs after lethal doses of picric acid (2,4,6-tri- nitrophenol) revealed yellow staining of the subcutaneous fat, lungs, in- testines, and blood vessels (Dennie et al., 19291. The results indicate that picric acid is distributed to many tissues of the body. Dennie et al. (1929) reported the presence of picric acid in the blood and urine of dogs administered a lethal dose of the trinitrophenol, and Harris et al. (1946) reported the presence of the trinitrophenol in the urine of humans follow- ing oral exposure. In a review of the early literature, Burrows and Dacre (1975) indicated that picric acid is eliminated from humans in both the free form and as picramic acid. In perfusion experiments with liver, kidney, and spleen, the liver exhibited the strongest capacity for reduction of 2,4,6-tri- nitrophenol to picramic acid. HEALTH ASPECTS Observations in Humans Mononitrophenols Myslak et al. (1971) reported on the excretion of 4-nitrophenol from a 19-year-old female following a suicidal oral dose of nitrobenzene. Large

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Toxicity of Selected Organic Contaminants in Drinking Water 237 quantities of 4-nitrophenol and 4-aminophenol were detected in the urine, and urinary excretion of 4-nitrophenol showed a half-life of approximately 84 hours. Din itrop h en ols Numerous cases of human poisoning by 2,4-dinitrophenol have been reported in the literature. The earliest cases of fatal dinitrophenol ~ntox- ication relate to its usage as a component of explosives during World War I. For example, 36 cases of fatal occupational dinitrophenol poisoning oc- cu~Ted among the employees of the munitions industry in Prance between 1916 and 1918 (Perkins, 1919~. In a literature review, van Oettingen (1949) stated that there had been 27 reported cases of fatal occupational dinitrophenol poisoning in the United States between 1914 and 1916. Later, Gisclard and Woodward (1946) reported two fatal cases of dinitro- phenol poisoning during manufacture of picric acid, when 2,4-dinitro- phenol was produced as an intermediate. Swamy (1953) described a case of suicidal poisoning by 2,4-dinitrophenol. Lethal doses for orally ingested 2,4-dinitrophenol in humans have been reported to be 14-43 mg/kg (Sax, 1968) and 61 mg/kg (Geiger, 1933~. The toxic manifestations of 2,4-dinitrophenol exposures, which hate been reviewed by Homer (1942), include subacute signs and symptoms such as gastrointestinal disturbances (nausea, vomiting, colic, diarrhea, anorexia), profuse sweating, weakness, dizziness, headache, and loss of weight. Acute poisoning has resulted in the sudden onset of pallor, burn- ing thirst, agitation, dyspnea, profuse sweating, and hyperpyrexia. In- tense and rapid onset of rigor mortis after death has also been observed. A physician who ingested a fatal overdose of dinitrophenol, estimated to be 4.3 g, died of hyperpyrexia with rectal temperatures at death exceeding 43.3°C (Geiger, 19331. Early in the 1930's, 2,4-dinitrophenol was highly recommended for the treatment of obesity. During the first 15 months following the introduc- tion of this use, an estimated 100,000 persons took the drug for weight reduction (Homer, 1942~. Typical treatment regimen for weight control consisted of one capsule containing 75 mg of 2,4-dinitrophenol or 100 me of the sodium salt taken 3 times daily after meals (2 to 5 mg/kg/day). Despite warnings of harmful side-effects caused by disruption of in- termediary metabolism, usage of dinitrophenol became very widespread. Homer (1942) reported a total of nine deaths resulting from the use of dinitrophenol as a weight-reducing agent. In the wake of reports of cataract development in humans taking the drug, dinitrophenol was fi- nally withdrawn from use in 1937 (Homer, 1942; Parascandola, 1974~.

238 DRINKING WATER AND HEALTH Tainter et al. (1933) administered 2,4-dinitrophenol to 113 obese pa- tients for as long as 4 months without evidence of cumulative or toxic effects. The most important side-effect noted was a maculopapular or ur- ticarial skin rash, which occurred in To of the patients. Symptoms sub- sided, however, within 2 to 5 days following withdrawal of the drug. Other effects noted were loss of taste and occasional gastrointestinal upset. No effects on liver or kidney function, pulse, blood pressure, or hematological parameters were seen. In a review of the acute and chronic toxicity of 2,4-dinitrophenol in humans, Homer (1942) indicated that nausea, vomiting, and loss of ap- petite were common in patients receiving dinitrophenol. Cutaneous le- sions were the most common side-effect (Woo to 23~o), responses ranging from mild to severe. Bone marrow effects in patients treated with 2,4-dinitrophenol also occurred, and eight cases of agranulocytosis were reported, three resulting in death. There were also 30 cases of neuritis, in- cluding aberrations of taste and neurological impairment affecting, par- ticularly, the feet and legs. Symptoms appeared after an average of 10 weeks of ordinary therapeutic doses and persisted for weeks and months. Electrocardiographic evidence of heart damage in some patients was also observed. The development of cataracts following 2,4-dinitrophenol therapy was first described by Homer et al. (19351. Subsequently, more than 100 cases of 2,4-dinitrophenol-induced cataract formation were reviewed by Homer (1942~. The cataracts occurred mainly in young women, and appeared during 2,4-dinitrophenol treatment or months or years afterward. The duration of 2,4-dinitrophenol treatment and the amount consumed varied tremendously. The length of treatment varied from 3 months to 24 months, averaging 11 months. Individual susceptibility rather than the length of treatment and the total dose seemed to be the most important factors in determining cataract formation. Homer (1942) estimated that that inci- dence of cataracts in patients taking 2,4-dinitrophenol was 1870. Tr~nitrophenols Gosselin et al. (1976) reported that severe poisoning has resulted from the ingestion of 14 mg/kg picric acid (2,4,6-trinitrophenol); however, details of the poisoning episode were not provided. According to Windholz et al. (1976), ingestion or percutaneous absorption of picric acid may cause nausea, vomiting, diarrhea, abdominal pain, oliguria, anuria, yellow staining of the skin, pruritus, skin eruptions, stupor, convulsions, and death. Perkins (1919) noted that picric acid was considerably less hazard- ous than 2,4-dinitrophenol in the munitions industry during World War I.

Toxicity of Selected Organic Contaminants in Drinking Water 239 An outbreak of hematuria among U.S. Navy personnel aboard ships anchored at Wakayama, Japan, was attributed to the ingestion of picric acid in drinking water (Harris et al., 1946~. Approximately 3 weeks prior to the outbreak, more than 90 metric tons of confiscated Japanese am- munition containing picric acid had been dumped in the immediate vicin- ity of the anchorage. The trichlorophenol was apparently carried over in the vapor phase into the freshwater supply upon distillation of the seawater. Levels of picric acid ranging from 2 to 20 mg/liter were found in the drinking water. During the 1920's and 1930,s, picric acid was used both alone and in combination with butyl aminobenzoate as an antiseptic surgical dressing for the treatment of burns. A serious dysfunction of the central nervous system following the topical application of picric acid was reported by Dennie et al. (1929~. The major effect of nonlethal doses of trinitrophenol appears to be an allergic or irritative dermatitis (Anonymous, 1937; Ehrenfried, 1911~. According to Dennie et al. (1929), approximately 4~O of the patients treated with picric acid were sensitive and developed a local dermatitis, which also appeared on unexposed areas. Intense itching, burning skin eruptions, and irritability were common. More severe reac- tions led to diffuse, often severe erythema and desquamation of affected areas (American Conference of Governmental Industrial Hygienists, 1971; Sulzberger and Wise, 1933~. These reactions lasted from several weeks to almost a year after exposure (Sulzberger and Wise, 1933~. Effects on the skin are apparent when picric acid is applied in concen- trations well below those necessary for oral systemic poisoning. Of 71 in- dividuals exposed to concentrations ranging from 0.0088 to 0.1947 mg/m3, dermatitis developed only among those exposed to the lower con- centrations, indicating that desensitization or adaptation reactions may occur (American Conference of Governmental Industrial Hygienists, 1971). Observations in Other Species Mononitropherzol Acute Effects 4-Nitrophenol is the most toxic of the mononitro- phenols, followed by 3-nitrophenol and 2-nitrophenol. The acute oral LDso for 2-, 3-, and 4-nitrophenol in the rat are 2,830, 930, and 350 mg/kg, respectively (Fairchild, 1977a; Vernot et al., 1977~. In the mouse, oral LDso's were 1,300, 1,410, and 470 mg/kg (Spector, 1956; Vernot et al., 1977~. Symptoms of 4-nitrophenol intoxication in animals include methemoglobinemia, shortness of breath, and initial stimulation followed

