Document jJvrXzQ24eVbVd0Y2Z2r3MMp

PB93154680 GREAT LAKES WATER QUALITY INITIATIVE CRITERIA DOCUMENTS FOR THE PROTECTION OF HUMAN HEALTH (February 1993 Draft) REPRODUCED BY: lV fe U.S, Department o l Commerce National Teehnieel Information Service Springfield, Virginia 22161 2/26/93 GREAT LAKES WATER QUALITY INITIATIVE WATER QUALITY CRITERIA FOR PROTECTION OF HUMAN HEALTH CRITERIA DOCUMENTS Benzene ................................................ . . . . 1 C h l o r d a n e .................................... .. Chlorobenzene ......................................... . . . . 17 Cyanides .............................................. Dieldrin .............................................. . . . . 43 2 , 4-Dimethylphenol ................................. . . . . 51 2 , 4-Dinitrophenol .................................... . . . . 57 Heptac h l o r ........................................... H e x a c h l o r o b en z e n e ................... . ............ H e x a c h l o r o e th a n e .................................... . . . . 79 Lindane ................................................ M e r c u r y ................................................ . . . 101 M e t h y l e n e chloride ................................. PCBs (class) ......................................... P e n t a c h l orophenol .................................... 2 , 3 , 7 , 8 --T C D D . . . ................................. . . . 143 T o l u e n e ............................... .. .............. . . . 157 T o x a p h e n e ............................................. T r i c h l o r o e t hy l e n e .................................... October 23, 1991 GREAT LAKES INITIATIVE TIER I HUMAN HEALTH CRITERIA FOR BENZENE CAS NO. 71-43-2 Tier 1 Human Noncancer Criterion Acute exposure to benzene vapors by humans is often associated with neurotoxicity characterized by loss of sensation, vertigo, headache and depression of the central nervous system. Hematopoietic toxicity involving changes in the bone marrow, spleen, and thymus has been associated with benzene exposure. Benzene has also been found to cause embryo/fetotoxicity in experimental animals (EPA, 1980). There are few chronic or oral studies available which examine the noncarcinogenic effects of benzene. The subchronic oral study by Wolf et al. (1956) was considered appropriate for Tier 1 HNC derivation. In this study, female Wistar rats in groups of 10 were administered benzene in olive oil by gavage for 5 days/week for six months. A g r o u p of 20 rats served as controls. Dose levels we r e 0, 1, 10, 50, and 100 m g / k g / b w / d a y . During this p e riod 1 hematologic examinations were performed on selected animals. Growth, body weight, organ weights, behavior, urea nitrogen in blood, histopathological changes and bone marrow counts were also evaluated. Rats exposed to 50 and 100 mg/kg/day exhibited leukopenia and erythrocytopenia, whereas these effects were marginal in the 10 mg/kg/day group. The l mg/kg/day dose/level was considered the NOAEL for this study. This dose is equivalent to 0.71 mg/kg/day after being adjusted for exposure for only 5 days/week. The findings of the Wolf et al. (1956) study are supported by the results of a chronic study by NTP (1986). In this study, C 5 7 B L 1 6 N mice and F344 rats (50 animals/sex/group) were a d m i n i s t e r e d benzene orally at doses of 0, 50, 100 or 200 mg/kg, 5 days/week, for 103 weeks. An additional group consisting of female rats and mice of both sexes were administered 25 mg/kg. Blood was taken for analysis from 10 animals/sex/group at various times during the study. The results of the study showed dose-related leukopenia in rats and mice of each sex for the first 18 months of the study. However, at 24 months, the numbers of white blood cells in high dose male rats, high dose female rats, and mid-dose male mice were higher than controls. Numbers of white blood cells in dosed female mice were not significantly different from controls. 2 Hematopoietic toxicity of benzene following exposure via inhalation was reported by Deichmann et al. (1963). In this study, groups of male and female Sprague-Dawley rats were exposed for 5 hours/day, 4 days/week for periods ranging from 5 weeks to 7 months to benzene concentrations of 0, 15, 29, 31, 44, 47, 61, 65, or 831 ppm. Several hemat o l o g i c parameters, and other parameters including body weight, food intake and blood benzene levels were determined periodically during the study. Following 2-4 weeks of exposure, groups exposed to benzene concentrations of 61 to 831 ppm demonstrated a significantly increased level of leukopenia. Hematopoietic effects were moderate in groups exposed to 44 to 47 ppm for 5-8 weeks. Exposure to 31 ppm benzene for over 4 months did not induce changes in the hematopoietic system and was considered a NOAEL for this study. Based on the conditions of exposure and an assumed absorption factor of 50% (EPA, 1987), a N O A E L of 2.35 mg/kg/day was calculated. This value is comparable to the value calculated in the Wolf et al. (1956) study. EPA (1985) suggested that the Wolf et al. (1956) and Chang (1972) studies could be used to establish a range of acceptable d a i l y intake (ADI) values. In the C h a n g (1972) study 119 workers occupationally exposed to benzene were examined. Hematologic al abnormalities we r e rep o r t e d in 28 of the workers exposed to benzene. These abnormalities 3 included 21 workers with anemia, 2 with leukopenia, and 5 with anemia and leukopenia. Based on an estimate of exposure duration and benzene concentration the researcher derived an exponential function which suggested a threshold level of 10 p p m for h e m a t o l o g i c effects. This study was not used for criterion development due to the route of exposure and the absence of reliable data on the actual exposure concentrations for the individual employees. Rozen et al. (1984) examined the effect of inhalation exposure to 0, 10, 31, 100 and 301 pp m benzene on B- and Tlymphocyte mitogen-induced blastogenesis in C57B1 mice. Exposure to benzene at all doses for 6 hours/day for 6 days resulted in a significant depression in femoral lipopolysaccharide (LPS)-induced B-colony forming ability while total numbers of B-lymphocytes were significantly depressed at 100 and 300 ppm. Splenic phytohemagglutinin (PHA)-induced blastogenesis was significantly depressed at 31, 100 and 300 p p m w h i l e total numbers of T - l y m p h o c y t es were significantly depressed at 100 and 300 ppm. This study was not used for risk assessment because it used the inhalation route of exposure and it is questionable whether the effects produced in this study are biologically significant and adverse. 4 Few studies using the oral route of exposure have examined the reproductive/developmental effects of benzene. Nawrot and Staples (1979) administered 0.3, 0.5 or 1.0 ml/kg/day (790, 1320 and 2640 mg/kg/day, respectively) benzene to pregnant CD-I mice during days 6-15 or 12-15 of gestation. Despite some maternal lethality and embryonic resorptions at the two higher doses, no evidence of teratology was seen. ADE = 0.71 mq/kg/d = 0.00071 mg/kg/d 1000 Where: Uncertainty Factor = 1000, composed of: 10 x for interspecies variability 10 x for intraspecies variability 10 x for subchronic to chronic extrapolation Drinking Water Sources: HNV = ADE x Wh x RSC = 0.00071 ma/kcr/d x 70 ka_________ w c + (FC X BAF) 2 1/d + (0.015 kg/d x 13 1/kg*) = 0.0226 mg/1 (rounded off to 0.02 mg/1 (Tier 1)) 5 Nondrinking Water Sources: HNV = A P E x Wh x RSC = 0.00071 ma/kcr/d x 70 ka________ W C + (FC x BAF) 0.01 1/d + (0.015 kg/d x 13 1/kg*) = 0.242 mg/1 (rounded off to 0.2 mg/1 (Tier 1)) Where: *BAF = 13 1/kg, p r ovided by EPA-Duluth and Minnesota PCA. Note: A relative source contribution (RSC) factor has not been utilized in these draft calculations. REFERENCES: Chang, I.W. 1972. Study on the threshold limit value of b e n zene and early diagnosis of benzene poisoning. J. Cath. Med. Coll. 23:429. Deichmann, W.B., W.E. MacDonald, and E. Bernal. 1963. hematopoietic tissue toxicity of benzene vapors. Toxicol. Appl. Pharmacol. 5:201-224. The International Agency for Research on Cancer (IARC). 1982. IARC Monograph: Evaluation of the Carcinogenic Risk of Chemicals to Humans, Volume 29, WHO Publications Center, USA, Albany, NY, pp 1-416. National Toxicology Program (NTP). 1986. Toxicology and Carcinogenesis Studies of Benzene in F344/N Rats and B6C3F1 Mice (Gavage Studies). NTP Technical Report Series. 289., National Toxicology Program, Research Triangle Park, NC. Nawrot, P.S. and R.F. Staples. 1979. Embryo-fetal toxicity and teratogenicity of benzene and toluene in the mouse. Teratology. 1 9 :41a. 6 Rozen, M.G., C .A. Snyder, and R.E. Albert. 1984. Depressions in B- and T-lymphocyte mitogen-induced blastogenesis in mice exposed to low concentrations of benzene. Toxicol. Letters. 20:343-349. U.S. Environmental Protection Agency (EPA). 1987. Integrated Risk Information System (IRIS database). Chemical file for benzene (CAS No. 71-43-2). Verification Date 10/9/87. Last Reviewed 10/9/87. U.S. Environmental Protection Agency (EPA). 1985. Drinking Water Criteria Document for Benzene (Final Draft). U.S. Environmental Protection Agency, PB86-118122. Washington, D.C. U.S. Environmental Protection Agency (EPA). 1980. Water Quality criteria Document for Benzene. Washington, DC. EPA 440/5-80-018. Ambient Wolf, M.A., V.K. Rowe, D.D. McCollister, R.L. Hollingsworth and F. Oyen. 1956. Toxicological studies of certain alkylated benzenes and benzene. A.M.A. Arch. Indust. Health. 14:357-398. Tier l Human Cancer Criterion According to the weight-of-evidence method for the classification of carcinogens, benzene is a Class A carcinogen (known human carcinogen) (IARC, 1982; EPA, 1987; 1989). This is based on sufficient evidence from epidemiologic studies on the incidence of non-lymphocytic leukemia from occupational exposure, and increased incidence of neoplasms in rats and mice exposed to benzene by inhalation and gavage (EPA, 1987). In addition, numerous studies have found a significant increase in chromosomal aberrations of bone marrow cells and peripheral lymphocytes from workers exposed to benzene (IARC, 1982). The data are sufficient to derive a Tier 1 HCC for benzene. 7 Numerous epidemiologic and case studies have shown a relationship between leukemia and exposure to benzene (IARC, 1982). The oral slope factor for benzene based on human data is estimated to be 2.9E-2 (mg/kg/day)-1 (EPA, 1987). The unit risk estimate is based on the geometric mean of four maximum likelihood point estimates using pooled data from the studies of Rinsky et al. (1981) and Ott et al. (1978)/ which was then adjusted for the results from the Wong et al. (1983) study as described by EPA (1987). The slope factor of the dose-response curve for the carcinogenic effects of benzene by the oral route, 2.9E-2 (mg/kg/day)-1 is used in the calculation of the HCC for benzene. RAD = l x 10 5______________ 2.9E-2 (mg/kg/d)-1 = 0.0003448 mg/kg/d - 0.00034 mg/kg/day Drinking Water Sources: H C V = R A D X Wh________ = 0.00034 mq/ k q / d x 70 kg_________ W C + (FC X BAF) 2 1/d + (0.015 kg / d x 13 1/kg*) = 0.0108 mg/1 (rounded off to 0.01 mg/1 (Tier 1)) 8 Nondrinking Water Sources: H C V = RAD x Wh________ = 0.00034 ma/ka/d x 70 ka______________ WC + (FC x BAF) 0.01 1/d + (0.015 kg/d x 13 1/kg*) = 0.116 mg/1 (rounded off to 0.1 mg/1 (Tier 1)) Where: PCA. *BAF = 13 1/kg provided by E PA-Duluth and M i n n esota REFERENCES: International Agency for Research on Cancer (IARC). 1982. IARC Monograph: Evaluation of the Carcinogenic Risk of Chemicals to Humans; Volume 29, W HO Publications Center, USA, Albany, NY, pp 1-416. Ott, M.G., J.C. Townsend, W.A. Fishbeck and R.A. Langner. 1978. Mortality among individuals occupationally exposed to benzene. Arch. Environ. Health., 33: 3-10. Rinsky. R . A . , R.J. Young and A.B. Smith. 1981. Leukemia in benzene workers. Am. J. Ind. Med. 2: 217-245. U.S. Environmental Protection Agency (EPA). 1987. Health Effects Assessment for Benzene. EPA/600/8-89/086. Cincinnati, OH. U.S. Environmental Protection Agency (EPA). 1987. Integrated Risk Information System (IRIS database). Chemical file for benzene (CAS No. 71-43-2). Verification Date 10/9/87. Last Reviewed 10/9/87. U.S. Environmental Protection Agency (EPA). 1980. Water Quality Criteria Document for Benzene. Washington, DC. EPA 440/5-80-018. Ambient Wong, O . , R.W. Morgan and M.D. Wharton. 1983. Comments on the NIOSH study of leukemia in benzene workers. Technical report submitted to Gulf Canada, Ltd., by Environmental Health Associates. 9 10 S e p tember 6, 1991 GREAT LAKES INITIATIVE HUMAN HEALTH CRITERIA FOR CHLORDANE CAS NO. 57-74-9 Tier l Human Noncancer Criterion A review of the available literature indicates that the most appropriate study for HNC derivation for chlordane is a study conducted by Velsicol Chemical Corporation (1983a). In this study, 80 Fischer 344 rats of each sex were admini s t e r e d 0, 1, 5 or 25 p p m chlordane for 130 weeks. Hematological, biochemical, urinary and pathological measurements were made on eight animals/sex/group at weeks 26 and 52. The same mea s u r e m e n t s were m a d e on animals which survived to week 130. Liver hypertrophy occurred in females at 5 ppm (0.273 mg/kg/d) and a NOEL of 1 ppm (0.055 mg/kg/d) was determined. No liver lesions were found in male rats and a NOEL of 25 ppm (0.1175 mg/kg/d) was determined. The NOEL for female rats is slightly lower than the NOELs determined for mice and dogs. A 24-month chronic study in ICR mice found NOELs of 0.123 mg/kg and 0.138 mg/kg for male and female mice, respectively (Velsicol, 1983b). A study using Beagle dogs found NOELs of 0.06 mg/kg and 0.09 mg/kg for male and female dogs, respectively (Wazeter, 1967). '\ Preceding Page Blank 1 ll Data on the reproductive and developmental effects of chlordane are limited. ATSDR (1989) cited a study by Usami et al. (1986) which showed no malformations in pups whose dams were administered 20, 40 or 80 mg/kg/d ch l o r d a n e from d a y 7 to 17 of gestation. The finding of this study suggests that exposure levels which may cause adverse effects on development are higher than the NOEL cited above for the Velsicol study (1983a). Other studies reported by Chernoff and Kavlock (1982) and Ingle (1952 as cited by EPA, 1990) also indicate that criteria derived from the chronic NOAEL of 0.055 mg/kg/d (Velsicol, 1983a) should be protective of potential developmental effects. The quality of the Velsicol (1983a) rat study was deemed sufficient to derive a Tier 1 HNC. This study was also used by EPA (1989; 1990) to derive the oral RfD for chlordane. The HNC was derived from the female rat NOEL using an uncertainty factor of 100 to account for intraspecies variability and interspecies extrapolation. ADE = 0.055 mq/kq/d = 0.00055 mg/kg/d = 550 ng/kg/d 100 Where: U n c e r tainty factor = 100, composed of: lOx for intraspecies variability lOx for interspecies extrapolation 12 Drinking Water Sources: HNV = ACE x l/i x RSC = 550 ng/kg/d x 70 kg x 0.8__________ UC + (FC x BAF) 2 L/d + <0.015 kg/d x 219,375 l/kg*> = 9.35 ng/l (rounded off to 9 ng/l (Tier 1)) Nondrinking Water Sources: HNV = APE x Uh x RSC = 550 noAo/d x 70 kg x 0.8_____________ UC + (FC x BAF) 0.01 l/d + (0.015 kg/d x 219,375 l/kg*) = 9.36 ng/l (rounded off to 9 ng/l (Tier 1)) Where: RSC = 0.8; the substance is persistent and bioaccumulative. *BAF = 219,375, provided by EPA-Duluth and Minnesota PCA. REFERENCES Agency for Toxic Substances and Disease Registry (ATSDR). 1989. Toxicological Profile for Chlordane. Department of Health and Human Services. U.S. Public Health Service. Chernoff, N. and R. J. Kavlock. 1982. An in v i v o teratology screenu t i l i zi n g pr e g n a n t mice. J. Toxicol. Environ. Hlth. 10:541-550. Ingle, L. 1952. Chr o n i c oral t o xicity of c h l o rdane to rats. Arch.Ind. Hyg. Occup. Med. 6:357-367. U.S. Environmental Protection Agency (EPA). 1989. Integrated Risklnformation System (IRIS database). Chemical file for chlordane (57-74-9). Verification Date 3/22/89. Last Revised 7/1/89. 13 U.S. Environmental Protection Agency (EPA). 1990. Drinking Water Criteria Document for Heptachlor, Heptachlor Epoxide and Chlordane. Revised November, 1990. ECAOCIN-406. Usami, M . , K. K a w a s h i m a , S. Nakaura, et al. 1986. Effects of chlordane on prenatal development of rats. (Abstract). Eisei Shikenso Hokoku. 104:68-73. Velsicol Chemical Corporation. 1983a. Yonemura, T . , F. T a k amura and Y. Takahashi. Thi r t y - m o n t h chronic toxicity and tumorigenicity test in rats by chlordane technical. (Unpublished study by Research Institute for Animal Science in Biochemistry and Toxicology, Japan). V e l sicol Chemical Corporation. 1983b. Inui, S., K. Yamazaki, T. Yonemura, et al. T w e n t y - f o u r month chronic toxicity and tumorigenicity test in mice by chlordane technical. (Unpublished study by Research Institute for Animal Science in Biochemistry and Toxicology, Japan). Wazeter, F.X, 1967. Two-Year Chronic Feeding Study in the Beagle Dog. Sponsored by Velsicol Chemical Corporation (Unpublished). Tier 1 Human Cancer Criterion There are inadequate data available to determine whether chlordane is a human c arcinogen (EPA, 1987; 1990). Chronic studies using four strains of mice (CD-I, B6C3F1, C5781/6N, ICR) of both sexes have shown an increase in the occurrence of liver tumors. Additional weight-of-evidence for chlordane's carcinogenicity is provided by its structural similarity to other compounds (i.e., dieldrin and heptachlor) which have been found to induce hepatocellular carcinomas in mice. The results of various mutagenicity studies are inconclusive as to this chemical's ability to cause mutagenic effects. The weight-of-evidence for 14 chlordane carcinogenicity is sufficient for B2 (probable human carcinogen) clas s ification (EPA, 1986; 1987; 1990). The data are sufficient to derive a Tier 1 HCC. Two key studies (NCI, 1977; Velsicol, 1973 as cited in EPA, 1986) found a significant increase in hepatocellular carcinomas in treatment groups when compared to controls. Both studies also showed a dose-response relationship between exposure of mice to chlordane and the occurrence of liver tumors. EPA (1986; 1987; 1990) calculated four separate slope factors from these key studies, and derived a recommended slope factor of 1.3 (mg/kg/d) ^ from the geometric mean of these slope factors. This method of computing a slope factor is used "in situations where no single study is judged most appropriate, yet several studies collectively support the estimate ..." (EPA, 1989). According to EPA (1989), the advantage of this method of determining the slope factor is that all relevant data are used in the computations. -5 -6 RAD = 1 x 10-- ______ = 7.7 x 10 mg/kg/d = 7.7 ng/kg/d 1.3 (mg/kg/d) 1 Drinking Water Sources: HCV s RAD x Uh______ = 7.7 na/ka/d x 70 kg_________________ WC + (FC x BAF) 2 l/d + (0.015 kg/d * 219,375 l/fcg*) 15 0.16 ng/l (rotnded off to 0.2 ng/l (Tier 1)) Nondrinking Water Sources: HCV RAD x Wh______ = 7.7 no/ka/d x 70 kg_____________________ WC + <FC x BAF) 0.01 l/d + <0.015 kg/d x 219,375 l/kg*> 0.16 ng/l (rotxided off to 0.2 ng/l (Tier 1)) *BAF = 219,375, provided by EPA-Duluth and Minnesota PCA. REFERENCES: National Cancer Institute (NCI). 1977. Bioassay of Chlordane forpossible carcinogenicity. NCI Carcinogenesis Tech. Rep. Ser. No. 8. U.S. DHEW Publ. No. (NIH) 77-808. Bethesda, MD. U.S. Environmental Protection Agency (EPA). 1986. CarcinogenicityAssessment of Chlordane and Heptachlor/Heptachlor Epoxide. Carcinogen Assessment Group. Office of Health and Environmental Assessment, Washington, DC. U.S. Environmental Protection Agency (EPA). 1987. Integrated Risk Information System (IRIS database). Chemical file for chlordane (57-74-9). Verification Date 4/1/87. Last Revised 1/1/91. U.S. Environmental Protection Agency (EPA). 1989. Risk A s s e s s m e n t Guidan c e for Superfund. V o l u m e 1. Human Health Evaluation Manual (Part A). Interim Final. OERR. EPA/540/1-89/002. U.S. Environmental Protection Agency (EPA). 1990. Drinking Water Criteria Document for Heptachlor, Heptachlor Epoxide and Chlordane. Revised November, 1990. ECAO-CIN-4 06. 16 N o v e m b e r 1, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR CHLOROBENZENE CAS NO. 108-90-7 Tier l Human Noncancer criterion A review of the available literature indicates inadequate human data for quantitative risk assessment of chlorobenzene. Humans exposed occupationally to chlorobenzene intermittently for up to 2 years displayed signs of neurotoxicity including numbness, cyanosis (from depression of the respiratory center), hyperesthesia and muscle spasms (Rozenbaum, 1947, as cited in ATSDR, 1990). While these findings provide qualitative evidence that chlorobenzene is potentially neurotoxic, specific exposure levels are not available to support quantitative risk assessment. Several animal studies were identified during a review of the available literature. The most appropriate basis for HNV derivation for chlorobenzene is the NOAEL from a subchronic dog study by Monsanto Company (1967). Young adult pure-bred beagle dogs (4/sex/group) were administered chlorobenzene in gelatin capsules, 5 days/week at doses of 17 0, 0.025, 0.050 and 0.250 ml / k g / d a y (converted per EPA (1989a) to 0, 27.25, 54.5 and 272.5 mg/kg/day, respectively) for 13 weeks. Four high-level dogs died or were sacrificed in moribund condition between the third and fourth weeks of the study. These deaths were preceded by anorexia, decreased activity, body weight loss, cachexia and coma. The four surviving high-level dogs exhibited temporary lack of appetite and body weight loss. Changes in hematology, clinical chemistry and urine analyses were observed at the high dose. Pathologic changes in the liver, kidney, gastrointestinal mucosa, and hematopoietic tissue were also observed in high-dose animals. At 54.5 mg/kg/day, slight hepatic alterations were observed, and severe cellular variations in the epithelium of the terminal proximal tubule in the kidney were also observed. The NOAEL and LOAEL for this study were 27.25 and 54.5 mg/kg/day, respectively. The database is judged to be sufficient for Tier 1 HNC derivation. The key study (Monsanto, 1967) provides a subchronic NOAEL which is supported by additional data. In a rat study performed by Monsanto Company and abstracted by Knapp et al. (1971), significant elevations were noted in liver and kidney weights of rats administered Dietary levels of 100 and 250 mg/kg/ d a y for 93 to 99 consecutive days. No remarkable histopathologic findings were reported. No 18 effects were noted among rats receiving 12.5 or 50 mg/kg/day. The NOAEL and LOAEL for this study were 50 and 100 mg/kg/day, respectively (EPA, 1 9 8 9 a ) . The authors concluded that in comparison to rats, dogs displayed a higher sensitivity to chlorobenzene toxicity. In a chronic study by the National T o x i c o l o g y Pro g r a m (NTP, 1985), groups of F344/N rats and B6C3F1 mice (50/sex/dose) were administered chlorobenzene by gavage in corn oil 5 days/week for 103 weeks. The male and female rats received 0, 60 or 120 mg/kg/day; the female mi c e r e c e i v e d 0, 60 or 120 m g / k g / d a y and the male mice rec e i v e d 0, 30 or 60 mg/kg/day. A statistically significant decrease in the survival of high-dose male rats was observed. Histological examination of the liver showed hepatocellular necrosis, graded as minimal to mild, in all groups. In male mice, mortality at both dose levels was increased to a statistically significant level. No other chlorobenzenerelated clinical toxicity was observed in mice. Several reproductive and developmental studies have been performed with chlorobenzene. John et al. (1984) exposed Fischer 344 rats and New Zealand White rabbits by inhalation to chlorobenzene for 6 h o u rs/day at doses of 0, 75, 210 or 590 ppm during periods of major organogenesis. Exposure to 590 ppm caused elevated liver weights in both species and 19 decreased body weight gain and feed consumption in rats. The developmental NOAEL was 590 ppm for both species (equivalent to 216 mg/kg/day for rats and 125 mg/kg/day for rabbits, as per EPA, 1969a). The maternal NOAELs were 210 ppm (equivalent to 77 mg/kg/day, as per EPA, 1989a) for rats and 75 ppm (equivalent to 16 mg/kg/day, as per EPA, 1989a) for rabbits. In a two-generation reproduction inhalation study in rats (Nair et al., 1987) groups of 30 male and 30 female sprague- Dawley CD rats (FQ generation) were exposed to chlorobenzene at target c o n c e ntrati on s of 0, 50, 150 or 450 p pm for 10 . weeks prior to mating and during mating, gestation, and lactation. Groups of 30 male and 30 female animals were exposed to the same concentrations of chlorobenzene as the F0 parents, initiated 1 week postweaning and lasting through mating, gestation and lactation. Hepatocellular hypertrophy and renal changes were observed among Fq and F^ male rats exposed to 150 and 450 ppm but exposure of rats to c hlorobenzene at levels of 50, 150 or 450 p p m did not have any adverse effects on reproductive performance or fertility of male and female rats. A reproductive NOAEL of 165 m g / k g / d a y and a systemic N O A E L of 18 m g / k g / d a y (conversions were from EPA, 1989a) were determined from this study. 20 The HNV is derived from the NOAEL dose of 27.25 mg/kg/day (converted to 19.46 mg/kg/day for 5 days/week administration) from the 13-week dog study by Monsanto Company (1967) with an uncertainty factor of 1000. Data from other studies (John et al., 1984; Nair et al., 1987) suggest that this value will be protective of reproductive/developmental effects. This approach is consistent with the risk assessment of chlorobenzene for the derivation of the oral RfD, drinking water equivalent level (DWEL), and maximum contaminant level goal (MCLG) by EPA (EPA,1989a; EPA, 1989b; EPA, 1989c). ADE = 19.46 ma/kq/d = 0.019 mg/kg/day 1,000 Where: Uncertainty Factor = 1,000, composed of: lOx for intraspecies variability lOx for interspecies extrapolation lOx for subchronic to chronic extrapolation Drinking Water Sources: H N V = ADE x W h x RSC = 0,019 m a / k a / d x 70 ka________ WC + (FC x BAF) 2 1/d + (0.015 kg/d x 49 1/kg*) = 0.49 mg/1 (rounded off to 0.5 mg/1 (Tier 1)) 21 Nondrinking Water Sources: H N V = A P E x Wh x RSC = 0.019 mq/k o / d x 70 kg_________ WC + (FC x BAF) 0.01 1/d + (0.015 kg/d x 49 1/kg*) = 1.79 mg/1 (rounded off to 1.8 mg/1 (Tier 1)) Where: *BAF = 49, pro v i d e d by EPA-Duluth and Mi n n e s o t a PCA. NOTE: A Relative Source Contribution (RSC) factor has not been utilized in these draft calculations. REFERENCES: Agency for Toxic Substances and Disease Registry (ATSDR). 1990. Toxicological Profile for Chlorobenzene. U.S. Public Health Service. ATSDR/TP-90/06. John, J . A . , W.C. Hayes, T.R. Hanley, Jr., K.A. Johnson, T.S. Gushow and K.S. Rao. 1984. Inhalation teratology study on monochlorobenzene in rats and rabbits. Toxicol. Appl. Pharmacol. 76(2):365-373. Knapp, Jr., W.K., W.M. Busey, and W. Kundzins. 1971. Subacute oral toxicity of monochlorobenzene in dogs and rats. Toxicol. Appl. Pharmacol. 