240 DRINKING WATER AND HEALTH by progressive depression (von Oettingen, 1949~. Grant (1959), however, was unable to detect methemoglobin formation after oral administration of 3-nitrophenol and 4-nitrophenol to rats. However, Smith et al. (1967) demonstrated that 2-aminophenol and 4-aminophenol, the reduction products of mononitrophenols, would produce methemoglobin in female mice. Methemoglobin formation, therefore, may depend upon the capac- ity of the animal to reduce the mononitrophenol to amines. Din itroph en ols The acute toxicity data for the six dinitrophenol isomers are summarized in Table VII-7. Formation of cataracts resulting from acute exposure to 2,4-dinitro- phenol was first demonstrated in animals almost 10 years after the prob- lem was known to exist in humans (Bettman, 1946; Feldman et al., 1959, 1960; Gehring and Buerge, 1969a; Ogino and Yasukura, 1957~. The 2,4-dinitrophenol-induced cataracts, first produced in ducks and chickens, differ from human cataracts in that they can be formed in acute exposures and appear within less than 1 hour. Furthermore, these lesions will disappear spontaneously in animals within 25 hours (Howard et al., 1976~. Therefore, the utility of the animal model for 2,4-dinitrophenol- induced cataracts is questionable. Tr7nitrophenols The toxicity data on trinitrophenols in animals are limited to 2,4,6-trini- trophenol (picric acid). The lowest published lethal doses (LDLo) for picric acid in the rabbit and cat are 120 and 500 mg/kg, respectively (von Oettingen, 1949), while the lethal dose in the dog is reported to be 100-125 mg/kg following subcutaneous injection (Dennie et al., 1929~. After an acute lethal dose of picric acid, dogs died from respiratory paralysis (Den- nie et al., 1929~. Autopsy results demonstrated the presence of yellow staining of the subcutaneous fat, lungs, intestines, and blood vessels. Swelling of the liver and glomerulitis were also observed. Sublethal doses of picric acid less than or equal to 50 mg/kg in dogs have resulted in transitory changes in the kidney, including glomerulitis and other changes in the kidney ultrastructure. Guinea pigs develop allergic reactions following topical treatment with picric acid (Chase and Maguire, 1972; Landsteiner and DiSomma, 1940; Maguire, 1973; Maguire and Chase, 1972~.

Toxicity of Selected Organic Contaminants in Drinking Water 241 Mononitrophenols Subchronic and Chronic Effects Ogino and Yasukura (1957) reported the development of cataracts in vitamin-C-deficient guinea pigs following the administration of 4-nitrophenol. Cataracts developed in two of three guinea pigs on day 7 and 11 following daily intraperitoneal administration of 4-nitrophenol in doses of 8.3 to 12.5 mg/kg. Administration of 4-nitrophenol over a 20-day test period produced cataracts while similar dosing with 2- and 4-nitrophenol did not. However, these investigators failed to report the results obtained from the control animals neither those totally untreated nor those treated with nitrophenols and a vitamin C supplement. Thus, their results must be viewed with caution. In contrast, Dietrich and Beutner (1946) found 2- and 4-nitrophenol to be devoid of cataract-forming activity in 7-day-old chicks. Animals in this study were fed a commercial brand of chick food containing 0.25~o nitro- phenol. Cataracts developed rapidly, within 24 to 48 hours after the animals were fed 2,4-dinitrophenol, but no cataracts developed with 3 weeks in animals fed the mononitrophenol isomers. Both 2-and 4-nitrophenol have been shown to inhibit porcine heart maleate dehydrogenase in vitro (Wedding et al., 1967), and depressed rec- tal temperatures have been observed in rats receiving any of the three isomeric nitrophenols (Cameron, 19581. Thus, these compounds presum- ably are not potent uncouplers of oxidative phosphorylation, in contrast to the chemically similar 2,4-dinitrophenol. Makhinya (1969) indicated that 2-, 3-, and 4-nitrophenol possess distinct cumulative properties. Chronic administration of any of the mononitrophenols to mammals caused alterations of neurohumoral regu- lation and pathological changes including colitis, enteritis, hepatitis, gastritis, hyperplasia of the spleen, and neuritis. Limiting doses for the disruption of conditioned reflex activity were established as 0.003 mg/kg for 4-nitrophenol and 4-nitrophenol and 0.00125 mg/kg for 4-nitro- phenol. However, this reference is retrievable only in abstract form. Dinitrophenols Spencer et al. (1948) studied chronic toxicity in rats fed diets containing 2,4-dinitrophenol at 0, 100, 200, 500, 1,000, and 2,000 ,ug/g for 6 months. Both hematological and pathological investigations were con- ducted on surviving animals. Rats maintained on diets containing 2,4-di- nitrophenol at 200 Agog (10 mg/kg/day) grew at a normal rate, and no adverse changes were noted at autopsy. Pathological changes were not

242 DRINKING WATER AND HEALTH found upon microscopic examination of tissues from rats receiving diets containing 2,4-dinitrophenol at 500 ~g/g (25 mg/kg/day); however, growth of these rats was 5% to 10% less than that of controls. There was also a slight depletion of body fat and a slight increase in kidney weight. At higher doses of 2,4-dinitrophenol (50 and 100 mg/kg/day), some rats died and survivors rapidly lost weight. Enlarged dark spleens, pathological le- sions of the liver and kidney' and testicular atrophy were also observed. Trinitrophenols - No subchronic or chronic toxicity data in animals are available. Mutagenicity Szybalski ( 1958) obtained negative results in tests to determine the ability of the three mononitrophenol isomers to induce streptomycin independence in streptomycin-dependent Escherichia coli. 4-Nitrophenol showed negative mutagenic activity in host-mediated activ- ity in mice using Salmonella tvphimurium and Serratia marcescens in- dicator organisms (Buselmaier et al.. 1973~. Both 2- and 4-nitrophenol failed to induce mutations in Salmonella, both with and without microsomal activation (Chill et al., 1978; McCann et al., 1975~. However, Fahrig (1974) demonstrated weak mutagenic activity with 4-nitrophenol in tests for mutagenic gene conversion in Saccharomyces cerevisiae. This test system allows the detection of a genetic alteration whose molecular mechanism is presumably based on the formation of single-strand breaks of DNA. Adler et al. (1976) showed some evidence of DNA damage by 4-nitrophenol in wild-type and repair-deficient strains of Proteus mirabilis. There have been some reports of effects on mitosis and chromosome fragmentation in plants. Sharma and Ghosh (1965) examined the mitotic effect of mononitrophenol isomers in root tips of Allium cepa. All three compounds induced mitosis in root tips, but only 4-nitrophenol induced detectable chromosome fragmentations. Nonetheless, the data indicate that the mononitrophenols do not pose a significant mutagenic hazard to humans. A severalfold increase in mutation frequency was produced by 2,4-di- nitrophenol in tests for back mutations from streptomycin dependence to independence in E. cold (Demerec et al., 19511. Using a system that measured the induction of unscheduled DNA synthesis in testes, Fried- man and Staub (1976) found no evidence for the mutagenicity of 2,4-di- nitrophenol. Intraperitoneal injection of 2,4-dinitrophenol in mice pro- duced chromatic-type breaks in bone marrow cells (Mitra and Manna, 19711; however, there was no linear relationship between the frequency of

Toxicity ot Selected Organic Contaminants in Drinking Water 243 chromosome aberrations and the dose of 2,4-dinitrophenol. 2,4-Di- nitrophenol also failed to induce DNA damage in an in-vitro alkaline elu- tion assay with Chinese hamsters V79 cells with or without P-450 or microsomal activation (Swenberg et al., 1976) and was negative in the Salmonella system (Chiu et al., 1978~. Streptomycin-independent mutants were produced in streptomycin- requiring E. cold after preincubation in the presence of picric acid (Demerec et al., 1951~. Yoshikawa et al. (1976) reported that mutation was induced in the Salmonella system when picric acid was tested in the presence of microsomes, but no activity was observed in their absence (Chiu et al., 19781. In contrast, Auerbach and Rob son (1949) failed to show mutagenicity in Drosophila. Carcinogenicity There are no data on the carcinogencity of mononi- trophenols; however, 4-nitrophenol has been selected by the National Cancer Institute for testing under its Carcinogenesis Bioassay Program. Boutwell and Bosch (1959) have studied the ability of a number of phenolic compounds to promote tumor formation in mouse skin following a single initiating dose of dimethylbenz[<x~anthracene. Although phenol was shown to have promoting activity in this system, both 2- and 4-nitro- phenol and 2,4-dinitrophenol failed to promote tumor development in the mice (Boutwell and Bosch, 1959; Stenback and Garcia, 1975~. In other experiments in mice, Spencer et al. (1948) did not detect tumor formation during a 6-month oral administration of 2,4-din*rophenol. The committee found no data on carcinogenic effects of picric acid. Teratogenicity No information on the teratogenicity of the mononitro phenols is available. The effects of 2,4-nitrophenol on fertility and gestation of female and fetal rats were examined by Wulff et al. (1935~. They administered 2,4-nitrophenol (20 mg/kg/day) to female rats for 8 days before introduc- ing the males. 2,4-Dinitrophenol was then administered orally twice daily until the litters were weaned. There was no effect on the weight gain of the mothers during pregnancy, the average number of offspring in each litter, nor were neonatal malformations detected. However, 25% of the offspring of treated females were stillborn, compared to only 6.8~o of those born to the controls. Moreover, the mortality during the nursing period among the viable offspring of the treated mothers was 30.9~o, compared to 13.4% for the young of the control mothers. Intraperitoneal (7.7 or 13.6 mg/kg) or oral (25.5 or 38.3 mg/kg) admin- istration of 2,4-dinitrophenol to pregnant mice during early organogenesis did not produce morphologic defects in the young, but embryotoxicity oc