19:393. Nair, R.S., J.A. Barter, R.E. Schroeder, A. Knezevich, and C.R. Stack. 1987. A two-generation reproduction study with monochlorobenzene vapor in rats. Fundam. Appl. Toxicol. 9(4):678-686. National Toxicology Program (NTP). 1985. Toxicological and Carcinogenesis Studies of Chlorobenzene in F344/N Rats and B6C3F1 Mice (Gavage Studies). NTP-TR No. 261, NIH Publication No. 86-2517. Rozenbaum, N.D., R.S. Blekh, S.N. Kremneva, et al. 1947. [Use of chlorobenzene as a solvent from the standpoint of industrial hygiene.] Gig. Sanit. 12:21-24. (Russian) As cited in ATSDR (1990). 22 U.S. Environmental Protection Agency (EPA). 1989a. Integrated Risk Information System (IRIS database). Chemical file for chlorobenzene (108-90-7). Verification Date 1/19/89. Last Reviewed 1/19/89. U.S. Environmental Protection Agency (EPA). 1989b. Monochlorobenzene. 54 FR No. 97. pp. 22087-22088. 22, 1989. May U.S. Environmental Protection Agency (EPA). 1989c. Health Effects Assessment for Chlorobenzene. PB90-142514/XAD. EPA/60/8-89/099 23 24 October 23, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR CYANIDES CAS NO. 57-12-5 Tier 1 yum Noncancer Criterion A review of the available literature indicates inadequate human data for quantitative risk assessment of cyanides. Qualitative data suggest that chronic dietary exposure to naturally occurring cyanogens in cassava results in thyroid abnormalities in African countries where cassava is a staple crop. Effects seen, including endemic goiter, cretinism and congenital hypothyroidism, were potentiated by low iodine intake and other dietary deficiencies. Tropical ataxic neuropathy has also been linked to chronic cyanide ingestion from cassava derivatives (Njoh, 1990). However, these data do not provide a dose-response relationship (EPA, 1985a; 1985b; ATSDR, 1988; Njoh, 1990). From animal studies on the chronic oral toxicity of inorganic salts of cyanide, the most appropriate basis for HNV derivation for cyanides is the HNOAEL from the chronic rat feeding study of Howard and Hanzal (1955). Weanling rats (10/sex/group) were offered food fumigated with hydrogen cyanide at dietary concentrations of 100 and 300 p p m HCN, averaging 73 and 183 mg C N " /kg diet) for tw o years. Dose levels were approximately 4.6 or 10.8 mg CN~/kg bw/day 25 Preceding Page Blank to females, and 3.6 or 7.5 mg C N " /kg bw/day to males (EPA, 1985b, 1990). At termination, hematological values were within normal limits and neither gross nor microscopic examination of tissues (including thyroid) revealed evidence of pathology due to exposure in any of the exposure groups. Therefore, the HNOAEL from this study is 10.8 mg CN~/kg bw/day. The database is judged to be sufficient for Tier 1 HNC derivation. The key study (Howard and Hanzal, 1955) provides a chronic HNOAEL which is supported and supplemented by other data. In a chronic study (Philbrick et al., 1979), ten male weanling rats were given 1500 ppm potassium cyanide in the diet for 11.5 months. The administered dose was approximately 75 mg KCN/kg bw/day, or 30 mg CN" /kg bw/day (ATSDR, 1988; EPA, 1985a; 1985b). The cyanide exposure in rats receiving either normal or restricted diet resulted in reduced body weight gain, decreased thyroid gland activity and increased thyroid weights. In a study by Tewe and Maner (1981b), pregnant Yorkshire pigs (6/dose group) were fed fresh cassava diets containing 0, 276 or 521 mg cyanide (added as KCN) per kg of fresh cassava offered during gestation and parturition, on the 26 110th day of gestation, two gilts per dose group were sacrificed and the fetuses were evaluated. The remaining gilts in each dose group were allowed to naturally deliver and were then maintained on a diet with no cyanide throughout the 56-day lactation period. No serious interference was observed with the production of the first litter of offspring by gilts receiving cassava diets containing up to 521 ppm added cyanide during gestation. Gilts in the high dose group, exhibited possible adverse effects on the thyroid (increased weight) and kidney (proliferation of glomerular cells). The high exposure level also suggests a LOAEL for developmental effects, as the 110-day-old fetuses had decreased relative spleen, thyroid and heart weights. The evidence that the thyroid may be a sensitive target organ in pigs lends support to the thyroid effects reported by Jackson (1988; see later discussion). However, data interpretation is difficult due to the small number of gilts evaluated (2/dose group). The administered levels of 276 and 521 ppm CN" in the diet convert to 7.7 and 17 mg C N " /kg b w / d a y for the gilts, using the reported animal body weights and food intake from the study. The NOAEL for this study was 276 ppm CN" in the diet, or approximately 7.7 mg CN"/kg bw/day. Another developmental study (Tewe and Maner, 1981a) reports that 500 ppm KCN administered in the diet to rats resulted 27 in a decreased protein efficiency ratio among offspring during the postweaning growth phase. The dose level has been converted to approximately 50 mg CN"/kg/day assuming a 10% food conversion factor (ATSDR, 1988), or approximately 10.6 mg CN~/kg/day per EPA (1985b). Other available oral studies are more limited in design, including the only relevant study with drinking water exposure. Palmer and Olson (1979) gave 7 male SpragueDawley rats 200 mg/1 KCN in water (or 80 mg/1 C N " ; 10 mg C N " /kg/day per EPA, 1985a) or 200 p p m KCN in diet (80 m g C N " /kg food; 8 mg CN" /kg/day per EPA, 1985a) for 21 days. At the end of the study, the only parameters evaluated were body weight and liver weight. Liver weights were increased over controls following drinking water exposure only. Although inadequate for criteria derivation, this study suggests that cyanide via drinking water is more potent for inducing effects than feeding studies, but is less potent than gavage exposures. In another noteworthy but limited study, the effects of oral cyanide on glucose metabolism, thyroid function and an array of behavioral indices were evaluated in miniature pigs (Jackson, 1988). Three swine/dose group (mixed sexes) received 0, 0.4, 0.7 or 1.2 mg C N " /kg/day by intraoral bolus as KCN in aqueous solution, daily for 24 weeks. The author 28 reports that treatment resulted in a dose-related increase in the fasting blood glucose level, dose-related decreases in T 3 and T4 thyroid hormones, and numerous altered behaviors. By Chi-square analyses these changes were determined to be significant, even in the low exposure group with regard to some parameters. The alterations noted were more pronounced in the high-dose group, particularly for thyroid hormone levels, the parameter most clearly indicative of adverse effects. Suppression of T 3 and T4 was dose-related with a much stronger response in the high-dose group (roughly double the effect seen in the mid-dose group at 24 weeks). However, the limitations of the study design and reporting are substantial (Papa, 1990, personal communication). Some of the most critical deficiencies in design or reporting include: small animal numbers per group; lack of body weight and organ weight data; exposure pattern as a single daily bolus dose; unclear biological significance of the reported biochemical effects; no report of the variance about the mean for T 3 , T4 , and blood glucose values; and the distribution of sexes among the four groups was not reported. The latter point appears critical because from the 12 animals d i s tributed evenly among t he four groups, five were females, and seven were castrated males. The castrated males were, therefore, unevenly represented among the groups and the castration effect on thyroid levels 29 and behavior is likely to be significant (Papa, 1990, personal communication). The 2-year dietary exposure study by Howard and Hanzal (1955) has been selected as the key study for the derivation of the risk assessment of cyanide in drinking water by EPA (1985a, 1985b, 1990). Those assessments have applied an additional uncertainty factor of 5 due to the dietary method of exposure. This is intended to account for the relative tolerance to cyanide when it is ingested with food rather than when it is ingested in drinking water. The value of 5 was based on an evaluation of cyanide-binding affinity of food and the GI absorption of cyanide in food versus drinking water by Dr. Ernest Foulkes of the University of Cincinnati who served as an external reviewer for EPA (1985a) (Papa, 1990, personal communication). This 5-fold (or 20%) adjustment factor for differential bioavailability between cyanide in feed and in drinking water is concluded to be appropriate and scientifically supportable in a more recent analysis (Pearsall and Chrostowski, 1990). The HNC is therefore derived from the HNOAEL dose of 10.8 mg CN"/kg/day in rats via feed (Howard and Hanzal, 1955), and an uncertainty factor of 500. This consists of the 5-fold 30 adjustment discussed above, and inter- and intraspecies extrapolation. ADE = 10.8 ma CN~/kq/d = 0.0216 mg/kg/d 500 Where: Uncertainty Factor = 500, composed of : lOx for intraspecies variability lOx for interspecies extrapolation 5x adjustment for bioavailability, feed vs. drinking water Drinking Water Sources: HNV = ADE x Wh x RSC = 0.022 m a / k a / d x 70 kg__________ wc + (FC x BAF) 2 1/d + (0.015 kg/d x 1 1/kg*) = 0.75 mg/1 (rounded off to 0.8 mg/1 (Tier 1)) Nondrinking Water Sources: HNV = A D E x W h x RSC = 0.022 ma/ k a / d x 70 kg______________ WC + (FC x BAF) 0.01 1/d + (0.015 kg/d x 1 1/kg*) = 60.5 mg/1 (rounded off to 60 mg / 1 (Tier 1)) 31 Where: *BAF = 1.0 by default. Per EPA-Duluth and Minnesota PCA, there are insufficient data to derive a BAF. NOTE: A Relative Source Contribution (RSC) factor has not been utilized in these draft calculations. REFERENCES: Agency for Toxic Substances and Disease Registry (ATSDR). 1988. Toxicological Profile for Cyanide. U.S. Public Health Service. ATSDR/TP-88/12. Howard, J. and R. Hanzal. 1955. Chronic to x i c i t y to rats of food treated w i t h hyd r o g e n cyanide. J. Agric. Food Chem. 13:325-329. Jackson, L. 1988. Behavioral effects of chronic sublethal dietary cyanide in an animal model: implications for humans consuming cassava (Manihot esculents) . Human Biology 60(4):597-614. Njoh, J. 1990. Tropical ataxic neuro p a t h y in Liberians. Trop. Geogr. Med. 42(1):92-94. Palmer, I. and O. Olson. 1979. Partial preven t i o n by cyanide of selenium poisoning in rats. Biochemical Biophysical Research Comm. 90(4):1379-1386. Papa, L. 1990. U.S. EPA, ORD, R e s e a r c h Physiologist. Personal communic a t i on w i t h R. Sills, M i c h i g a n Department of Natural Resources. Pearsall, L. and P. Chrostowski. bioavailability of cyanide. Prepared by Clement Assoc., Boston M A . 1990. The oral Unpublished report. Inc., for Boston Gas Co., Philbrick, D. et al. 1979. Effect of p r o l onged cyanide and thiocy a n a t e feeding in rats. J. Toxicol. Env. Health 5:579-592. Tewe, 0. and J. Maner. 1981a. L o n g -term and carry-over effect of dietary inorganic cyanide (KCN) in the life cycle performance and metabolism of rats. Toxicol. Applied Pharmacol. 58:1-7. . 32 Tewe, 0. and J. Maner. 1981b. Performance and pathophysiological changes in pregnant pigs fed cassava diets containing different levels of cyanide. Res. Vet. Sci. 30(2): 147-151. U.S. Environmental Protection Agency (EPA). 1985a. Drinking Water Criteria Document for Cyanide. Final Draft. NT I S . EPA-60Q/X-84-192-1. PB 86-117793. U.S. Environmental Protection Agency (EPA). 1985b. Integrated Risk Information System (IRIS database). Chemical file for Cyanide, free (57-12-5). Verification Date 8/5/85. Last Reviewed 8/5/85. U.S. Environmental Protection Agency (EPA). 1990. Federal Register 55(143):30370-30448. July 25, 1990. National Primary and Secondary Drinking Water Regulations; Synthetic Organic Chemicals and Inorganic Chemicals. . Proposed Rule. 33 34 / S e p t ember 6, 1991 GREAT LAKES INITIATIVE HUMAN HEALTH CRITERIA FOR P ,P 7- D I C H L O R O D I P H E N Y L T R I C H L O R O E T H A N E (DDT) CAS NO. 50-29-3 Tier l Human Noncancer Criterion A review of the available literature indicates that the most appropriate basis for HNC derivation for DDT is the NOAEL from the subchronic rat feeding study of Laug et al. (1950). Weanling rats (15/sex/group) were fed commercial-grade DDT (81% p ,p '- D D T , 19% o,p'-DDT) at levels of 0, 1, 5, 10 or 50 ppm for 15-27 weeks. The critical toxic effect was liver toxicity, demonstrated as relatively mild dose-dependent histopathologic changes in hepatocytes at doses of 5 ppm and higher. These included hepatocellular hypertrophy, increased cytoplasmic oxyphilia, and peripheral basophilic cytoplasmic granules. The NOEL was 1 ppm, or 0.05 mg/kg bw/day assuming a food consumption rate of 5% body weight per day. The LOAEL was 5 ppm (0.25 mg/kg bw/day). The database is judged to be sufficient for Tier 1 HNC derivation. The key study (Laug et al., 1950) provides a subchronic (greater than 90 day) NOEL which is supported and supplemented by other data. In a 2-year rat dietary exposure 35 Preceding Page Blank I study (Fitzhugh, 1948) rats were exposed to 10-800 ppm DDT in feed, resulting in liver lesions at all dose levels with a LOAEL of 10 ppm (0.5 mg/kg bw/day). The available mammalian reproduction and developmental studies of DDT indicate that an HNC derived from the critical effect of liver toxicity will be protective of potential human reproductive/ developmental effects (EPA, 1985). T he HNC is based on the subchronic rat NOEL of 0.05 mg/kg bw/day, with a total uncertainty factor of 100. An uncertainty factor for subchronic to chronic conversion is not included because of the corroborating chronic study in the database. This approach is consistent with the oral RfD development by EPA (1985). ADE = 0.05 mq/ka/d = 0.0005 mg/kg/d = 0.5 ug/kg/d 100 Where: Uncertainty Factor = 100, composed of: lOx for intraspecies variability lOx for interspecies extrapolation Drinking Water Sources: HNV = = ADE X Wh x RSC WC + (FC x BAF) 0.5 ua/kg/d x 70 ka x 0.8_____________ 2 1/d + (0.015 kg/d x 1,913,8751/kg*) 36 0.00098 ug/1 (rounded off to 1.0 ng/1 (Tier 1)) Nondrinking Water Sources: HNV = APE x Wh x RSC WC + (FC X BAF) 0.5 uq/kq/d x 70 kg x 0.8___________________ 0.01 1/d + (0.015 kg/d x 1,913,875 1/kg*) = 0.00098 ug/1 (rounded off to 1.0 ng/1 (Tier 1)) Where: RSC = 0.8; the substance is persistent and bioaccumulative *BAF = 1,913,875, provided by EPA-Duluth and Minnesota PCA. References: Fitzhugh, O. 1948. Use of DDT insecticides on food products. Industrial and Engineering Chemistry. 