244 DRINKING WATER AND HEALTH curred at the higher dosage levels (Gibson, 1973~. The higher concentra- tions also produced overt signs of toxicity, including hyperexcitability and hyperthermia in the dams, but they were not lethal. Bowman (1967) studied the effect of 2,4-dinitrophenol on the develop- ing chick embryo in vitro. Concentrations of 8 mg/liter or 370 mg/liter resulted in degeneration and sometimes complete absence of neural tissue accompanied by a reduction in the number of somites. Malformations such as hemiophthalmus and cross beak were induced in chick embryos following administration of 0.5 ,ug of 2,4-dinitrophenol per egg into the yolk sac at the 48th hour of incubation (Miyamoto et al., 1975~. Following examination of purified myelin in the malformed embryos, these in- vestigators suggested that exposure to dinitrophenol resulted in impaired embryonic myelination. No information on possible teratogenic effects of 2,4,6-trinitrophenoJ was found. CONCLUSIONS AND RECOMMENDATIONS Despite their high rates of production and use in a variety of products, data on the subacute and chronic toxicity of the nitrophenols are inade- quate. Suggested No-Adverse-Response Level (SNARE) More on itrop h en ols 24-Hour Exposure There are no adequate data from which to calcu- late this SNARL. 7-Day Exposure Ogino and Yasukura (1957) observed the develop- ment of cataracts in guinea pigs given daily intraperitoneal doses of 4-nitrophenol at 8.3 to 12.5 mg/kg for 6 to 11 days. Despite the absence of control animals and other shortcomings of these data, they will still be used for the SNARL calculation. Using the 8.3 mg/kg/day dosage, as- suming all the mononitrophenol came from the water during this period, a 70-kg human, and an uncertainty factor of 1,000, one may calculate the 7-day SNARL as: 8.3 mg/kg X 70 kg 1,000 X 2 liters = 0.29 mg/liter.

Toxicity of Selected Organic Contaminants in Drinking Water 245 Chronic Exposure This cannot be calculated due to the lack of ade- quate subacute or chronic data for any of the mononitrophenols. Dinitrophenols 24-Hour Exposure 7-Day Exposure No data are available for calculation. No data are available~for calculation. Chronic Exposure No good chronic toxicity data for 2,4-dinitrophenol are available. However, one can estimate an approximate chronic ex- posure limit from information on toxicity in humans and animals (Table VII-8. Several investigators (Homer, 1942; Homer et al., 1935; Tainter, 1933) have summarized the side-effects observed during the therapeutic use of 2,4-dinitrophenol for the treatment of obesity. At daily dosages as low as 2 mg/kg/day, skin lesions, hematological effects, neuritis, and cataracts developed in some patients over treatment periods as long as 4 months. Using this value and an uncertainty factor of 100 (since this is not a "no-toxic-effect" dose), and assuming a 20~o intake from water and a daily consumption of 2 liters by a 70-kg adult, one may calculate the chronic SNARL as: 2 mg/kg X 70 kg X 0.2 100 X 2 liters = 0.14 mg/liter. Spencer (1948) fed 2,4-dinitrophenol to rats for 6 months and observed no toxic effect at a dose of 10 mg/kg/day. Using this value and an uncer- tainty factor of 1,000, and assuming a 20~o intake from water and a daily consumption of 2 liters by a 70-kg adult, one may calculate the chronic SNARL as: 10 mg/kg X 70 kg X 0 2 = 0.07 mg/liter 1,000 X 2 liters which is in good agreement with the chronic SNARL calculated from ex- posure data on humans. Tr~nitrophenol 24-Hour Exposure Only acute toxicity data on 2,4,6-trinitrophenol (picric acid) are available. Severe poisoning in humans results from an

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Toxicity of Selected Organic Contaminants in Drinking Water 247 oral dose as low as 14 mg/kg (Gosselin et al., 1976~. Using this value and an uncertainty factor of 100 (since this is not a "no-toxic-effect" dose) and assuming 100% of intake during this period from a daily water consump- tion of 2 liters by a 70-kg adult, one may estimate the 24-hour SNARL as: 14 mg/kg X 70 kg 100 X 2 liters = 4.9 mg/liter. 7-Day Exposure Harris et al. (1946) reported that U.S. Navy person- nel developed hematuria upon drinking water that contained picric acid in concentrations ranging from 2 to 20 mg/liter. Assuming that this ex- posure occurred over several days and that 2 mg/liter represents the minimum concentration in water, one may estimate a 7-day SNARL, us- ing an uncertainty factor of 10, as: 2 mg/liter = 0 2 mg/liter. Chronic Exposure This cannot be calculated due to the lack of ade- quate chronic exposure data. Petroleum Products (Crude and Refined) The increasing demand for and consumption of petroleum and its by- products have greatly increased the risk of contamination of drinking water supplies. In 1974, world and U. S. crude oil production was estimated at 20,538 and 3,203 million barrels, respectively (U.S. Bureau of Mines, 1976~. In the same year, the estimated U.S. demand for refined petroleum products was: residual fuel oil, 963 million barrels; gasoline, 2,402 million barrels; distillate fuel oil 1,016 million barrels; kerosene and jet fuel, 427 million barrels; and lubricants, including grease, 56.7 million barrels. Crude oils vary widely in both physical and chemical properties. They are generally considered to be a complex mixture of hydrocarbons ranging in molecular weight from that of methane (e.g., 16.04) to possibly 100,000 or more (Kallio, 1976~. In petroleum, there are fewer compounds contain- ing nitrogen (amines), sulfur (sulfides), and oxygen (phenols). Kallio (1976) estimated that there may be as many as 1 million discrete com- pounds in crude oil. The hydrocarbon portion of petroleum is composed of three major classes of hydrocarbons-alkalies, alicyclics, and aromatics. Alkanes or paraffins average approximately 20~o of the oil fraction of crude oils but

248 DRINKING WATER AND HEALTH range in amounts from practically zero to close to lOO~o in different oils. About one-half of the alkane fraction is comprised of normal or straight- chain hydrocarbons, and the other half contains branched-chain hydro- carbons, both ranging from Cal to approximately C40. The content of ali- cyclic hydrocarbons (naphthenes, cycloalkanes, or cycloparaffins) also varies among crudes, but it is generally considered to be about 50~o of the oil fraction, consisting mostly of cyclopentane or cyclohexane types and smaller amounts of seven- and eight-membered ring hydrocarbons. Aro- matics generally do not account for more than 20% of the oil fraction. Benzene predominates in the aromatic fraction, but polynuclear aromatic compounds of up to nine rings have been identified. Refined petroleum solvents that may contaminate drinking water sup- plies may be classified according to the following major divisions: petroleum ether, rubber solvent, varnishmakers' or painters' naphtha, mineral spirits, Stoddard solvents, and kerosene (National Institute for Occupational Safety and Health, 1977a). Petroleum ether is a refined petroleum solvent with a boiling point ranging from 30°C to 60°C and is typically composed of 80~o pentane and 20% isohexane. Rubber solvent is a refined petroleum solvent with a boiling point ranging from 45°C to 125°C and is composed of hydrocarbons whose carbon chain lengths range from Cs to C7. Varnishmakers' or painters' naphtha is a refined petroleum solvent with a boiling point ranging from 95°C to 160°C and contains hydrocarbons of chain lengths from Cs to Cal. Mineral spirits comprise a fraction with a boiling point ranging from 150°C to 200°C, while that of Stoddard solvents boils in the range of 160°C to 210°C and contains mainly C7 to C~2 hydrocarbons. Kerosene is a refined petroleum solvent with a boiling point ranging from 175°C to 325°C. Gasoline, the major refined petroleum solvent, normally contains more than 200 different hydrocarbons, mainly in the Cs to Cg fraction. Its boil- ing point ranges from 26°C to 204°C (PEDCo. Environmental, Inc., 19771. Alkanes and aromatic compounds generally constitute the largest fraction of gasoline, but olefins and alicyclic compounds are also present. Analysis showed the mean compos*ion of 15 premium grade gasolines to be: alkanes, 50370; aromatic compounds, 27~o; olefins, ll~o; and alicyclic compounds, 11% (PEDCo. Environmental, Inc., 1977~. In addition, gasoline typically contains a variety of additives for improving engine per- forrnance, of which lead alkyls are predominant. The composition of gasoline varies as a function of the crude oil, the refinery process, the gasoline blending makeup for different grades, the grade of the gasoline, the climate of the marketing region, and the brand. Thus, gasoline is a blend consisting of a mixture of various blending stocks, notably catalytically cracked gasoline, light, straight-run gasoline,

Toxicity of Selected Organic Contaminants in Drinking Water 249 hydrocracked gasoline, and thermally cracked gasoline, reformats, and alkylate. Since the highest octane blending components in gasoline are aromatics and branched-chain alkanes, refineries attempt to produce a blending stream rich in these components. Lower octane blending stocks can be used in the production of leaded gasoline because alkyl lead ad- ditives increase gasoline octane ratings. However, the increasing demands for unleaded gasoline are forcing refineries to produce more fractions rich in aromatics and branched-chain alkanes since the octane of unleaded gasoline depends solely on the natural octane rating of its blending com- ponents. Approximately 325 inorganic and organic fuel additives were registered in 1972 (National Academy of Sciences, 1976a). These compounds, which can be classified into about 15 different chemical types, are added for such purposes as antiknock components, antioxidants, surfactants' and deposit modifiers. They are usually added to fuel in very small amounts ranging from a few micrograms per milliliter to a few hundred micro- grams per liter. The average tetraethyl lead content, however, is one part per 1,300 parts of gasoline. EPA analyses of trace elements in SO gasolines showed that only lead and sulfur were present in concentrations greater than a few micrograms per milliliter (Jungers et al., 1975~. Concentrations of lead in these gasoline samples ranged from 132 to 763 ,ug/ml, whereas the sulfur concentration fell between 4 and 720 ,ug/ml. Drinking water may be contaminated by crude oil or its products, either accidentally or operationally, wherever oil is produced, transported, stored, processed, or used. Although there is a greater probability of freshwater contamination by refined petroleum products than by crude oil itself, the pollutants encountered in freshwater inevitably include the complete range of chemicals present in oils and their products as well as various chemical additives. In nearly all cases, the petroleum-derived pollutants have undergone some degree of weathering. Crude oils and refined petroleum products exposed on surface waters are subjected simultaneously to several physical and chemical processes that can diminish their volume as well as alter their composition. In general, components with lower boiling points evaporate. The remaining pollutants may emulsify with water while undergoing dissolution, oxida- tion, and biological degradation. The hydrocarbons in petroleum are af- fected at different rates, depending on their molecular weights and struc- tures, and the rates at which these processes occur, in turn, are dependent to a considerable extent on environmental conditions including temper- ature and wind velocities. When studying the toxic hazards resulting from contamination of drinking water by crude and refined petroleum, one must consider the

250 DRINKING WATER AND HEALTH relative solubilities of the various constituents. With the exception of C2-C4 compounds, the aliphatic, olefinic, and alicyclic hydrocarbons are relatively insoluble in water (Anderson et al., 19741. By contrast, the solubility of the aromatic constituents of petroleum, especially benzene and substituted benzenes, is quite high. Although hydrocarbons are not highly soluble in water, this property should be considered because of the volume of water that is available to the compounds. Solubilities are highest for the low-molecular-weight aromatic compounds, then decrease through the series: aromatics, alicyclic compounds, branched-chain alkalies, and n-alkanes. Within each series, solubility decreases with in- creasing molecular weight. An example of these differences is shown in Table VII-9, in which the hydrocarbon contents of water that has been equilibrated with two different crude oils and No. 2 fuel oil are compared. The solubility of hexane, a typical alkane, in the aqueous fraction is substantially lower than that of either benzene or toluene, two of the prominent components of the aromatic fraction. These results are in line with the relative water solubilities of the three pure hydrocarbons. In studies of experimental oil spills, benzene is solubilized and removed from the surface oil slick most rapidly of all of the major petroleum components (McAuliffe, 1977~. The chemical processes affecting petroleum constituents are extremely complex and not well understood. They are most likely photochemical ox- idations and polymerizations. The products of oxidation are generally more soluble than the parent hydrocarbons and, hence, are more easily removed from the aquatic environment by further oxidation and by microbial degradation. Microbial metabolism initially affects only the n-alkanes. Resistance to TABLE VII-9 Specific Hydrocarbon Content (,u/ml) of Water-Soluble Fractions Equilibrated with Crude and Refined Oils Crude Oil Crude Oil No. 2 Fuel Water Solubility No. 1 in No. 2 in Oil in in Freshwater Compound Seawatera Seawatera Seawatera at 25°Cb Hexane 0.09 0.29 0.014 9.5 Total saturates 9.86 11.62 0.54 Benzene 6.75 3.36 0.55 1,780 Toluene 4.13 3.62 1.04 515 Total aromatics 13.90 10.03 5.74 aData from Anderson et al., 1974. bData from McAuliffe, 1966.

Toxicity of Selected Organic Contaminants in Drinking Water 251 microbial attack increases through the series of n-alkanes, branched- chain cycloalkanes, and alicyclic and aromatic compounds. This biolog- ical process, unlike the physical processes, is not necessarily concentrated on the lower-molecular-weight components, but the rate of attack is quite slow. HEALTH ASPECTS An in-depth consideration of the hazards associated with contamination of drinking water by crude oil or refined petroleum products is a very com- plex and difficult undertaking that is beyond the scope of this volume. Not only are there literally tens of thousands of different hydrocarbons present in crude oil and various refined petroleum products, but crude oil also contains countless numbers of oxygen, nitrogen, and sulfur com- pounds. Moreover, hydrocarbons are rearranged by catalytic processes during refining to form new compounds, and hundreds of different chemicals are also added to petroleum solvents, lubricating oils, and other by-products to impart desirable qualities. These substances all have dif- ferent solubilities, volatilities, and, most importantly, toxicological prop- erties. Since most of the toxicity data on the constituents of petroleum and petroleum solvents involve inhalation exposure, there is relatively limited information concerning the oral toxicity of these compounds. Despite this deficiency, one can make approximations that greatly simplify the estima- tion of toxic hazards. Crude oil and refined petroleum products contain four major groups of hydrocarbon components: alkanes, olefins, and alicyclic and aromatic compounds (PEDCo. Environmental, Inc., 1977~. Various types of chemi- cal additives are also present in gasoline and other refined petroleum products. Small amounts of nitrogen, sulfur, and oxygen compounds are present in crude petroleum, but most of them are removed during the refining process. The alkane or paraffin fraction, containing primarily aliphatic hydrocarbons from C3 to Cal, has relatively low toxicity. Alkanes of five or more carbons have strong narcotic properties following inhalation ex- posure. Recent toxicological and epidemiolog~cal evidence suggests that acute intoxication by these alkalies involves a transient depression of the central nervous system. Chronic intoxication with alkalies may lead to development of a more persistent polyneuropathy (National Institute for Occupational Safety and Health, 1977b). Polyneuropathy has been ob- served in workers in shoe factories and various leather industries following inhalation of Cs-C7 aliphatic hydrocarbons, which are generally used as solvents for leather adhesives (Buiatti et al., 1978~. Although polyneurop

252 DRINKING WATER AND HEALTH athy has been attributed to exposure to n-hexane, recent evidence suggests that such neuropathies can be caused by other alkanes and their isomers (National Institute for Occupational Safety and Health, 1977b). Animal studies have shown that peripheral neuropathies can be produced with n-hexane (Schaumburg and Spencer, 1976~. Recent studies suggest that the hydrocarbon-induced polyneuropathies can be attributed primarily to the neurotoxic effect of 2,5-hexanedione, which is a metabolite of n-hex- ane (Perbellini et al., 1980; Spencer et al., 1978~. Presumably, similar ketone metabolites of other hydrocarbons also produce neuropathy. In general, straight-chain alkalies appear to be more toxic than the branched . . cnam Isomers. The olefin or alkene fraction contains unsaturated aliphatic hydrocar- bons. These compounds exhibit little toxicity other than weak anesthetic properties (PEDCo. Environmental, Inc., 1977~. The naphthenes or cycloparaffins are saturated and unsaturated ali- cyclic hydrocarbons that resemble aliphatic hydrocarbons in their toxicity since they act as general anesthetics and have depressant effects on the central nervous system with a relatively low degree of acute toxicity (PED- Co. Environmental, Inc., 1977~. These compounds are not cumulative and little if any significant toxicity has been noted upon prolonged ex- posure to naphthene vapors. Aromatic hydrocarbons have generally been regarded as the most toxic fraction of petroleum and petroleum solvents (PEDCo. Environmental, Inc., 1977~. They are also the most soluble in water. The aromatic fraction contains benzene, alkyl derivatives of benzene, and small quantities of various polynuclear aromatic hydrocarbons. Benzene, because of its volatility, unique myelotoxicity, and carcinogenic potential, is the most toxic component. The toxicity of toluene, the xylenes, and other alkylated benzene derivatives is considerably lower. Tetraethyl and tetramethyl lead and organic halogen compounds are added to gasolines as antiknock agents to improve performance. Although these compounds have a fairly high degree of toxicity, their concentrations in gasolines are quite low. Moreover, these additives have a relatively low solubility in water. Ethylene dibromide, one of the organic halogens added to gasoline, has been identified as an animal carcinogen and is also mutagenic (National Academy of Sciences, 1980a). Thus an assessment of the toxicity of drinking water contaminated by crude oil or refined petroleum products should focus on the aromatic frac- tion. Since benzene is the most acutely toxic member of the aromatic frac- tion and also has the highest solubility in water, the toxicity of drinking water polluted by crude or refined petroleum will be determined largely on the basis of benzene content. Since toluene is also found in high concen- trations in such contaminated water, it is also considered in this section.