40(4):704-705. Laug, E., A. Nelson, 0. Fitzhugh and F. Kunze. 1950. Liver cell alteration and DDT storage in the fat of the rat induced by dietary levels of 1-50 p p m DDT. J. Pharmacol. Exp. Therap. 98:268-273. U.S. Environmental Protection Agency (EPA). 1985. Integrated Risk Information System (IRIS). Chemical file for DDT (50-29-3). Verification Date 12/18/85. Last Revised 9/30/87. 37 Tier 1 Human Cancer Criterion A review of the available literature for DDT carcinogenicity reveals a lack of adequate epidemiological data and an extensive database of chronic oral rodent bioassays. These studies indicate that the induction of liver tumors is the most consistent and significant tumorigenic response to DDT in rodents. EPA (1987) has classified the weight of evidence of DDT carcinogenicity as B2 based on multiple positive studies in two species (mice and rats), with ancillary evidence including promoting activity, genotoxicity, and structural relation to other rodent liver carcinogens. Therefore, the data are sufficient for Tier 1 HCC derivation. The animal bioassay providing the highest slope factor estimation is the multigeneration mouse feeding study of Tarjan and Kemeny (1969). The predominant tumor types were leukemias and lung tumors; a significant liver response was not seen. EPA (1980) derived ambient water quality criteria from the slope factor of 8.422 (mg/kg/day) * from this study. EPA (1986a) evaluated the carcinogenicity of DDT and other related compounds and determined that the Tarjan and Kemeny (1969) study was not the most appropriate basis for quantitative risk assessment. The study's findings were not 38 consistent with the numerous other positive bioassays in terms of the organ site (lung/leukemia versus liver) and the slope factor (about an order of magnitude greater). This slope factor was judged to be a statistical outlier in relation to the liver tumor induction data from six key studies, and the quality and validity of the study was also questionable. EPA (1986a) derived a slope factor from the consistent finding of liver tumor induction in rats and mice, for which the six key studies provided slope factors within a 13-fold range. The recommended slope factor of 3.4 E-l (mg/kg/day) ^ was derived as the geometric mean of ten slope factors from those six studies (Turusov et al., 1973; Terracini et al., 1973; Thorpe and Walker, 1973; Tomatis and Turusov, 1975; Cabral et al., 1982; Rossi et a l . , 1977). The averaging procedure was followed because no further database refinement or rejection could be logically made, and the geometric average of the values was viewed as the best rational estimate of the slope factor (EPA, 1 9 8 6 a ) . The EPA's CRAVE workgroup has reviewed and accepted this approach to slope factor estimation as a method to include all relevant data (EPA, 1987). This averaging approach to slope factor estimation utilizing multiple studies, species, strains and sexes has not g e n e r a l l y been recommended in earlier EPA guide l i n e s (EPA, 1980; 1 9 8 6 b ) . However, more recently, EPA (1989) has 39 stated: "Occasionally, in situations where no single study is judged most appropriate, yet several studies collectively support the estimate, the geometric mean of estimates from all studies may be adopted as the slope. This practice insures the inclusion of all relevant data" (EPA, 1989). In the specific case of DDT, the averaging process as applied to the best available studies may be the most reasonable means of quantitatively characterizing the carcinogenicity of DDT (Schoeny, 1991; Holder, 1991; Bayard, 1991). The Tier 1 Human Cancer Criteria for DDT are derived from the slope factor of 3.4 E-l (mg/kg/d) 1 based on rodent liver tumor induction in the six key studies. RAD = 1 x 1Cr^ _____________ = 2.94 X 10 5 mg/kg/d 3.4 x 10 ^ (mg/kg/d) ^ - 29.4 ng/kg/d Drinking Water Sources: HCV = RAD x Uh_______ = 29.4 nq/ko/d x 70 ko_________________ VC + (FC x BAF) 2 l/d + (0.015 kg/d x 1,913,875 l/kg*) 0.0716 ng/l (rotnded off to 0.07 ng/l (Tier 1>) 40 Nondrinking Water Sources: HCV = RAD x Uh_____ = 29.4 ng/kq/d x 70 kg_________________ WC + (FC x BAF) 0.01 l/d + (0.015 kg/d x 1,913,875 l/ltg*) = 0.0716 ng/l (rounded off to 0.07 ng/t (Tier 1)) Where: *BAF = 1,913,875, provided by EPA-Duluth and Minnesota PCA. References: Bayard, S. 1991. T o x i c o l o g i s t/ S t a t i s t i c i a n w i t h the U.S. EPA Office of Research and Development, Human Health A s s e s s m e n t Group. Personal c o m m u n i c a t i on w i t h R. Sills, Michigan Department of Natural Resources. Cabral, J. et al. 1982. Effects of long-term intake of DDT on rats. Tumori 68:11-17. Holder, J. 1991. Toxicologist with the U.S. E P A Office of Research and Development, Human Health Assessment Group. Personal c o m munication w i t h R. Sills, Mi c h i g a n Department of Natural Resources. Rossi, L. et al. 1977. Long-term a d m i n i s t r a t io n of DDT or phenobarbital-Na in W i star rats. Int. J. Cancer. 19:179-185. Schoeny, R. 1991. U.S. EPA Environmental Criteria Assessment Office, Chair of the Cancer Risk Assessment Verification Endeavor (CRAVE) workgroup. Personal communication with R. Sills, Mi c h i g a n D e p a r t m e n t of Natural Resources. Tarjan, R. and T. Kemeny. 1969. M u l t i g e n e r a ti o n studies on DDT in mice. Food Cosmet. Toxicol. 7:215-222. Terracini, B. et al. 1973. The effects of long-term feeding of DDT to BALB/c mice. Int. J. Cancer. 11:747-764. 41 Thorpe, E. and A. Walker. 1973. The t oxicology of dieldrin. II. Comparative long-term oral tox i c i t y ' studies in mice with dieldrin, DDT, phnobarbital, beta-BHC and gamma-BHC. Food Cosmet. Toxicol. 11:433 442. Tomatis, L. and V. Turusov. 1975. Studies on the carcinogenicity of DDT. Gann Monograph on Cancer Research. 17:219-241. Turusov, V. et al. 1973. Tumors in CF-1 mice exposed for six consecutive generations to DDT. J. Natl. Cancer Inst. 51:983-998. U.S. Environmental Protection Agency (EPA). 1980. 45 Federal Register No. 231, pp. 79347-79356. Appendix - Guidelines and Methodology Used in the Preparation the Consent Decree Water Criteria Documents. c of U.S. Environmental Protection Agency (EPA). 1986a. Assessment of the Carcinogenicity of Dicofol (Kelthane), DDT, DDE, and DDD (TDE). OHEA/ORD. EPA/600/6-86/001. PB 87-110904. The U.S. Environmental Protection Agency (EPA). 1986b. 51 Federal Register No. 185, pp. 33992-34003. Guidelines for Carcinogen Risk Assessment. U.S. Environmental Protection Agency (EPA). 1987. Intergrated Risk Information System (IRIS database). Chemical file for DDT (59-29-3). Verification Date 6/24/87. Last Revised 5/1/91. U.S. Environmental Protection Agency (EPA). 1989. Risk A s s e s s m e n t Guidance for Superfund. V o l u m e 1. H u m a n Health Evaluation Manual (Part A). Interim Final. OERR. EPA/540/1-89/002. 42 October 23, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR DIELDRIN CAS NO. 60-57-1 Tier 1 TTiiman N o n c ancer Criterion A review of the available literature indicates that the most appropriate study for HNC derivation for dieldrin is a two year study conducted by Walker et al. (1969). In this study, 25 Carworth Farm "E" rats of each sex were administered 0.1, 1.0 or 10.0 ppm dieldrin in their diet and 45 rats of each sex were used as controls. At the end of two years, the females exposed to 1.0 and 10.0 ppm had increased liver weights and liver-to-body weight ratios. Histopathological examination of these animals found changes in perenchymal cells which included focal proliferation and focal hyperplasia. A NOAEL of 0.1 ppm (estimated to be 0.005 mg/kg/day) was determined from the study. In support of this value a systemic NOEL of 0.005 mg/kg/day was calculated for dogs in the same study. Studies examining the reproductive effects of dieldrin are lacking (EPA, 1987). A review of studies which examine the developmental effects of dieldrin in mice (Chernoff et al., 1975; Dix et al., 1977) and rats (Harr et al., 1970; Chernoff et al., 1975) suggest that exposure levels which 43 may result in adverse developmental effects are higher than the NOAEL determined in the Walker et al. (1969) study. The quality of the Walker et al. (1969) study was deemed sufficient to derive a Tier 1 HNC. This study was also used by EPA (1987) to derive the oral RfD for dieldrin. The HNC was derived from the NOAEL (0.005 mg/kg/day) using an uncertainty factor of 100 to account for intraspecies variability and interspecies extrapolation. ADE = 0.005 ma/kq/d 100 = 0.00005 mg/kg/d = 50 ng/kg/d Where: Uncertainty Factor = 100, composed of: lOx for intraspecies variability lOx for interspecies extrapolation Drinking Water Sources: HNV = ADE x Wh x RSC = 50 n a / k a / d x 7 0 k g x 0 .8 ______________ WC + (FC X BAF) 2 1/d + (0.015 kg/d x 28,171 1/kg*) = 6.6 ng/1 (rounded off to 7 ng/1 (Tier 1)) 44 Nondrinking Water Sources: HNV = APE x Wh x RSC = 50 nq/kq/d x 70 kq x 0.8_____________ WC + (FC x BAF) 0.01 1/d +(0.015 kg/d x 28,171 1/kg*) = 6 . 6 ng/1 (rounded off to 7 ng/1 (Tier 1)) Where: RSC = 0.8; the substance is persistent and bioaccumulative. *BAF = 28,171, provided by EPA-Duluth and Minnesota PCA. REFERENCES: Chernoff, N . , R.J. Kavlock, J.R. Kathrein, J.M. Dunn and J.K. Haseman. 1975. Prenatal effects of dieldrin and photo-dieldrin in mice and rats. Toxicol, Appl. Pharmacol. 31:302-308. Dix, K.M., C.L. Van Der Paus and W. B. McCarthy. 1977. Toxicity studies with dieldrin: teratological studies in mice dosed orally with HEOD. Teratology 16:57-62. Harr, J . R . , R.R. Claeys, J.F. Bone and T.W. McCorcle. 1970. Dieldrin toxicosis: Rat reproduction. Am. J. Vet. Res. 31:181-189. U.S. Environmental Protection Agency (EPA). 1987. Integrated Risk Information System (IRIS database). Chemical file for dieldrin (60-57-1). Verification Date 4/16/87, Last Revised 9/1/90. Walker, A.I.T., D.E. Stevenson, J. Robinson, E. T h o r p e and M. Roberts. 1969. The t oxicology and pha r m a c o d y n am i c s of dieldrin (HEOD): Two year oral exposures of rats and dogs. Toxicol. Appl. Pharmacol. 15:345-373. 45 Tier 1 bii*ti Cancer Criterion According to EPA (1987a), there are inadequate data available to ascertain whether dieldrin is a human carcinogen. However, chronic studies have shown that dieldrin induces the formation of liver tumors in seven strains of mice when administered orally. Additional support for d i e l d r i n 7s carci n o g e n i ci t y is p r o v i d e d by its structural similarit y to other compounds (i.e. h e p t a c h l o r and chlordane) which have been found to induce tumors in rodents. Dieldrin has also produced a positive response in several mutagenicity studies. The weight-of-evidence for dieldrin carcinogenicity is sufficient for B2 (probable human carcinogen) classification (EPA, 1 9 8 7 a ) . The data are sufficient to derive a Tier 1 HCC. Six key studies (Davis, 1965 as reevaluated by Reuber and cited in Epstein, 1975; Walker et al., 1972; Thorpe and Walker, 1973; NCI, 1978; Tennekes et al., 1981; Meierhenry et al., 1983) have reported liver tumor induction in mice exposed orally to dieldrin. EPA (1987a; 1987b) calculated 13 d i f f e r e n t slope factors u sing data from these studies. The calculated slope factors were within an eight-fold range. EPA (1987a; 1987b) calculated a single oral slope factor of 1.6 x 101 (mg/kg/day)-1 by taking the geometric 46 mean of the 13 slope factors co m p u t e d from t he k e y studies. This method of computing a slope factor is used "in situations where no single study is judged most appropriate, yet several studies collectively support the estimate..." (EPA, 1989). Accordin g to EPA (1989), the a d v antage of this method of determining the slope factor is that all relevant data are used in the computation. RAD = l x 10~5_______________ = 6 x 10" 7 m g / k g / d = 0.6 ng/kg/d 1.6 x 101 (mg/kg/d)"1 Drinking Water Sources: H C V = RAD x Wh________ = 0.6 na/kq/d x 70 kg WC + (FC x BAF) 2 1/d + (0.015 kg/d x 28,171 1/kg*) = 0.099 ng/1 (rounded off to 0.1 ng/1 (Tier 1)) Nondrinking Water Sources: H C V = R A D x Wh________ = 0.6 nq/ka/d x 70 ka____________________ WC + (FC X BAF) 0.01 1/d +(0.015 kg/d X 28,171 1/kg*) = 0.099 ng/1 (rounded off to o.l ng/1 (Tier 1)) *BAF = 28,171, provided by EPA-Duluth and Minnesota PCA. 47 REFERENCES: Davis, K.J. 1965. Pathology report on mice fed aldrin, dieldrin, heptachlor or heptachlor epoxide for two years. Internal FDA m e m o r a n d u m to Dr. A. J. Lehman. July 19. As cited in: Epstein, 1975; EPA, 1987a. Epstein, S.S., 1975. The carcinogenicity of dieldrin. Part 1. Sei. Total Environ. 4:1-52. Meierhenry, E.F., B.H. Reuber, M.E. Gershwin, L.S. Hsieh and S.W.French. 