Toxicity of Selected Organic Contaminants in Drinking Water 253 METABOLISM Benzene The metabolism of benzene, which has been reviewed by the U.S. En- vironmental Protection Agency (1979a), Rusch et al. (1977), Snyder and Kocsis (1975), and Snyder et al. (1977), is widely accepted as a prere- quisite to its toxicity. It is metabolized by cytochrome P-450 monooxy- genases to form the highly reactive arena-oxide-type metabolite, benzene oxide. The oxide can spontaneously rearrange to form phenol, undergo enzymatic hydration followed by dehydrogenation to catechol, react en- zymatically to form a glutathione conjugate, or bind covalently with cellular macromolecules. Sulfate and glucuronide conjugates are also formed. The specific metaboliteks) of benzene that induce leukemia or other toxicities have not yet been identified, but likely candidates include benzene oxide, catechol, and hydroquinone or the corresponding semi- quinones (U.S. Environmental Protection Agency, 1979a). HEALTH ASPECTS Observations in Humans Chronic Effects The toxicity of benzene has been reviewed in several reports (National Academy of Sciences, 1976b, 1977, 1980a; Snyder and Kocsis, 1975; U.S. Environmental Protection Agency, 1979a). The hema- tological toxicity following chronic exposure to benzene in humans is well established. Reported effects include myelocytic anemia, thrombo- cytopenia, or leukopenia and acute myelogenous and monocytic leuke- mia. The data thus suggest that benzene is a leukemogen in humans. Observations in Other Species The toxicity of benzene in laboratory animals has been summarized in the first and third volumes of Drinking Water and Health (National Academy of Sciences, 1977, 1980a) and by the U.S. Environmental Protection Agency (1979a). Acute Effects No new information has become available since those reports were published. Chronic Effects In most of the studies on the chronic toxicity of benzene to animals, investigators used inhalation exposures. An exception was the study of Wolfe et al. (1956), who administered benzene orally to

254 DRINKING WATER AND HEALTH rats 5 days/week for 6 months at daily doses of 1, 10, 50, and 100 mg/kg body weight. Leukopenia and erythrocytopenia were observed at the 50 mg/kg dosage, while slight leukopenia was seen at 10 mg/kg/day. Mutagenicity There is no information in addition to that previously published (National Academy of Sciences, 1977, 1980a; U.S. En- vironmental Protection Agency, 1979a). Carcinogenicity No new information has become available. Toluene Toluene, as well as several other alkyl benzenes, were recently reviewed in depth (National Academy of Sciences, 1980b). Because of that review, another review by the U.S. Environmental Protection Agency (1979b) and the calculations for acute and chronic SNARL's in Drinking Water and Health, Vol. 3 (National Academy of Sciences, 1980a), no further con- sideration is necessary at this time. CONCLUSIONS AND RECOMMENDATIONS One may estimate the toxicity of drinking water contaminated by a crude oil or a refined petroleum solvent on the basis of its benzene and toluene contents. However, such estimates must be regarded only as approxima- tions. More precise estimates of risk can be made only after there have been much more detailed and thorough evaluations of acute and chronic toxicities of all of the thousands of chemicals present in crude and refined petroleum products, singly and in combination. There are obviously still many unknown factors associated with the toxicity of petroleum products. This is illustrated by the recent, puzzling finding of a high frequency of brain tumors among workers in the petrochemical industry (Fox, 1980~. Other substantial limitations associated with the calculation of 24-hour and 7-day SNARL's for both benzene and toluene include the inadequacy of the oral toxicity data and the lack of data on synergistic interactions among petroleum constituents. These short-term SNARL calculations also ignore the carcinogenicity of benzene in humans. No chronic SNARL has been calculated since benzene is probably a leukemogen in humans. Suggested No-Adverse-Response Level {SNARLJ 24-Hour Exposure There are no adequate data from which to calcu- late a 24-hour SNARL.

Toxicity of Selected Organic Contaminants in Drinking Water 255 7-Day Exposure Oral administration of benzene to rats 5 days/week for 6 months produced a minimal toxic effect at a dose of 10 mg/kg body weight. Adjusting to a 7-day exposure, 10 mg/kg/day X 5/7 days = 7.1 mg/kg/day. Applying a safety factor of 1,000, assuming that 2 liters/day of drinking water is the only source of benzene during this period for a 70-kg human, one may calculate the 7-day SNARL for humans as: 7. ~ mg/kg X 70 kg 0 25 /lit This committee has reevaluated the data on benzene to arrive at this 7-day SNARL. It elected to use the threshold dose rather than the leukopenic dose. This, coupled with a more conservative safety factor, results in a 7-day SNARL of 0.25 mg/liter as compared to the value of 12.6 mg/liter in Drinking Water and Health, Vol. 3 (National Academy of Sciences, 1980a). Chronic Exposure No chronic SNARL can be calculated because benzene may be a leukemogen in humans. Polynuclear Aromatic Hydrocarbons These compounds were reviewed previously by the Criteria and Standards Division, Office of Water Planning and Standards, U.S. Environmental Protection Agency (1979e). The committee has reviewed and discussed that document for accuracy and completeness. The following pages sum- marize those findings and are augmented, when necessary, by additional information. Polynuclear aromatic hydrocarbons (PAH's) are a diverse class of com- pounds consisting of substituted and unsubstituted polycyclic and heterocyclic aromatic rings. PAM's are formed as a result of incomplete combustion of organic compounds in the presence of insufficient oxygen. This leads to the production of C-H free radicals, which can polymerize to produce various PAH's. Among the PAM's, benzo~aipyrene (BaP) is the most thoroughly studied because of its ubiquitous presence in the environ- ment and its high carcinogenic activities in laboratory animals. The toxic- ity of BaP was reviewed in the first volume of Drinking Water and Health (National Academy of Sciences, 1977~. Although small amounts of PAM's originate from natural or endog- enous sources, most of the PAM's in surface waters are derived from human activity. Discharges of raw and industrial wastewater, atmospheric fallout and precipitation, road runoff, and leachate from polluted soils all

256 DRINKING WATER AND HEALTH contain substantial concentrations of PAM's (Andelman and Suess, 1970), thereby contributing to the contamination of surface waters by these com- pounds. The PAM's are extremely insoluble in water. For example, in relatively clean water, the solubility of BaP is only approximately 10 ng/liter (Andelman and Snodgrass, 1974~. Despite the relative insolubility of the PAM's, their concentrations in surface waters can be increased by the ac- tion of detergents and other surfactants. Total PAH concentrations in sur- face waters have been found to range from 0.14 to 2.5 ,ug/liter, whereas BaP concentrations range from 0.0006 to 0.35 ,ug/liter (Borneff and Kunte, 1964, 1965; Harrison et al., 1975~. Similar concentrations exist in surface waters used for drinking water supplies; however, water treatment in the United States significantly lowers these concentrations to a range of 0.003 to 0.14 ,ug/liter (Basu and Saxena, 1977, 1978; Borneff and Kunte, 1964; Harrison et al., 1975~. In surface waters, one-third of the total PAM's is bound to large suspended particles, one-third is bound to finally dispersed particles, and the remainder is present in dissolved form (U.S. Environmental Protection Agency, 1980~. The usual sedimentation, floc- culation, and filtration process removes a good share of the PAM's present in the water. In addition, from 50~o to 60% of PAM's such as BaP are removed by chlorination of the water (U.S. Environmental Protection Agency, 1979e). PAM's are generally quite stable in water, remaining in solution over long periods. Iluitsky et al. (1971) showed that after 35 to 40 days, Ho to 20% of an initial BaP concentration of 10 ,ug/liter remained in water. As described in the earlier chapters on distribution systems, finished waters from various treatment sites are transported to the consumers through a variety of pipelines. PAM's leach from the tar or asphalt linings of these pipes (Basu and Saxena, 1977, 1978; Sorrell et al., 1980; U.S. En- vironmental Protection Agency, 1979e), resulting in increased concentra- tions of these compounds in water reaching the consumers. On the other hand, cement-lined pipes produce lower PAH concentrations, possibly because PAM's are adsorbed from the water. Basu and Saxena (1977, 1978) analyzed BaP and five other PAM's in finished drinking water from 15 U. S. cities. BaP levels ranged from < 0.1 to 2.1 ng/liter. The total concentration of carcinogenic PAM's ranged from 0.2 to 11.3 ng/liter, whereas levels of total PAM's ranged from 0.3 to 138 ng/liter. Sorrell et al. (1980) surveyed other data on the concentra- tions of PAM's in finished and distributed drinking waters. Phenanthrene was found in the highest concentrations in drinking water (3-3,300 ng/ liter), whereas concentrations of most other PAM's were less than 1 ng/ liter.

Toxicity of Selected Organic Contaminants in Drinking Water 257 Levels of PAM's detected in U.S. drinking waters are well below the limit of 200 ng/liter recommended by the World Health Organization (1970~. Furthermore, Shabad and Iltnitskii (1970) stated that the amount of carcinogenic PAM's in water ingested by humans is typically only 0.1 To of the amount consumed in foods. Thus, if the total PAH uptake from food is 4.15 mg/year (U.S. Environmental Protection Agency, 1979e), the uptake of PAM's in drinking water by humans would probably not exceed 4 ~g/year. METAB OLI SM Studies in animals indicate that structurally related PAM's, such as BaP, chrysene, 7,12-dimethylbenz~aJanthracene (DMBA), benz~a janthracene, and 3-methylcholanthrene (3MC), are readily absorbed from the in- testinal tract and tend to localize primarily in body fat and fatty tissues such as the breast (Bock and Dao, 1961; Kotin et al., 1959; Schlede et al., 1970a,b). Disappearance of BaP from the blood and liver of rats following a single intravenous injection is very rapid (Schlede et al., 1970a), having a half- life in blood of less than 5 minutes and a half-life in liver of 10 minutes. In both blood and liver, however, the initial rapid elimination phase is followed by a slower disappearance phase, lasting 6 hours or more. Schlede and coworkers (1970a) concluded that a rapid equilibrium is established between BaP in blood and that in liver and that the com- pound's fast disappearance from the blood is due to both metabolism and distribution in tissues. The distribution of radioactivity in rats after administration of labeled dibenz~aJanthracene, DMBA, and 3MC by stomach tube was comprehen- sively studied by Daniel et al. (19671. The major route of excretion was found to occur via the bile into the feces. There was a rather prolonged retention of radioactivity in body fat, ovaries, and adrenals. Early physicochemical calculations to explain the carcinogenicity of various PAM's were based on the chemical reactivity of certain regions of the molecule (Pullman and Pullman, 19551. This concept, however, did not appear to hold true for many of the PAH's. More recently, it was learned that PAM's are metabolized by enzyme-mediated oxidative mech- anisms to form reactive electrophiles (Lehr et al., 1978~. These reactive metabol*es can then covalently interact with cellular constituents such as RNA, DNA, and proteins, ultimately leading to tumor formation (Miller, 1978). The necessity for metabolic activation to express the carcinogenesis of PAH has prompted the investigation of PAH metabolism in numerous