1983. Dieldrin-induced mallory bodies in hepatic tumors of mice of different strains. Hepatology. 3:90-95. National Cancer Institute. 1978. Bioassays of aldrin and dieldrin for possible carcinogenicity. DHEW P ublication No. (NIH) 78-822. National Cancer Institute Carcinogenesis Technical Report Series, No. 22. N C I - C G - T R - 2 2 . Tennekes, H . A . , A.S. Wright, K.M. Dix and J.H. Koeman, 1981. Effects of dieldrin, diet and bedding on enzyme function and tumor incidence in livers of male CF-1 mice. Cancer Res. 41:3615-3620. Thorpe, E. and A.I.T. Walker. 1973. The t oxicology of d i e l d r i n ( H E O D ) . Part II. Comparative long-term oral toxicology studies in mice with dieldrin, DDT, phenobarbitone, beta-BHC and gamma-BHC. Food Cosmet. Toxicol. 11:433-441. U.S. Environmental Protection Agency (EPA). 1987a. Integrated Risk Information System (IRIS database). Chemical file for chlordane (57-74-9). Verification Date 3/5/87. Last Revised 1/1/91. U.S. Environmental Protection Agency (EPA). 1987b. Carcinogenicity Assessment of Aldrin and Dieldrin. Prepared by Carcinogen Assessment Group, Office of Health and Environmental Assessment, Washington, DC for Hazard Evaluation Division, Office of Pesticide Programs, Office of Pesticides and Toxic Substances. OHEA-C-205. U.S. Environmental Protection Agency (EPA). 1989. Risk A s s e s s m e n t Guidance for Superfund. V o l u m e 1. Human Health Evaluation Manual (Part A). Interim Final. OERR. EPA/540/1-89/002. 48 Walker, A.I.T., E. Thorpe and D.E. Stevenson. 1972. The toxicology of dieldrin (HEOD). I. Lo n g - t e r m oral toxicity studies in mice. Food Cosmet. Toxicol. 11:415-432. 49 50 Oct o b e r 23, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR 2,4-DIMETHYLPHENOL CAS NO. 105-67-9 Tier l Human Noncancer Criterion A review of the available literature indicates that HNV derivation for 2,4-dimethylphenol (2,4-DMP) is most appropriately based on the subchronic oral mouse study conducted by EPA (1989). Groups consisting of 30 male and 30 female albino mice were administered 2,4-DMP by gavage at dose levels of 0, 5, 50 or 250 m g / k g/day for 90 days. At day 30, an interim sacrifice was p e r f o r m e d on at least 8 males and 9 females from each group. Effects examined included mortality, clinical signs, body weights, food consumption, ophthalmology, hematology, clinical chemistry, organ weights, and gross histopathology. Toxicologically relevant clinical signs observed only after week 6 at 250 mg/kg/day in both sexes included squinting, lethargy, prostration, and ataxia, with onset shortly after dosing. Statistically significant l o w e r mean corpuscular volume and mean corpuscular hemoglobin concentrations were observed in female mice at 250 rag/kg/day during the final but not during the interim sacrifice. At interim sacrifice, the blood urea 51 Preceding Page Blank nitrogen (BUN) levels for females at 50 and 250 mg / k g / d a y were significantly lower than the vehicle controls, while at the final sacrifice, the BUN levels for females at 50 mg/kg/day were significantly higher than the vehicle control group. For only the low-dose (5 mg/kg/day) m a l e s at the interim sacrifice, cholesterol levels were significantly higher than the vehicle control group. Increased adrenal weights w e r e observed in low-dose (5 mg/kg/day) but not midto high-dose females when compared to vehicle control animals. Since the reported changes in BUN, serum cholesterol and adrenal weights were not dose- or timedependent, they may be interpreted to be spurious findings. The NOAEL and LOAEL for this study were 50 and 250 mg/kg/day, respectively, based on clinical signs and hematological changes. The database is judged to be sufficient for Tier 1 HNC deriva t i o n because the key study (EPA, 1989) pro v i d e s a subchronic NOAEL. However, there is a paucity of supplemental and supportive data. No useful chronic, reproductive or developmental studies are available. The overall findings from the 90-day study (EPA, 1989) compare favorably with the results of a 14-day mice gavage study (EPA, 1987; as cited in EPA, 1989; EPA, 1990) conducted at the same laboratory. In the 14-day study, the only toxicological signs observed in males and females 52 administered 250 mg/kg/day were lethargy, prostration, and ataxia. This is the same dose at which critical effects were found in the 90-day study (EPA, 1989). The HNV is derived from the NOAEL dose of 50 mg/kg/day from the 90-day gavage mouse study by EPA (1989) with an uncertainty factor of 3000. This approach is consistent with the derivation of the oral RfD for 2,4-DMP by EPA (1990). ADE = 50 mq/ka/d = 0.017 mg/kg/day 3,000 Where: Uncertainty Factor = 3,000, composed of: lOx for intraspecies variability lOx for interspecies extrapolation lOx for subchronic to chronic extrapolation 3x for substantial gaps in the database Drinking Water Sources: H N V = ADE x Wh x RSC = 0.017 mq/ k a / d x 70 ka________ WC + (FC x BAF) 2 1/d + (0.015 kg/d x 156 1/kg*) = 0.274 mg/1 (rounded off to 0.3 mg/1 (Tier 1)) 53 Nondrinking Water Sources: HNV = APE x Wh x RSC = 0,017 mg/kg/d x TO kg_________ UC + (FC x BAF) 0.01 L/d + (0.015 kg/d x 156 l/kg*) = 0.51 mg/l (rounded off to 0.5 mg/l (Tier 1)) Where: *BAF = 156, provided by EPA-Duluth and Minnesota PCA. NOTE: A R e l ative Source Contribution (RSCJ factor has not been utilized in these draft calculations. 54 REFERENCES : U.S. Environmental Protection Agency (EPA). 1990. Integrated Risk Information System (IRIS database). Chemical file for 2,4-dimethylphenol (105-67-9). Verification Date 2/21/90. Last Reviewed 2/21/90. U.S. Environmental Protection Agency (EPA). 1989. Ninety- Day Gavage Study in Albino Mice Using 2,4Dimethylphenol. Study No. 410-2831, prepared by Dynamac Corporation, Rockville, MD, for the Office of Solid Waste and Emergency Response, Washington, DC. U.S. Environmental Protection Agency (EPA). 1987. Fourteen-Day Gavage Study in Albino Mice Using 2,4- Dimethylphenol. Study No. 410-2830, prepared by Dynamac Corporation, Rockville, MD, for the Office of Solid Waste and Emergency Response, Washington, DC. As cited in EPA (1989, 1990). U.S. Environmental Protection Agency (EPA). 1980. Ambient Water Quality Criteria for 2,4-Dimethylphenol. Office of Water Regulations and Standards, Criteria and Standards Division, Washington, DC. EPA 440/5-80-044. PB81-117558. 55 56 Octo b e r 23, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR 2,4-DINITR0PHEN0L CAS NO. 51-28-5 Tier 1 TTunian Noncancer Criterion A review of the available literature on the toxic effects and therapeutic use of 2,4-dinitrophenol (2,4-DNP) indicates that the HNC derivation is most appropriately based upon the human dose-response following exposure to 2,4-DNP as reviewed by Horner (1942). Numerous studies on 2,4-DNP and its toxic effects on humans are available (Horner, 1942; SRC, 1981). Commonly-reported toxic effects included gastrointestinal disturbances (nausea, vomiting, loss of appetite), cutaneous rashes, neuritis, agranulocytosis of the bone marrow, and jaundice. Liver and kidney and cardiovascular damage was rarely reported. Evidence of cardiovascular effects was limited to abnormal electrocardiograms indicating functional abnormalities of the heart, although fragmentation of the heart muscle was reported in cases of fatal poisoning. Nine cases of mortality resulting from 2,4-DNP poisoning were cited. Death usually occurred within 24 hours after the onset of such toxic manifestations as dizziness, fatigue, dyspnea, high temperature, intense thirst, and excessive perspiration. 57 Preceding Page Blank In the study by Horner (1942), bilateral cataract formation was frequently observed in patients receiving 2,4-DNP as a weight-loss agent. The study reported that cataracts developed in more than 164 persons after the use of dinitrophenol, an estimated incidence of 0.86 percent. The study did not include a control group, however the researcher noted that this type of cataract is not expected to occur in some of the age groups which exhibited cataracts in the study. Formation of cataracts occurred either during dosing or within several months to a year after the final dose was taken. Cataracts were observed in patients receiving as little as 2 mg/kg bw/day which was the lower range of the recommended therapeutic dose for obesity. This LOAEL determined from the Horner (1942) study was deemed sufficient for the derivation of a Tier 1 HNC. In a 6-month feeding study, male rats (from the Breeding and Laboratory Institute, Brooklyn, NY) were administered 2,4DNP at d i e t a r y levels of 0, 100, 200, 500 and 1000 p p m for 178-179 days (Spencer et a l . , 1948). There w e r e 14, 12, 12, 9 and 14 rats per dietary level, respectively. An additional 10 rats were fed 2000 p pm but after 24 days this group experienced 40% mortality and the remaining animals at 2000 ppm were sacrificed and examined at this time. These animals were emaciated and had empty gastrointestinal tracts, enlarged spleens with hemosiderosis, testicular 58 atrophy, and increased levels of blood urea nitrogen. Rats fed 1000 ppm 2,4-DNP suffered a reduction in body weight gain of 10-15%, a slight depletion of body fat, a very slight increase in the average weight of the kidneys, and a very slight decrease in the weight of the heart. Blood urea nitrogen levels were elevated in 2/14 animals at 1000 ppm. Reduced growth occurred at 500 ppm and a significant increase (between 91% and 92% above controls) in kidney weights occurred at all dietary concentrations. The authors concluded that the male rats maintained for six months on diets containing 200 ppm (and presumably 100 ppm) showed no appreciable ill effects. However, because there was a statistically significant increase in kidney weights at all dietary concentrations, the dose of 100 ppm may be considered the LOAEL for this study. Using a food consumption value of 0.08 k g / k g b w (EPA, 1988), the LOAEL for the Spencer et al. (1948) study was 8 mg/kg bw/day. This is very close to the LOAEL of 2.0 mg/kg bw/day which was calculated using the human data from Horner (1942). EPA (1980) derived an Acceptable Daily Intake (ADI) from an estimated NOAEL of 5.4 mg/kg/day (100 ppm group) from the study by Spencer et al. (1948). In a teratology study with 2,4-DNP, Gibson (1973) reported that neither intraperitoneal (7.7 and 13.6 mg/kg/day) nor oral (25.5 and 38.2 mg/kg/day) doses of 2,4-DNP administered 59 to pregnant Swiss-Webster mice during early organogenesis (days 10-12 of gestation) produced morphological defects. However, the higher intraperitoneal dose was embryotoxic and the higher intraperitoneal and oral doses produced overt signs of toxicity (hyperexcitability and hyperthermia) in the dams. The HNV is derived from the LOAEL (2.0 mg/kg bw/day) determined from the human data summarized by Horner (1942) using an uncertainty factor of 1000. This approach is consistent with the derivation of the oral RfD for 2,4-DNP by EPA (1986). ADE = 2 mg/kg/d = 0.002 mg/kg/d = 2 ug/kg/d 1000 Where: Uncertainty Factor = 1000, composed of : lOx for subchronic to chronic conversion lOx for intraspecies variability lOx for LOAEL to NOAEL conversion Drinking Water Sources: H N V = A D E x W h x R S C = 2 ua/k a / d x 70 kg__________________ WC + (FC X BAF) 2 1/d + (0.015 kg/d X 4.B 1/kg*) 60 = 67.6 ug/1 (rounded off to 0.07 mg/1 (Tier 1)) Nondrinking Water Sources: HNV = A P E x Wh x RSC = 2 uq/kq/d X 70 kq_____________________ WC + (FC x BAF) 0.01 1/d + (0.015 kg/d x 4.8 1/kg*) = 1707 ug/1 (rounded off to 1.7 mg/1 (Tier 1)) Where: PCA. *BAF = 4.8, provided by EPA-Duluth and Minnesota NOTE: A Relative Source Contribution (RSC) factor has not been utilized in these draft calculations. REFERENCES: Gibson, J.E. 1973. Teratology studies in mice with 2secbutyl-4, 6-dinitrophenol (dinoseb). Food Cosmet. Toxicol.11:31-43. Horner, W.D. 1942. Dinitrophenol and its relation to formation of cataracts. Arch. Ophthal. 27:1097-1121. Spencer, H.C., V.K. Rowe, E.M, Adams and D.D. Irish. 1948. Toxicological studies on laboratory animals of certain alkyl dinitrophenols used in agriculture. J. Indus. Hyg. Toxicol. 30:10-25. Syracuse Research Corporation (SRC), Center for Chemical Hazard Assessment. 1981. Information Profiles on Potential Occupational Hazards: Nitrophenols. Prepared for National Institute for Occupational Safety and Health (NIOSH), Rockville, MD. PB89-215842/XAD. PHS-NIOSH-- 210-79-0030. 61 U-S. Environmental Protection Agency (EPA). 1988. Recommendations For And Documentation Of Biological Values For Use In Risk Assessment. PB88-179874. U.S. Environmental Protection Agency (EPA). 1986. Integrated Risk Information System (IRIS database). Chemical file for 2,4-dinitrophenol (51-28-5). Verification Date 2/5/86. Last Reviewed 2/5/86, U.S. Environmental Protection Agency (EPA). 1980. Ambient Water Quality Criteria Document for Nitrophenols. Prepared by the Office of Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH for the Office of Water Regulations and Standards, Criteria and Standards Division, Washington, DC. EPA 440/5-80-063. 62 Octo b e r 31, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR HEPTACHLOR CAS NO. 76-44-8 Tier l Human Noncancer criterion A review of the available literature indicates that the most appropriate study for HNC derivation for heptachlor is a two-year study conducted by Velsicol Chemical Corporation (1955) as cited by EPA (1987). In this study, CF strain rats (20/sex/group) w e r e fed diets c o n t a i n i n g 0, 1.5, 3, 5, 7, or 10 ppm heptachlor. The N O E L for m a l e rats w a s 3 ppm (0.15 mg/kg/d) with the liver-to-body weight ratio reported as the sensitive endpoint. The LEL for this critical effect was 5 ppm (0.25 mg/kg/d). Two studies suggest that exposure levels which may cause adverse effects on reproduction are higher than the NOEL determined in the study cited above. EPA (1987) cited a NOEL of 0.25 mg/kg/d in a one-generation rat reproduction test (Velsicol, 1955) and a NOEL of 0.5 mg/kg/d in a threegeneration rat reproduction test (Velsicol, 1967). However, Green (1970) found adverse effects on reproduction in rats which were administered heptachlor in feed at 5 ppm 63 (approximately 0.25 mg/kg/d). This value was considered a LOAEL and no NOAEL was established in this study. With respect to developmental toxicity, Yamaguchi et al. (1987) found no teratogenic effects in the offspring of female rats dosed o rally w i t h 5, 10 or 20 m g / k g heptac h l o r from days 7 17 of gestation. The quality of the Velsicol (1955) study and supporting data was deemed sufficient to derive a Tier l HNC. This study was also used by EPA (1987) to derive the oral RfD for heptachlor. ADE = 0.15 mq/kq/d = 0.0015 mg/kg/d = 1.5 ug/kg/d 100 Where: Uncertainty Factor = 100, composed of: lOx for intraspecies variability lOx for interspecies extrapolation Drinking Water Sources: H N V = A D E X w h X RSC = 1.5 ua/ k a / d x 70 ka x 0.8____________ WC + (FC x BAF) 2 1/d + (0.015 kg/d x 19,097 1/kg*) = 0.29 ug/1 (rounded off to 0.3 ug/1 (Tier 1)) 64 Nondrinking Water Sources: HNV = APE x Mh x RSC = 1.5 uq/k<3/d x 70 kg x 0.8___________ UC + (FC * BAF) 0.01 t/d + <0.015 Icg/d x 19,097 l/kg*) = 0.29 ug/l (rounded off to 0.3 ug/l (Tier 1)) Where: RSC = 0.8; the substance is persistent and bioaccumulative. *BAF = 19,097, provided by EPA-Duluth and Minnesota PCA. REFERENCES: Green, V. 1970. Effects of Pesticides on Rat and Chick Embryo. In: Hemphill, D. (Ed.). 1970. Trace Substances in Environmental Health-Ill. Proceedings of University of Missouri's 3rd Annual Conference on Trace Substances in Environmental Health, June 24-26, 1969. U.S. Environmental Protection Agency (EPA). 1987. Integrated Risk Information System (IRIS database). Chemical file for heptachlor (76-44-8). Verification Date 4/16/87. Last Revised 3/1/91. Velsicol Chemical Corporation. 1955. MRID No. 0062599. Available from EPA. Write to FOI, EPA, Washington, DC 20460. Velsicol Chemical Corporation. 1967. MRID No. 00147058. Available from EPA. Write to FOI, EPA, Washington, DC 20460. Yamaguchi, M . , S. Tanaka, K. Kawashima, S. N a k a u r a and A. Takanaka. 1987. Effects of heptachlor on fetal development of rats. Natl. Inst. Hyg. Sci. 105:33-36. 65 Tier 1 Human Cancar Criterion There are inadequate data available to ascertain whether heptachlor is a human carcinogen. However, heptachlor exposure has caused a significant increase in hepatocellular carcinomas in two strains of mice. In addition, heptachlor is structurally related to other compounds (i.e., dieldrin and chlordane) which have been found to induce the formation of liver tumors in mice. Heptachlor has produced negative results in gene mutation assays, mouse dominant lethal assays and in vitro D N A repair assays (EPA, 1980; 1986; 1987). According to the weight of evidence approach employed by EPA (1986; 1987; 1990), there is sufficient evidence to classify heptachlor as a B2 carcinogen (probable human carcinogen). There is also sufficient evidence to derive a Tier 1 HCC. Two key studies (Davis, 1965 as cited by Epstein, 1976; NCI, 1977) have reported hepatocellular tumors in mice exposed orally to heptachlor. EPA (1986; 1987; 1990) calculated four slope factors using data from these studies. The slope factors ranged from 0.83 to 14.9 (mg/kg/d)- 1 . EPA (1987) calculated a single slope factor of 4.5 (mg/kg/day) 1 by taking the geometric mean of the four slope factors from the key studies. This method of computing a slope factor is used "in situations where no single study is judged most appropriate, yet several studies collectively support the 66 estimate ..." (EPA, 1989). The advantage of this m e t h o d of determining the slope factor is that all relevant data are used in the computations (EPA, 1989). R A D = 1 x 10~5________ = 0.0000022 m g / k g / d = 2 . 2 ng/k g / d 4.5 (mg/kg/d)-1 Drinking Water Sources: H C V = RAD x Wh________ = 2.2 na/ka/d x 70 ka__________________ WC + (FC X BAF) 2 1/d + (0.015 kg/d X 19,097 1/kg*) = 0.534 ng/1 (rounded off to 0.5 ng/1 (Tier 1)) Nondrinking Water Sources: HCV = RAD x Uh_______ = 2 . 2 no/ka/d x 70 kg_________________ UC + (FC * BAF) 0.01 l/d + (0.015 kg/d x 19,097 t/kg*) = 0.538 ng/l {roinjed off to 0.5 ng/l (Tier 1)) *BAF = 19,097, provided by EPA-Duluth and Minnesota PCA. REFERENCES: Davis, K. 1965. P a t h o l o g y Report on Mice Fed Aldrin, Dieldrin, Heptachlor and Heptachlor Epoxide for Two Years. Internal FDA m e m o r a n d u m to Dr. A. J. Lehman, J u l y 19. Epstein, S.S. 1976. Carcinogenicity of heptachlor and chlordane. Sei. Total Environ. 6:103-154. 67 National Cancer Institute (NCI). 1977. Bioassay of Heptachlor for Possible Carcinogenicity. NCI Carcinogenesis Tech. Rep. Ser. No. 9. U.S. Environmental Protection Agency (EPA). 1980. Ambient Water Quality Criteria for Heptachlor. Criteria and Standards Office. Washington, DC. E PA 440/5-80-052. U.S. Environmental Protection Agency (EPA). 1986. Carcinogenicity Assessment of Chlordane and Heptachlor/Heptachlor Epoxide. Carcinogen Assessment Group. Office of Health and Environmental Assessment, Washington, DC. U.S. Environmental Protection Agency (EPA). 1987. Integrated Risk Information System (IRIS database). Chemical file for heptachlor (76-44-8). Verification Date 4/1/87. Last Revised 1/1/91. U.S. Environmental Protection Agency (EPA). 1989. Risk A s s e s s m e n t Guida n c e for Superfund. V o lume 1. Human Health Evaluation Manual (Part A ) . Interim Final. OERR. EPA/540/1-89/002. U.S. Environmental Protection Agency (EPA). 1990. Drinking Water Criteria Document for Heptachlor, Heptachlor Epoxide and Chlordane. Revised November, 1990. ECAOCIN-406. 68 October 31, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR HEXACHLOROBENZENE CAS NO, 118-74-1 Tier 1 yppcancer Criteria A r eview of the available information on h e x a c h l o r o b e n z e n e (HCB) toxicity, including reviews by EPA (1980; 1985a; 1985b; 1988), indicates that the database is sufficient for Tier l HNC derivation. The best available data consist of laboratory animal studies. The principle human data on HCB toxicity consist of widespread toxic effects among several thousand Turkish citizens exposed to HCB via consumption of fungicide-treated grain during 1955-1959. The resulting effects included porphyria cutanea tarda (PCT), neurotoxicity, liver damage, and increased infant mortality. The exposure has been estimated at 50-200 mg/day over an extended period, without further description of the dosage estimation method (Cam and Nigogosyan, 1963). The human data cannot be used for quantitative risk assessment because accurate exposure data are not available (EPA, 1988). The available literature indicates that HCB is a potent developmental toxicant in several animal species. K i t c h i n et al. (1982) exposed rats to 0, 60, 80, 100, 120 and 140 p p m H CB in feed (doses were approximately 4.5-10 mg/kg/day). They reported 69 a dose-dependent increase in mortality of pups in the Fla and Flb litters. Grant et al. (1977) conducted a 4-generation r e production study w i t h rats at food HCB levels in the diet of 0, 10, 20, 40, 80, 160, 320 and 640 ppm. They concluded that 20 ppm (about 1.5 mg/kg/day) was a NOAEL, while 40 ppm (about 3 mg/kg/day) resulted in increased liver weights and aniline hydroxylase activities in weanlings. Rush et al. (1983) exposed m i n k to 0, 1 or 5 p pm in feed (about 0.16 or 0.78 mg/kg/day), resulting in profound effects on kit survivability to weaning at the high dose. Mortality was 8.2 percent, 4.1 percent and 77.4 percent among controls, low dose, and high dose groups, respectively. Bleavins et al. (1984a, 1984b) also reported that mink, as well as ferrets, are highly sensitive to the developmental effects of HCB. Mink were found to be more sensitive than ferrets, while both appeared more sensitive than rats according to published data. The most profound effects reported were decreased mink birth weights at adult dietary levels as low as 1 ppm and a dose-related increase in kit mortality at three weeks of age among both mink and ferrets at levels as low as 1 ppm (about 0.14 and 0.11 mg/kg/day for mink and ferrets, respectively). Additionally, effects were seen on the levels of hypothalamic dopamine of mink kits and on hypothalamic serotonin in adult mink. These changes were statistically significant at levels as low as 1 ppm in feed. A NOAEL was not reported. ; 70 Arnold et al. (1985) exposed male and female rats to dietary HCB levels of 0, 0.32, 1.6, 8.0 or 40 p pm for 90 days p r i o r to mating and until 21 days after parturition (at w e a n i n g ) . The offspring were exposed in utero, from maternal nursing, and from their diets for the remainder of their lifetime. The total study period was 130 weeks. A NOAEL was reported at 1.6 ppm (about 0.08 m g / k g / d a y ) . At 8 p p m (0.29 mg/kg/day) the p a rental (Fo) males demonstrated increased heart and liver weights and the FI generation had an increased incidence of hepatic centrilobular basophilic chromogenesis. The 40 ppm FI groups showed increases in pup mortality, hepatic centrilobular basophilic chromogenesis, and severe chronic nephritis (males only). The effects of HCB on adult animals has been further demonstrated in many other studies, a few of which report NOAELs. KuiperGoodman et al. (1977) exposed rats via the diet to 0, 0.5, 2, 8 and 32 m g / k g bw/day for up to 15 weeks. A N O A E L w as rep o r t e d at 0.5 mg/kg/day, while at the higher dose levels, increased tissue porphyrin, increased organ weights and increased severity of centrilobular liver lesions were noted. Grant et al. (1974) exposed rats to HCB at diet a r y levels of 10, 20, 40, 80 a nd 160 ppm for 9-10 months. Porphyria was induced at levels as low as 20 p p m (about 1 m g / k g / d a y ) . The data selected for HNC determination are from the Arnold et al. (1985) study. This study involved the e x posure of adult 71 Sprague-Dawley rats and subsequent exposure of the offspring in utero, via lactation, and via the diet. Cross-fostering studies have demonstrated that the neonate is particularly sensitive to the toxic effects of HCB. The transfer of HCB to neonates via the milk of exposed adults has also been shown to be significant (Bailey et al. 1980; Bleavins et al. 1982). The A r n o l d et al. (1985) study demonstrated NOAELs of 0.32 and 1.6 ppm in feed (estimated to be 0.016 and 0.08 mg/kg/day). Therefore, the NOAEL of 0.08 mg/kg/day, with an uncertainty factor of 100 (lOx for each inter- and intraspecies extrapolation) is used for HNC derivation. Consideration was also given to the mink and ferret data (Rush et al., 1983; Bleavins et al., 1984a, 1984b) which demonstrate that these species are highly sensitive to the developmental toxic effects of HCB. Adverse effects on the development of mink and ferrets have been reported at doses only slightly higher than the rat NOAEL of 0.08 mg/kg/day. However, the rat NOAEL is utilized preferentially because the SpragueDawley rat, unlike the mink or ferret, is a human-surrogate species which has been extensively studied as an animal model for toxicity testing, and because the Arnold et al. (1985) study is of very good quality. The selection of this key study is consistent with the development of the RfD for HCB by EPA (1988). ADE = 0.08 ma/ka/d = 0.0008 mg/kg/d = 0.8 ug/kg/d 100 72 Where: Uncertainty factor = 100, composed of: lOx for intraspecies variability lOx for interspecies extrapolation Drinking Water Sources: H NV = A P E X Wh X RS C = 0.8 u a / k q / d X 70 k a X 0.8_____________ WC + (FC X BAF) 2 1/d + (0.015 kg/d X 208,590 1/kg*) = 0.014 ug/1 (rounded off to 10 ng/1 (Tier 1)) Nondrinking Water Sources: HNV = APE x W h x RSC = 0.8 ua/ka/d x 70 ka x 0.8________________ WC + (FC x BAF) 0.01 1/d + (0.015 kg/d x 208,590 1/kg*) = 0.014 ug/1 (rounded off to 10 ng/1 (Tier 1)) Where: RSC = 0.8; the substance is persistent and bioaccumulative. *BAF = 208,590, provided by EPA-Duluth and Minnesota PCA. 73 REFERENCES: Arnold, D.L. et al. 1985. L o n g - t e r m toxicity of hexachlorobenzene in the rat and the effect of dietary v i tamin A. Fd, Chem. Toxic. 23(9):779-793. Bailey, J , , V. Knauf, W. Mueller and W. Hobson. 1980. T r ansfer of hexachlorobenzene and polychlorinated biphenyls to nursing infant rhesus monkeys: Enhanced toxicity. Environ. Res. 21(1): 190-196. Bleavins, M.R., W.J. Breslin, R.J. Aulerich and R.K. Ringer. 1982. Excretion and placental and mammary transfer of hexachlorobenzene in the European ferret (Mustela putorius f u r o ) . J. Toxicol. Environ. Health. 10:929-940. Bleavins, M . , R. Auler i c h and R. Ringer. 1984a. Effects of chronic dietary hexachlorobenzene exposure on the reproductive performance and survivability of mink and European ferrets. Arch. Environ. Contam. Toxicol. 13:357-365. Bleavins, M. et al. 1984b. Effects of dietary hexach l o r o b en z e n e exposure on regional brain biogenic amine concentrations in mink and European ferrets. Toxicol. Environ. Hlth. 14:363 377. Cam, C. and G. Nigogosyan. 1963. A c q u i r e d toxic po r p h y r i a cutaneatarda due to hexachlorobenzene. Report of 348 cases caused by this fungicide. J. Am. Med. Assoc. 