258 DRINKING WATER AND HEALTH animal models and human tissues. From these studies has emerged a gen- eral understanding of the mechanisms involved in the biotransformation of these compounds. It is now known that PAM's are metabolized initially by the cytochrome-P-450-dependent monooxygenases, which are localized in the endoplasmic reticulum. These enzymes are often designated as aryl hydrocarbon hydroxylases (Conney, 1967~. Their activity is readily induced by exposure to various chemicals. Although they are found in most mam- malian tissues, they are located predominantly in the liver. In the initial step of the activation process, the hepatic cytochrome-P-450-dependent monooxygenases oxidize PAM's to reactive epoxide metabolites (Lehr et al., 1978; Levin et al., 1977a; Selkirk et al., 1971, 1975; Sims and Grover, 1974; Thakker et al., 1977~. The PAH epoxides can undergo various fur- ther reactions, including hydration by the enzyme epoxide hydrolase, which is also located in the endoplasmic reticulum, to form trans- dihydrodiols. These in turn are oxidized by the cytochrome-P-450- dependent monooxygenases to form the highly reactive dial epoxides, the ultimate carcinogens. A schematic representation of the principal metabolic pathways involved in the activation of BaP is shown below. Benzo~aipyrene r Cytochrome P-450 enzymes Benzo~a]pyrene-7,8-oxide l Epoxide hydrolase Benzo~aipyrene-7,8-dihydrodiol l Cytochrome P-450 enzymes Benzotaipyrene-7,8-diol-9,10-oxide Carcinogenesis The metabolic profile of BaP, the most representative and well-studied compound of the PAM's, has been fairly well established. Known metabo- lites of BaP include five phenols, 1-, 3-, 6-, 7-, and 9-hydroxy BaP; three dihydrodiols, the ~-~-enantiomers of BaP-4,5-, 7,8-, and 9,10-trans- dihydrodiols; and three quinones, BaP-1,6-quinone, BaP-3,6-quinone, and BaP-6,12-quinone (Holder et al., 1974; Selkirk et al., 197-4; Sims, 1970; Thakker et al., 1976; Yang et al., 1978a,b). Considerable new evidence implicates the diol epoxide ~ + ~-7d,8<x-dihydroxy-9cx,10cz-epoxy- 7,8,9,10-tetrahydro-BaP as the ultimate carcinogen derived from BaP (Huberman et al., 1976; Jerina et al., 1976; Kapitulnik et al., 1978a,b; Levin et al., 1976a,b, 1977b; Thakker et al., 1977, 1979; Yang et al., 1978a,b). Epoxides are the initial cytochrome-P-450-catalyzed oxidation

Toxicity of Selected Organic Contaminants in Drinking Water 259 products of BaP (Jerina and Daly, 1974; Sims and Grover, 1974~. They rearrange nonenzymatically to form phenols, are hydrated by epoxide hydrolase to trans-dihydrodiols, or form glutathione conjugates catalyzed by glutathione S-transferase (Bend et al., 1976; Sims and Grover, 1974~. Some phenols are converted to glucuronide conjugates, a reaction that is catalyzed by UDP-glucuronyl transferase. Others form sulfate conjugates by the action of sulfotransferases (Cohen et al., 1976~. Certain epoxides on saturated, angular benzo-rings that form part of a "bay-region" are important in the activation of various PAM's (Jerina et al., 1978; Lehr and Jerina, 1977; Lehr et al., 1978; Wood et al., 1979~. The bay region is typified by the hindered region in the 10 and 11 positions of BaP. Diol epoxides on saturated benzo-rings that form part of a bay region of a hydrocarbon are very active chemically and are more readily converted to carbonium ions than are epoxides not located in the bay region. Therefore, they are more potent alkylating agents with greater mutagenic and carcinogenic activities. Molecular orbital calculations can be used to predict the relative carcinogenic potential of a series of PAM's from their relative tendencies to form carbonium ion from their bay region dial epoxides. HEALTH ASPECTS Observations in Humans Although exposure to PAM's occurs predominantly by direct ingestion of the compounds in food and in drinking water, there are no studies to document the possible carcinogenic risk to humans by these routes of ex- posure. It is known only that significant quantities of PAM's can be in- gested by humans and that ingestion by animals of similar amounts may result in cancers at various sites in the body. Convincing evidence from air pollution studies indicates an excess of mortality from lung cancer among workers exposed to large amounts of PAH-containing materials such coal gas, tars, soot, and coke-oven emis- sions (Doll et al., 1965, 1972; Hammond et al., 1976; Henry et al., 1931; Kawai et al., 1967; Kennaway, 1925; Kennaway and Kennaway, 1936; Kuroda, 1937; Mazumdar et al., 1975; Redmond et al., 1972, 1976; Reid and Buck, 1956~. However, there is no definite proof that the PAM's pre- sent on these materials are responsible for the observed lesions. Never- theless, our understanding of the characteristics of PAH-induced tumors in animals and their close resemblance to carcinomas of the same target organs in humans strongly suggests that PAM's pose a carcinogenic threat to humans, regardless of the route of exposure. There is now also good

260 DRINKING WATER AND HEALTH evidence of the overwhelming importance of cigarette smoking in the etiology of lung cancer in humans, and cigarette smoke is known to con- tain PAH's. Observations in Other Species Acute Effects There is little information on the acute toxicity of PAH's. It has been reported that the acute oral LD50 for DMBA in mice is 350 mg/kg (National Institute for Occupational Safety and Health, 1978d). Most adult male mice in a dominant lethal study survived intraper~toneal doses of BaP administered in tricap~lin in doses of 500, 750, and 1,000 mg/kg (Epstein et al., 1972~. Robinson et al. (1975) found that mouse strains that were "responsive" to the induction of cytochrome-P-450 and related monooxygenases by PAM's had significantly shorter survival times following 500 mg/kg intraperitoneal doses of BaP, 3MC, or DMBA than did "nonresponsive" mouse strains. At lower, single intraperitoneal doses of BaP at 100 mg/kg BaP and 3MC at 300 mg/kg, however, the dif- ferences in survival times were not evident. Subchronic and Chronic Effects Almost all mice strains that were "responsive" to induction to cytochrome P-450 by PAM's were still alive at 180 days following daily oral intake of BaP, 3MC, or DMBA at doses of 120 mg/kg/day, whereas most of the "nonresponsive" strains died within 20 days after the start of the experiment (Robinson et al., 1975~. The mechanism for this effect is not known, but the PAH-exposed nonrespon- sive mice showed a rapid loss in body weight, increased binding of BaP metabolites to DNA, and leukopenia accompanied by a marked decrease in myeloid precursors in bone marrow. PAM's at doses of 150 mg/kg or more produce systemic toxicity, which is manifested by the inhib*ion of body growth in rats and mice (Haddow et al., 1937~. Tissue damage resulting from the administration of various PAM's to laboratory animals is often widespread and severe, although there may be select*e organ destruction such as adrenal necrosis and Iym- phoid tissue damage. Current opinion favors a concept that normally pro- liferating tissues, e.g., intestinal epithelium, bone marrow, Iymphoid organs, and testes, are preferential targets for PAM's and that this suscep- tibility may be due to the specific attack on DNA by cells in the S-phase of mitosis (Philips et al., 1973~. Target organs for the toxic action of PAM's are diverse because of the extensive distribution of these compounds within the body and because of their selective attack on proliferating cells. Damage to the hematopoietic and Iymphoid systems of laboratory animals resulting from exposure to

Toxicity of Selected Organic Contaminants in Drinking Water 261 PAM's is a particularly interesting observation. Yasuhira (1964) described severe degeneration of the thymus and markedly reduced weight of the spleen and mesenteric lymph nodes of CF1, Swiss, and C57BL mice given a single intraperitoneal injection of 3MC (0.3-1.0 mg/animal, i.e., ap- proximately 150-500 mg/kg body weight) between 12 hours and 9 days after birth. Degeneration of young cells in the bone marrow and retarda- tion of thyroid gland development were also noted. Newborn mice were highly susceptible to the toxic effects of the PAM's; many of them died from wasting disease following treatment. After 50-day-old female Sprague-Dawley rats were fed one dose of DMBA at either 112 or 133 mg/kg body weight, pancytopenia accom- panied by severe depression of hematopoietic and Iymphoid precursors developed within weeks (Cawein and Sydnor, 1968~. Female Sprague- Dawley rats receiving oral doses of DMBA at 300 mg/kg and male rats receiving an intravenous injection of DMBA at 50 mg/kg incurred injury to the intestinal epithelium, extreme atrophy of portions of the hematopoietic system, shrinkage of the Iymphoid organs, agranulocytosis, Iymphopenia, and progressive anemia (Philips et al., 1973~. Mortality among female rats was approximately 65~o. Sixty-day-old adult rats that were given 20 mg of DMBA orally and 5 mg intravenously developed tran- sient degenerative changes in the testes, which were most evident between 38 and 40 days after treatment. Lesions of the testes were highly specific, involving destruction of spermatogonia and resting spermatocytes the only testicular cells that actively synthesize DNA. Neither the remaining germinal cells nor the interstitial cells were damaged. Surprisingly, no testicular damage was produced by single oral doses of BaP at 100 mg and 3MC at 105 ma. Numerous investigators have demonstrated that carcinogenic PAM's can produce an immunosuppressive effect. This was first observed by Malmgren et al. (1952), who administered high doses of 3MC and dibenz~a,hianthracene to mice. Subsequent studies established that single carcinogenic doses of 3MC, DMBA, and BaP given to sheep caused pro- longed depression in the immune response of red blood cells (Stjernsward, 1966, 1969~. Noncarcinogenic hydrocarbons, such as benzo~e~pyrene and anthracene, produced no immunosuppressive act**y. There is substan- tial evidence indicating that the degree of immunosuppression is cor- related with the carcinogenic potency of the PAM's (Balwin, 1973~. Mutagenicit;y The results obtained with several in-vitro mutagenesis test systems, especially the Ames Salmonella typhimu~ium assay, support the belief that most carcinogenic chemicals are mutagenic as well. For the PAM's, the Ames assay has been very effective in detecting the parent

262 DRINKING WATER AND HEALTH structures and biotransformation products that possess carcinogenic ac- tivity (Brookes, 1977; McCann and Ames, 1976; McCann et al., 1975; Teranishi et al., 1975; Wislocki et al., 1976b; Wood et al., 1976a). The use of S. typhimur~um strains to detect chemically induced mutagens and microsomal preparations to provide metabolic activation has also made possible investigations of the mechanism of PAH-induced mutagenesis. In particular, an exhaustive survey of the mutagenicity of all the possible ox- idative metabolites of BaP has helped to confirm the belief that dial epox- ide intermediates are the ultimate mutagens/carcinogens derived from PAM's (Jerina et al., 1976; Levin et al., 1977a,b; Thakker et al., 1976; Wislocki et al., 1976a,b; Wood et al., 1976a,b). The mutagenic activity of the PAM's and their derivatives has also been examined in various mammalian cell line cultures. These studies have been conducted primarily in Chinese hamster cell lines, either V79 cells derived from male lung tissue or CHO cells derived from the ovary. Using mammalian cells, Huberman and Sachs (1974, 1976) have demonstrated the mutagenicity of a number of carcinogenic PAH's. In later studies, Huberman et al. (1977) showed a correlation between the degree of car- cinogenicity and the frequency of induced somatic mutations. The Chinese hamster V79 cells have also served to clarify the metabo- lism of BaP and the formation of the highly mutagenic and carcinogenic BaP-7,8-diol-9,10-epoxide metabolites as the ultimate mutagens/carcino- gens (Brookes, 1977; Hube~man et al., 1976, 1977; Jerina et al., 1976; Levin et al., 1976a; Wood et al., 1976a,b). Carcinogenicity Many PAM's produce tumors in the skin and most epithelial tissues of practically all species of animals tested. Malignancies are often induced by acute exposure to microgram quantities of these PAH's. Latency periods can be short (4-8 weeks), and the tumors produced may resemble carcinomas in humans. Historically, studies of the carcin- ogenicity of PAM's have focused primarily on effects on the skin or lungs. Moreover, the compounds are frequently injected subcutaneously or in- tramuscularly to produce sarcomas at the injection site. Ingestion has not been a preferred route of administration for the bioassay of PAH's. The tumorigenic effects of PAM's when applied to the skin of animals have been known for decades. Iball (1939) compared the carcinogenicity of a series of PAM's and observed that tumorigenic potency of the com- pounds in mouse skin to be: DMBA > 3MC > BaP > cholanthrene, etc. An additional compilation of tumorigenicity of the PAM's is provided in a monograph by the International Agency for Research on Cancer (1973~. The carcinogenicity of PAM's resulting from oral intake has not been studied as thoroughly as it has been for other routes of administration. Never

Toxicity of Selected Organic Contaminants in Drinking Water 263 theless, tumors at various sites do result when BaP is administered orally to rodents (International Agency for Research on Cancer, 1973~. Tumors appeared in mice after a 0.2 mg single oral dose of BaP and after concen- trations ranging from 50 to 250 ,ug/g had been fed to them in the diet for approximately 100 days. Stomach tumors, leukemias, and lung adenomas were produced in these animals. Other studies showed the carcinogenicity of BaP in the rat and hamster following oral administration. However, BaP produced carcinomas less effectively than other PAM's, notably DMBA, 3MC, and dibenz~a,h Janthracene. An examination of comparative carcinogenicity within the same tumor model system can provide valuable insight concerning relative risks of various PAH's. Shimkin and Stoner (1975) attempted this by injecting mice intravenously with one, approximately 0.25-mg dose of aqueous dispersions of PAH's. In this test system, 3MC displayed the greatest lung-tumor-forming capability, followed closely by dibenz~a, h janthra- cene. BaP was considerably less potent. Teratogenicity BaP had little effect on developing embryos in several mammalian and nonmammalian species, although there appeared to be resorption of embryos in the rat (Rigdon and Rennels, 1964~. Mice fed 1 ,ug/g BaP in the diet over their entire life-span reproduced normally, no malformations were observed in their offspring, and no resorption of em- bryos was evident (Ridgon and Neal, 19651. BaP, 3MC, and DMBA ad- ministered in single intraperitoneal doses of 80 mgikg in corn oil destroyed primordial oocytes in mouse ovaries. Thus, these PAM's are capable of producing premature ovarian failure in rodents. DMBA and its hydroxymethyl derivatives are teratogenic in the rat (Bird et al., 1970; Currie et al., 19701. CONCLUSIONS AND RECOMMENDATIONS The attempt to develop a drinking water criteria for PAM's as a class is hindered by several deficiencies in the scientific data base: · The PAM's are comprised of many compounds capable of inducing diverse biological effects and having different carcinogenic potential. A "representative" PAH mixture has not been defined. · The extrapolation to humans of results obtained from studies with BaP (or any other pure PAH) in animals may not be valid given the diverse nature of PAH mixtures. · Few acute, subchronic, or chronic toxicity studies in animals have been conducted with oral exposure to defined PAH mixtures.

264 DRINKING WATER AND HEALTH · No definitive acute toxicity data are available for pure PAH's. · There are no data concerning the effects on humans resulting from exposure to defined PAH mixtures or to individual PAH's. In the absence of data for oral exposure to defined PAH mixtures, assessment of exposure to such mixtures must be based on or derived from exposure to a single PAH. However, it is not possible to determine 24-hour or 7-day SNARL's for this class of compounds because the acute toxicity data for pure PAM's and PAH mixtures are inadequate. Moreover, be- cause many of the PAM's are highly mutagenic and potent carcinogens in animals and good evidence implicates them in human cancer, chronic SNARL's also cannot be determined. However, the solubilities of the PAM's are quite low, approximately 10 ng/liter for BaP (Andelman and Snodgrass, 1974), and their presence in drinking water probably con- stitutes only a small percentage (O.l~o) of the daily exposure of humans to PAM's (Shabad and Iltnitskii, 1970~. 24-Hour Exposure Insufficient data are available for calculation. 7-Day Exposure Insufficient data are available for calculation. Chronic Exposure No chronic SNARL can be calculated because many of the PAM's are proven carcinogens in animals and are strongly suspected as being carcinogenic in humans. 2,4,6-Trichtorophenol (C6H3CI3O) This compound was reviewed previously by the Criteria and Standards Division, Office of Water Planning and Standards, U.S. Environmental Protection Agency (1979b). The committee has reviewed and discussed that document for accuracy and completeness. The following pages sum- marize those findings and are augmented by additional information deemed necessary by the committee. 2,4,6-Trichlorophenol, also called Dowcide IS, Penaclor, and Omal, is used commercially as a fungicide, slimicide, and bactericide, as a preser- vative for wood and glue, and as a protection against mildew in textiles (Windholz et al., 1976~. Production of this chemical was discontinued in 1975 by the Dow Chemical Company, the only manufacturer of 2,4,6-tri- chlorophenol in the United States, because of the high cost of removing highly toxic dioxin impurities. 2,4,6-Trichlorophenol was prepared commercially by direct chlorina- tion of phenol (Kirk and Othmer, 1967~. Polychlorinated dioxins and