183:88-91. Grant, D. et al. 1974. Effects of he x a c h l o r o b en z e n e on liver porphyrin levels and microsomal enzymes in the rat. Environ. Physiol. Biochem. 4:159-165. Grant, D . , W. Phillips and G. Hatina. 1977. Effect of hexachlorobenzene on reproduction in the rat. Arch. Environ. Contam. Toxicol. 5(2):207-216. Kitchin, K, et al. 1982. Of f s p r i n g m o r t a l i t y and m a t e r n a l lung pathology in female rats fed hexachlorobenzene. Toxicol. 23:33-39. Kuiper-Goodman, T. et al. 1977. subacute t oxicity of hexachlorobenzene in the rat. Toxicol, and Appl. Pharmacol. 40:529-549. Rush, G. et al. 1983. Perinatal hexach l o r o b en z e n e tox i c i t y in the mink. Environ. Res. 31:116-124. 74 U.S. Environmental Protection Agency (EPA). 1980. Ambient Water Quality Criteria for Chlorinated Benzenes. EPA 440/5-80028. U.S. Environmental Protection Agency (EPA). 1985a. Drinking Water Criteria Document for Hexachlorobenzene (Final Draft). EPA - 600/X-84-179-1. PB-86-117777. U.S. Environmental Protection Agency (EPA). 1985b. Health Assessment Document for Chlorinated Benzenes. EPA/600/884/015F. U.S. Environmental Protection Agency (EPA). 1988. Integrated Risk Information System (IRIS database). Chemical file for hexachlorobenzene (118-74-1). Verification Date 5/26/88. Last Revised 4/1/91. Tier l Human cancer criterion A review of the available literature indicates that there are inadequate epidemiological studies and sufficient animal carcinogenicity data, supporting a B2 weight-of-evidence classification (EPA, 1989). The animal bioassays, which have been comprehensively reviewed and summarized by EPA (1980; 1985a; 1985b; 1989), indicate that HCB induces tumors of the liver predominantly, with neoplasm induction of the thyroid and kidney also reported. The data are judged sufficient for Tier 1 HCC derivation. EPA (1991) derived an oral slope factor of 1.6 per (mg/kg)/day from a chronic rat bioassay demonstrating hepatocellular carcinoma induction (Erturk et al., 1986). This slope factor is among the h i ghest of those derived for H CB from 14 di f f e r e n t datasets, which fell within a range of 8.3 E-2 to 1.7 E+0 (EPA, 75 1989). This dataset was also selected for slope factor estimation because the study was well-conducted and the tumors were malignancies of the primary target organ (liver cancers). In the key study, Erturk et al. (1986; abstracts previously published as Lambrecht et al., 1983a; 1983b) exposed groups of 94 S p r a g u e-Dawley rats/sex/do se to HCB via feed at 0, 75 or 150 ppm in the diet for up to two years. Treated animals of both sexes surviving past 12 months showed significant increases in liver and renal tumors. Females were far more susceptible to hepatocarcinogenicity while males were generally more sensitive to renal carcinogenicity. The slope factor of 1.6 per (mg/kg)/day is derived from the induction of hepatocellular carcinomas in female rats. This is consistent with EPA (1989). RAD = 1 x 10~5________ = 6.2 x 10-6 m g / k g / d = 6.2 ng/kg/d 1.6 (mg/kg/d)"* Drinking Water Sources: H C V = RAD x Wh________ = 6.2 nq/ko/d x 70 kg__________________ WC + (FC x BAF) 2 1/d + (0.015 kg/d x 208,590 1/kg*) = 0.14 ng/1 (rounded off to 0.1 ng/1 (Tier 1)) 76 Nondrinking Water Sources: HCV = RAD x Wh________ = 6.2 ng/ka/d x 70 kg_______________________ WC + (FC x BAF) 0.01 1/d + (0.015 kg/d x 208,590 1/kg*) = 0.14 ng/1 (rounded off to 0.1 ng/1 (Tier 1)) Where: *BAF = 208,590, provided by EPA-Duluth and Minnesota PCA. REFERENCES: Erturk, E. et a l . 1986. O n c o genicity of h e x a c h l o r o b e n z e n e . In: Hexachlorobenzene: Proc. int. S y m p . , C.R. Morris and J.R.P. Cabral, Eds. IARC Scientific Publ. No. 77, Oxford University Press, Oxford. pp. 417-423. Lambrecht, R . , et al. 1983a. Renal tumors in rats chronically exposed to hexachlorobenzene (HCB). Proceedings of the American Association for Cancer Research 24:59. Lambrecht, R . , et al. 1983b. Hepatocarcinogenicity of chronically administered hexachlorobenzene in rats. Federa t i o n Proceedings. 4 2 ( 4 ) :786. U.S. Environmental Protection Agency (EPA). 1980. Ambient Water Quality Criteria for Chlorinated Benzenes. EPA 440/5-80028. U.S. Environmental Protection Agency (EPA). 1985a. Drinking Water Criteria Document for Hexachlorobenzene (Final Draft). EPA-600/X-84-179-1. NTIS: PB 86-117777. U.S. Environmental Protection Agency (EPA). 1985b. Health Assessment Document for Chlorinated Benzenes. EPA/600/884/015F. U.S. Environmental Protection Agency (EPA). 1989. Integrated Risk Information System (IRIS database). Chemical file for hexachlorobenzene (118-74-1). Verification Date 3/1/89. Last Revised 3/1/91. 77 78 October 31, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR HEXACHLOROETHANE CAS NO. 67-72-1 Tier 1 iTiiinn Noncancer Criteria A review of the available literature indicates that the most appropriate basis for HNV derivation for hexachloroethane (HCE) is the NOAEL from a 16-week dietary study in rats (Gorzinski et al., 1985; Gorzinski et al., 1980, as cited in EPA, 1991). In this study male and female CDF Fischer 344 rats (10/sex/group) were administered a diet containing HCE at t arget levels of 0, 3, 30 and 100 m g / k g / d a y for 16 weeks. EPA (1991) reported that actual dose levels were analyzed to be a p p r o ximately 0, 1.3, 20 and 82 mg/kg/day. From analysis of eating patterns and measurement of the time-related loss of HCE from the diets, a conservative estimate of exposure was determined by the investigators as 0, 1, 15 and 62 mg/kg/day (Gorzinski et al., 1985). The results indicate that male rats were slightly more sensitive than female rats to the nephrotoxic properties of HCE. Renal toxicity observed at 15 and 62 mg / k g / d a y in male rats i ncluded pale and mottled kidneys; significant increases in absolute and relative kidney weights; slight to moderate renal tubular -- -------- ------- ^ Preceding Page Blank | J 79 atrophy and degeneration with or without peritubular fibrosis; a slight to moderate increase in renal tubular cytoplasmic clumping and droplet formation; and scattered or isolated renal tubules with slight hypertrophy and/or dilation of the proximal convoluted tubules. Liver weights were increased in mal e rats given 62 mg/kg/day. The liver exhibited a slight swelling of the hepatocytes in males given 15 or 62 mg/kg/day. Evidence of renal toxicity in female rats consisted of very slight renal tubular atrophy and degeneration observed histopathologically at the highest dose level. Female rats given 62 mg/kg/day also had an increase in relative liver weight ratios unaccompanied by microscopic alterations. Based on this study, a NOAEL of 1 mg/kg/day was derived for liver and kidney toxicity in male rats. While EPA (1991) indicates a NOAEL of 1.3 mg/kg/day based on the analyzed low dose, the estimated NOAEL of l.o mg/kg/day (Gorzinski et al., 1985; EPA, 1987) is used in the HNV derivation. In a chronic (78-week) gavage study with rats and mice, the National Cancer Institute (NCI, 1978) administered HCE in a cyclic manner to 50 male and 50 female Osborne-Mendel rats and continuously to 50 male and 50 female B6C3F1 mice. The rats received HCE in corn oil at doses of 250 and 500 mg/kg/day, 5 days per week for a period of 22 consecutive weeks, followed by a 1-week, treatment-free interval. 80 Thereafter, until the end of the 78 weeks, the rats were intubated for 4 consecutive weeks followed by 1 treatment free week, in a cyclical pattern, for a total of 66 weeks of HCE treatment. The time-weighted-average doses for the rats for the 78-week period were 212 and 423 mg/kg/day. The mice were intubated orally with HCE in corn oil at initial levels of 500 and 1000 mg/kg/day for 8 weeks with these doses increased to 600 and 1200 mg/kg/day, respectively, for the remaining 70 experimental weeks. A time-weighted-average dose of 590 and 1179 mg/kg/day for the low and high doses, respectively, was reported. The dosing regimes were followed by an observation period of 33 or 34 weeks for rats and 12 or 13 weeks for mice. Renal tubular nephr o p a t h y was observed during histopathological examination at the termination of the study in all groups of treated animals. In rats, significant pathology and mortality at both dose levels in the males precluded the development of a NOAEL or LOAEL for HCE. For the mice, due to the occurrence of hepatocellular carcinoma and non-neoplastic toxic nephropathy in both sexes at both dose levels, neither a NOAEL nor a LOAEL could be determined. Because of the inconclusive nature of results from the NCI (1978) study, additional toxicological and carcinogenesis studies were conducted by administering HCE in corn oil by gavage to groups of male and female F344/N rats 81 (50/sex/group) 5 days per w e e k for 2 years (NTP, 1989). The male rats received doses of 0, 10 or 20 m g / k g / d a y w h i l e the females received doses of 0, 80 or 160 mg/kg/day. The foremost toxic effect was kidney toxicity, demonstrated by increased incidences of mineralization and hyperplasia of the pelvic transitional epithelium in dosed male rats, increased severity of renal tubule hyperplasia in high dosed male rats, and increased incidence and severity of renal tubule hyperplasia in female rats. The LOAEL was 10 mg/kg bw/day for male rats. In this study, it was hypothesized that the increased sensitivity of male rats to the renal t o x icity of HCE was a result of the accumulation of <*2u " globulin in hyaline droplets synthesized by the liver and secreted into the blood (EPA, 1991). It is then apparently filtered through the glomeruli and partially reabsorbed through the proximal tubules. In the presence of HCE, as well as several nonpolar hydrocarbons such as decalin and gasoline, a 2u"globulin accumulates in hyaline droplets in the renal tubular cells. a 2u-Globulin is an excretory protein in male but not female rats. This may explain the male's greater sensitivity to kidney damage from HCE. In a 13-week rat study, also by NTP (1989), groups of 10 F 344/N rats of each sex were administered 0, 47, 94, 188, 375 or 750 mg/kg HCE in corn oil by gavage, 5 days/week for 13 weeks. Five/10 male rats and 2/10 female rats at 750 82 mg/kg/day died before the end of the study. The final mean body weight of male rats that received 750 mg/kg/day was 19% lower than that of vehicle controls. Compound-related clinical signs for both sexes included hyperactivity at doses of > 94 mg/kg/day and convulsions at 375 mg/kg/day. The relative weights of liver, heart and kidney were increased for exposed males and females. Kidney lesions were seen in all dosed male groups, and the severity increased with dose. Papillary necrosis and tubular cell necrosis and degeneration in the kidney and hemorrhagic necrosis in the urinary bladder were observed in the five male rats at 750 mg/kg/day which died before the end of the study. At all lower doses in males, hyaline droplets, tubular regeneration, and granular casts were present in the kidney. No chemical-r e la t e d k i dney lesions were obs e r v e d in females. Foci of hepatocellular necrosis were observed in several male and female rats at > 188 mg/kg/day. Weeks et al. (1979) studied the effects of repeated exposure to H C E vapor in 25 male and 25 female rats, 4 m a l e dogs, 10 male guinea pigs, 20 ma l e or female quail and 22 pre g n a n t rats per exposure group. The animals were exposed for 6 hours/day, 5 days/week for 6 weeks and doses were analyzed at 0, 15, 48 or 260 ppm of the HCE v apor (equivalent to 0, 145, 465 or 2515 m g / m 3 ; EPA, 1991). Toxic effects at the 83 h i ghest concentrations included tremors and. other n eurotoxic signs. No effects were observed at < 465 mg/m3 . Weeks et al. (1979) performed an oral study, in which HCE doses of 100, 320 and 1000 mg/kg/day were administered by g avage to rabbits for 12 days. The two highest doses resulted in liver degeneration and necrosis, toxic tubular nephrosis of the convoluted tubules of the corticomedullary region of the kidney, minimal tubular nephrocalcinosis, and decreased body weights. The NOAEL for this study was 100 mg/kg/day based on the effects of HCE on the kidneys of male rabbits. The database is judged to be sufficient for Tier 1 HNC derivation. The key study (Gorzinski et a l . , 1985) provides a subchronic NOAEL which is supported and supplemented by chronic toxicity data (EPA 1989; EPA, 1991; NCI, 1978; NTP, 1989; Weeks et a l . , 1979). The HNC is based on the subchronic rat NOAEL of 1 mg/kg/day (0.71 mg/kg/day converted for 5 day administration/week), with a total uncertainty factor of 1000. The LOAEL of 15 mg/kg/day resulted in male rat renal toxicity. It may be argued that the high sensitivity of this endpoint is peculiar to male rats, secondary to hyaline droplet formation and a2uglobulin accumulation. However, liver affects also occurred at a LOEL of 15 mg/kg/day. The use of the NOAEL of 1 84 mg/kg/day for risk assessment is consistent with the oral RfD development by EPA (1987) and the Lifetime Health A d v i s o r y (EPA, 1991). ADE = Q .71 mo/ka/d = 0.00071 mg/kg/d = 0.71 ug/kd/d 1,000 where: Uncertainty Factor = 1,000, composed of: lOx for intraspecies variability lOx for interspecies extrapolation lOx for subchronic exposure duration Drinking Water Sources: HNV = ADE x Wh x RSC = 0.71 ua/ko/d x 70 ka x 0.8 WC + (FC X BAF) 2 1/d + (0.015 kg/d X 950 1/kg*) = 2.45 ug/1 (rounded off to 2 ug/1 (Tier 1)) Nondrinking Water Sources: HNV * ADE x Wh x RSC = 0.71 uo/kg/d x 70 kg x 0.8_________ WC + (FC * BAF) 0.01 l/d + (0.015 kg/d * 950 L/kg*> = 2.79 ug/l (rounded off to 3 ug/l (Tier 1)) Where: RSC = 0.8; HCE is bioaccumulative. 