Toxicity of Selected Organic Contaminants in Drinking Water 265 dibenzofurans may be formed during the chemical synthesis of certain chlorophenols, and these highly toxic contaminants were present in technical 2,4,6-trichlorophenol (Rappe et al., 1979~. Heating or burning 2,4,6-trichlorophenol also resulted in the formation of dioxins (Larger et al., 1973; Rappe et al., 1979~. 2,4,6-Trichlorophenol is formed along with other chlorophenols by the chlorination of water or sewage containing phenol (Burttschell et al., 1959~. In the Netherlands, Piet and De Grunt (1975) found unspecified isomers of trichlorophenols in surface waters in concentrations ranging from 0.003 to 0.1 ~g/liter. 2,4,6-Trichlorophenol is also a prominent constituent of effluent from pulp mills, where it is formed during the multistep bleaching processes that remove colored breakdown products of lignin from the pulp (Landner et al., 1977; Lindstrom and Nordin, 19761. In addition, it is found as a degradation or metabolism product of the pes- ticides hexachlorobenzene (Engst et al., 1976) and lindane and its isomers (poster and Saha, 1978; Stein et al., 1977; Tanaka et al., 1977~. 2,4,6-Trichlorophenol has a strong phenolic odor. In water, the threshold for this odor is 100 ~g/liter at 30°C (Hoak, 1957~. This com- pound is acidic, having a pKa of 6.4 (Dodgson et al., 1950~. It has a melting point of 69°C and a boiling point of 246°C, and is soluble in water to 90 mg/liter (Roberts et al., 1977; Windholz et al., 19761. Its vapor pressure is 1 mm Hg at 75°C (Weast and Astle, 1978~. METABOLISM As to be expected from its high octanol-water partition coefficient of 4898:1 (Roberts et al., 1977), 2,4,6-trichlorophenol readily accumulates in fish (Landner et al., 1977) and penetrates the human epidermis (Roberts et al., 1977) and the rabbit eye (Ismail et al., 19751. Little information is available on the absorption, distribution, excre- tion, and metabolism of 2,4,6-trichlorophenol in mammals. Korte et al. (1978) showed that it cleared rapidly from rats, most if it being excreted in the urine. They added a 1 ,ug/g concentration of labeled phenol to the diet of rats for a 3-day period. Eighty-two percent of the dose was excreted in the urine and 22~o in the feces. Radioactive trichlorophenol was not detected in liver, lung, or fat samples examined 5 days after the last dose was administered. Daly et al. (1965) reported that the phenol is metabo- lized to 3,5-dichlorocatechol in the rabbit. Studies by Dodgson et al. (1950) in the rabbit have shown that 2,4,6-trichlorophenol does not form a sulfate conjugate in vivo because of its low pK value.

266 DRINKING WATER AND HEALTH HEALTH ASPECTS Observations in Humans No data are available. Observations in Other Species Acute [;ffects The acute oral LDso for 2,4,6-trichlorophenol in the rat is 820 mg/kg, whereas the LDso for an intraperitoneal dose in the same species is 276 mg/kg (Farquharson et al., 1958; Lewis and Tatken, 19781. In other studies in the rat, the compound produced convulsions when in- jected intraperitoneally (Farquharson et al., 1958) and lethal doses pro duced restlessness, hyperpyrexia, increased rates of respiration, tremors, dyspnea, and coma, continuing until death (Patty, 1963~. All of the trichlorophenol isomers stimulate oxygen consumption by the rat brain in vitro and produce hyperpyrexia in vivo (Farquharson et al., 19581. Onset of rigor mortis in 2,4,6-trichlorophenol-intoxicated animals is said to be characteristically rapid. 2,4,6-Trichlorophenol is a potent uncoupler of oxidative phosphoryla- tion, as reflected by the 18 AM value for 50~o inhibition of mitrochondrial ATP formation in vitro (Mitsuda et al., 1963~. Hexokinase and lactate dehydrogenase are also inhibited by low concentrations of 2,4,6-tri- chlorophenol in vitro (Stockdale and Selwyn, 1971~. Subchronic and Chronic Effects In preliminary subchronic feeding studies, diets containing 2,4,6-trichlorophenol were given ad libitum to groups of male and female F334 rats and B6C36~ mice for 7 weeks (Na- tional Cancer Institute, 1979~. The investigators continued to observe the animals for an additional week after the feeding was stopped. The diets given to the rats contained from 10,000 to 46,000 ~g/g technical grade 2,4,6-trichlorophenol; those given to the mice contained from 6,800 to 31,500 ,ug/g. A significant reduction in growth rate was observed in rats fed 10,000 Agog and in male mice receiving 14,700 ~g/g. Assuming the rats weighed 0.4 kg and that they consumed 0.02 kg food per day (Na- tional Academy of Sciences, 1977), the minimum toxic dose for rats was 500 mg/kg/day. Adult male Sprague-Dawley rats were given daily oral doses of 2,4,6-tri- chlorophenol at 0, 25, 100, and 200 mg/kg for 14 days with no adverse ef- fect on hepatic monooxygenase and UDP-glucuronyl transferase activities or on cytochrome P-450 levels (Carlson, 1978~. In vitro, the compound in- hibited the pesticide O-ethyl O-p-nitrophenyl phenylphosphonothioate

Toxicity of Selected Organic Contaminants in Drinking Water 267 (EPN) detoxification and demethylation of p-nitroanisole. Measurement of hepatic glucose-6-phosphatase and serum sorbitol dehydrogenase pro- vided no evidence of hepatotoxicity. Mutagenicity The mutagenicity of 2,4,6-trichlorophenol was tested in Salmonella typhim?'r~um strains TA98, TA100, TA1535, and TA1537 (Rasanen et al.' 1977~. The compound failed to produce a significant in- crease In revertant colonies, either in the presence or absence of rat liver S-9 fraction. Fahrig et al. (1978), however, found that 2,4,6-trichloro- phenol concentrations of 400 mg/liter increased the mutation rate in the strain of Saccharomyces cerevisiae, but there was no effect on the rate of intragenic recombination. By contrast, Simmon et al. (1978) failed to detect mutagenicity of 2,4,6-trichlorophenol in assays with S. cerevisiae or in the standard Salmonella microsome assay with strains TA1535, TA1537, TA1538, TA98, and TA100. Carcinogenicity Boutwell and Bosch (1959) showed that application of 2,4,6-trichlorophenol to the skin over a lS-week period failed to increase the incidence of papillomas in mice pretreated with the initiator, 7,12-dimethylbenzofajanthracene. Oral administration of technical 2,4,6- trichlorophenol to C57BL/6 and C3H/Anf mice at a dose of 32 to 100 ma/ kg/day for approximately 18 months resulted in an elevation of tumor in- cidence, but the results were not regarded as definitive (Innes et al., 1969). In a more recent study, technical 2,4,6-trichlorophenol was ad- ministered in the diet of male and female F344 rats and male B6C3F~ mice at concentrations of 5,000 and 10,000 ~g/g for 105 to 107 weeks (National Cancer Institute, 1979~. Female B6C3F~ mice were initially fed 2,4,6-trichlorophenol at 10,000 or 20,000 ~g/g, but at 38 weeks, because of reduced weight gain, the dieta~y levels were reduced to 2,500 and 5,000 ,ug/g, respectively. Administration of the lowered dosage rate was con- tinued for 67 weeks. The time-weighted average doses for the female mice was either 5,214 or 10,428 ,ug/g, respectively. Mean body weights of dosed rats and mice of each sex were lower than the corresponding controls and were dose related. The investigators concluded that under the conditions of this bioassay technical 2,4,6-trichlorophenol was carcinogenic in male F344 rats, inducing lymphomas or leukemias. The test chemical was car- cinogenic in both sexes of B6C3F~ mice, inducing hepatocellular car- cinomas or adenomas. The dioxin content of the 2,4,6-trichlorophenol used in these studies was not reported. Teratogenicity No data available.

268 DRINKING WATER AND HEALTH CONCLUSIONS AND RECOMMENDATIONS Suggested No-Adverse-Response Level (SNARLJ The acute toxicity of 2,4,6-trichlorophenol is relatively low, but there is only very limited information on its subchronic or chronic toxicity. Technical 2,4,6-trichlorophenol has been shown to be carcinogenic to rats and mice; however, it is not clear whether this activity is derived from dioxins or other impurities present in the commercial preparation. 24-Hour Exposure The minimum toxic dose for 2,4,6-trichlorophenol in the rat is 500 mg/kg (National Cancer Institute, 19791. Using this value, an uncertainty factor of 1,000, assuming that 2 liters of drinking water daily provides the only source during this period for a 70-kg human, one may calculate the 24-hour SNARL as: 500 mg/kg X 70 kg = 17.5 mg/liter. 7-Day Exposure No data are available for calculation. Using the acute 24-hour SNARL of 17.5 mg/liter for 2,4,6-trichlorophenol and dividing by 7 days: 17.5 mg/1iter = 2.5 mg/liter. Chronic Exposure No chronic SNARL can be calculated because tests with technical 2,4,6-trichlorophenol have yielded positive results in · · · · ~ aroma carcinogen City ~ Assays. REFERENCES Adler, B., R. Braun, J. Schoeneich, and H. Boehme. 1976. Repair-defective mutants of Pro- teus mirabilis as a prescreening system for the detection of potential carcinogens. Biol. Zentralbl. 95:463-469. [Chem. Absts. 86:26749, 1977.] Agustin, J.S., and C.Y. Lim-Sylianco. 1978. Mutagenic and clastogenic effects of chloroform. Bull. Philipp. Biochem. Soc. 1:17-23. Aitio, A. 1973. Glucuronide synthesis in the rat and guinea pig lung. Xenobiotica 3:13-22. Amdur, M.L. 1959. Accidental group exposure to acetonitrile-A clinical study. J. Occup. Med. 1:627-633.

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