85 Where: *BAF = 950 1/kg, provided by EPA-Duluth and Minnesota PCA. REFERENCES: Gorzinski, S.J., R.J. Nolan, S.B. McCollister, D.C. Morden, E.A. Hermann, D.A. Dittenbar, R.V. Kainis, J.E. Battjes and R.J. Kociba. 1980. Hexachloroethane: Results of a 16-Week Toxicity Study in the Diet of CDF Fischer 344 Rats. Toxicology Research Laboratory, Dow Chemical U.S . A . , Midland, MI. As cited in EPA (1991). Gorzinski, S . J . , R.J. Nolan, S.B. McCollister, R.J, Kociba and J.L. Mattsson. 1985. Subchronic oral toxicity, tissue distribution and clearance of hexachloroethane in the rat. Drug and Chem. Toxicol. 8(3):155-169. National Cancer Institute (NCI). 1978. Bioassay of Hexachloroethane for Possible Carcinogenicity. NCI C a r c i nogenesis Technical Report Series No. 68, NCI- C G T R - 6 8 , DHEW Publication No. (NIH) 78-1318. National Toxicology Program (NTP). 1989. Toxicology and Carcinogenesis Studies of Hexachloroethane (CAS No. 67 72-1) in F344/N Rats (Gavage Studies). NTP Technical Report. N T P - T R - 3 6 1 , NIH/PUB-89-2816, Order No. PB90- 170895, 117 pp. U.S. Environmental Protection Agency (EPA). 1991. Hexachloroethane. Health Advisory. Office of Drinking Water, Washington, DC. PB91-159657/XAD. U.S. Environmental Protection Agency (EPA). 1989. Health and Environmental Effects Document for Hexachloroethane. Environmental Criteria and Assessment Office, Cincinnati, OH. EPA/600/8-88/043. PB88-178736/GAR. ECAO-CIN-G041. U.S. Environmental Protection Agency (EPA). 1987. Integrated Risk Information System (IRIS database). Chemical file for hexachloroethane (67-72-1). Verification Date 4/16/87. Last Reviewed 4/16/91. Weeks, M.H., R.A. Angerhofer, R. Bishop, J. T h o m a s i n o and C.R. Pope. 1979. The toxicity of hexachloroethane in laboratory animals. Amer. Ind. Hyg. Assoc. J. 40(3) :187-199. 86 Tier 1 Human Cancer Criterion A review of the available literature for HCE carcinogenicityreveals a lack of adequate epidemiological data and two chronic oral rodent bioassays (NCI, 1978; NTP, 1989). EPA (1986) has classified HCE as a class C carcinogen (possible human carcinogen), based on the observation of carcinomas in one mouse strain after oral exposure (NCI, 1978). The data are judged to be sufficient for Tier l HCC derivation. In a NCI study (NCI, 1978), Osborne-Mendel rats and B6C3F1 mice were orally intubated with HCE in corn oil. Groups of 50 rats per sex per dose were administered HCE over a 78week period with an exposure protocol involving intermittent treatment-free intervals. The time-weighted-average doses were 212 or 423 mg/kg/day. The rats were then observed for an additional 33-34 weeks. Groups of 50 mice per sex per dose were administered HCE 5 days per week for 78 weeks at time-weighted-average doses of 590 or 1179 mg/kg/day, and were then observed for an additional 12-13 weeks. Due to an unusually high mortality rate among the male control mice, the results in treated groups were compared against both the vehicle control group from this study as well as a pooled vehicle control group from several concurrent studies. A statistically significant increase in the incidence of hepatocellular carcinoma was reported in both sexes of the mice (only males exhibited a dose-related trend) while 87 tumorigenicity was not observed in rats of either sex. The increased incidence was significant by the Cochran-Armitage test for both sexes of mice against both control groups and by the Fisher exact tests for both sexes as compared to the pooled controls. Survival of low- and high-dose male and female rats in this study was reduced compared with that of the vehicle controls. Because findings from NCI (1978) in rats were inconclusive, additional studies on toxicity and carcinogenesis were conducted in F344/N rats by administering HCE in corn oil by gavage to groups of males and females for 2 years (NTP, 1989). HCE was administered 5 days/week in c o m oil by g a v a g e at 0, 10 or 20 m g / k g bw to groups of 50 m a l e rats, and at 0, 80 or 160 m g HCE/kg b w to groups of 50 female rats. The incidences of renal adenomas and carcinomas alone and in combination increased in the high dose male group. One of the carcinomas in the high dose group metastasized to the lung. No compound-related neoplasms were observed in females. The incidence of pheochromocytomas of the adrenal gland in low dose male rats was significantly greater than that in vehicle controls, and the incidences for both dosed groups were greater than the mean historical control incidence rates. The renal lesions were considered by NTP to be indicative of HCE carcinogenicity while the pheochromocytomas were judged to be supportive evidence for 88 carcinogenic effects. On the basis of these data, NTP concluded that there was clear evidence of carcinogenicity for HCE in the male rat and no evidence of carcinogenicity in female rats. Renal tubule hyperplasia was observed at an increased incidence in high dose male rats. These lesions have been described as characteristic of the hyaline droplet nephropathy that is associated with an accumulation of l i ver-generated a 2il-globulin in the c y t o p l a s m of tubular epithelial cells (NTP, 1989). Using this assumption, it can be hypothesized that the male rat renal tumors were a secondary effect to hyaline droplet formation and that they may not be relevant to human risk assessment. The Tier 1 Human Cancer Criteria for HCE are derived from the slope factor of 1.4 E-2 (mg/kg/d)-1 based on a doseresponse data-set for hepatocellular carcinoma induction in male mice from the NCI study (NCI, 1978; EPA, 1986). RAD = 1 x 10~5________________ = 7.14 x 10-4 m g / k g / d 1.4 x 10-2 (mg/kg/d)-1 Drinking Water Sources: H CV = RAD x Wh________ = 7.14 X 10~4 m a / k g / d x 70 ka WC + (FC x BAF) 2 1/d + (0.015 kg/d x 950 1/kg*) 89 = 3.1 x 10" 3 mg/1 (rounded off to 3 ug/1 (Tier 1)) Nondrinking Water Sources: HCV = rad x wh______ = 7.14 x 10'* nw/kg/d x 70 tea________ UC + CFC x BAF) 0.01 L/d + (0.015 kg/d x 950 l/kg*) = 3.5 x 10's mg/t (rounded off to 4 yg/l (Tier 1)) Where: *BAF = 950 l/kgf provided by EPA-Duluth and Minnesota PCA. 90 REFERENCES; National Cancer Institute (NCI). 1978. Bioassay of Hexachloroethane for Possible Carcinogenicity. NCI Carcinogenesis Technical Report Series No. 68, NCI-CGTR-68, DHEW Publication No. (NIH) 78-1318. National Toxicology Program (NTP). 1989. Toxicology and Carcinogenesis Studies of Hexachloroethane (CAS No. 67 72-1) in F344/N Rats (Gavage Studies). NTP Technical Report. NTP-TR-361, NIH/PUB-89-2816, Order No. PB90170895, 117 pp. U.S. Environmental Protection Agency (EPA). 1986. Integrated Risk Information System (IRIS database). Chemical file for hexachloroethane (67-72-1). Verification Date 7/23/86. Last Reviewed 7/23/86. 91 92 Oct o b e r 31, 1991 GREAT LAKES INITIATIVE HUMAN HEALTH CRITERIA FOR LINDANE (G A M M A - H E X A C H L O R O C Y C L O H E X A N E ) CAS NO. 58-89-9 Tier 1 Human Noncancer Criterion A review of the available literature indicates that the most appropriate study for the derivation of the HNV for lindane is a subchronic study conducted by Zoecon Corporation (1983) as evaluated by EPA (1991) and summarized by EPA (1986). In this study, Wistar KFM-Ham (outbred) SPF rats (20/sex/dose) were administered 0, 0.2, 0.8, 4, 20 or 100 p p m lindane in the feed. Fifteen animals/sex/group were sacrificed after 12 weeks. The remaining rats were fed the control diet for an additional six weeks before sacrifice. Rats exposed to 20 and 100 ppm lindane had a greater incidence of liver hypertrophy, kidney tubular degeneration, hyaline droplets, tubular distension, interstitial nephritis and basophilic tubules than did the controls. The NOAEL for this study was 4 ppm. This dose was estimated to be equivalent to 0.29 mg/kg/d for the male and 0.33 mg/kg/d for the female rats. Two chronic studies which examined the effects of lindane on rats and dogs were cited by EPA (1986). A two-year study by Fitzhugh (1950) reported a NOAEL of 2.5 mg/kg/d in Wistar rats with liver weights and liver damage evaluated as the Preceding page W a n k 93 endpoints. In a two-year study in beagle dogs, Rivett et al. (1978) reported a NOAEL of 1.6 mg/kg/d for liver toxicity. A review of the database on developmental and reproductive effects of lindane suggests that these effects may occur at levels higher than the NOAEL calculated in the study conducted by Zoecon Corporation (1983). Palmer et al. (1978a) found no adverse effects on reproductive function and development following exposure of female rats to lindane in the feed at levels of 1.25, 2.5 and 5 mg/kg/d for three generations. Khera et al. (1979) found no reproductive effects in wistar rats exposed to lindane at levels ranging from 6.25 to 25 mg/kg from the 6th to the 15th day of gestation. No adverse effects were found in a teratogenicity study on pregnant rabbits fed lindane on g e s t ation days 6-18 at levels of 5, 10 and 15 m g / k g (Palmer et al., 1978b). However, Sircar and Lahiri (1989) reported that even the lowest exposure group (3.75 mg/kg/d) of Swiss mice receiving lindane during gestation experienced reproductive failure. The quality of the study conducted by Zoecon Corporation (1983) was deemed sufficient to derive a Tier 1 HNC. The results of studies which examine the reproductive or developmental effects of lindane are either negative or 94 indicative of possible effects at doses substantially higher than the NOAEL reported by Zoecon Corporation (1983). A l t h o u g h subchronic in duration (12 w e e k s ) , the key study is supported by chronic studies in the database. This study was also used by EPA (1986) to derive the oral RfD for lindane. The HNC was derived from the female rat NOAEL (0.33 mg/kg/d) using an uncertainty factor of 1000 to account for intraspecies variability, interspecies extrapolation and the extrapolation from a subchronic to chronic study. The magnitude of this unce r t a i n t y factor is expected to result in adequate protection from any potential reproductive or developmental effects as well as chronic noncancer effects. ADE = 0.33 ma/ka/d = 0.00033 mg/kg/d = 0.33 ug/kg/d 1000 Where: Uncertainty Factor = 1000, composed of: lOx for subchronic to chronic extrapolation lOx for intraspecies variability lOx for interspecies extrapolation Drinking Water Sources: HNV = A D E x Wh x RSC = 0.33 uq/ k q / d x 70 kq x 0.8_________ WC + (FC x BAF) 2 1/d + (0.015 kg/d X 1,628 1/kg*) 95 = 0.69 ug/1 (rounded off to 0.7 ug/1 (Tier l)) Nondrinking Water Sources: H N V = A P E x W h X RSC = 0.33 uc/kq/d it 70 ka x 0.8____________ WC + (FC x BAF) 0.01 1/d +(0.015 kg/d x 1,628 1/kg*) = 0.76 ug/1 (rounded off to 0.8 ug/1 (Tier 1)) Where: RSC * 0.8; the substance is biaccumulative. *BAF = 1,628, provided by EPA-Duluth and Minnesota PCA. REFERENCES: Fitzhugh, O.G., A.A. N e l s o n and J. P. Frawley. 1950. The chronic toxicities of technical benzene hexachloride and its alpha, beta and gamma isomers. J. Pharm. Exp. Ther. 100:59-66. Khera, K.S., C. Whalen, G. Trivett and G. Angers. 1979. Teratogenicity studies on pesticidal formulations of dimethoate, diuron and lindane in rats. Bull. Environ. Contain. Toxicol. 22 (4-5): 522-529. Palmer, A.K., D.D. Cozens, E.J.F. Spicer and A.N. Worden. 1978a. Effects of lindane upon reproductive function in a 3-generation study in rats. 10(l):45-54. Palmer, A . K . , A.M. Bottomley, A.N. Worden, H. Fr o h b e r g and A. Bauer. 1978b. Effect of lindane on p r e g n a n c y in the rabbit and rat. Toxicol. 9(3):239-247. Rivett, K.F., H. Chesterman, D.N. Kellett, A.J. N e w m a n and A.N. Worden. 1978. Effects of feeding lindane to dogs for periods of up to two years. Toxicol. 9:273-289. 96 sircar, s. and P. Lahiri. 1989. Lindane (gamma-HCH) causes reproductive failure and fetotoxicity in mice. Toxicol. 59:171-177. U.S. Environmental Protection Agency (EPA). 1986. Integrated Risk Information System (IRIS database). Chemical file for lindane (58-89-9). Verification Date 1/22/86. Last Revised 3/1/88. U.S. Environmental Protection Agency (EPA). 1991. Data Evaluation Record (DER) for lindane. Office ofPesticide Programs. Zoecon Corporation. 1983. Unpublished report. MRID No. 00128356. Available from EPA. Write to FOI, EPA, Washington, D.C. 20460. Tier 2 Human Cancer level of protection There are inadequate data available to ascertain whether lindane is a human carcinogen (IARC, 1982; EPA, 1985; ATSDR, 1989). The preponderance of evidence indicates that lindane is carcinogenic to mice (EPA, 1985). Animal bioassays conducted by Thorpe and Walker (1973), NCI (1977), Goto et al. (1972) and Hanada et al. (1973) provide evidence that lindane induces liver tumors. Possible relevance to humans is indicated by the occurrence of a carcinogenic metabolite (2,4,6-trichlorophenol) in humans and other species following exposure to lindane (EPA, 1980; 1985; ATSDR, 1989). The weight-of-evidence for lindane carcinogenicity is report e d l y sufficient for "B2-C" c l a s s i f i c a t io n (EPA, 1985; 1991). The database was not judged sufficient for B2 (probable human carcinogen) classification because of limitations in the quality of the bioassay data and also 97 because the results of most mutagenicity tests have been negative (EPA, 1985). According to EPA (1985), "the weightof-evidence appears to be closer to a Category C carcinogen than to Category B2 c arcinogen". The cancer risk assessment for lindane is u n d e r revi e w by EPA (EPA, 1990). For this initiative, the data are sufficient to derive a Tier 2- level of protection because liver tumor induction has been found in more than one bioassay, in both sexes and in multiple strains of mice. A Tier 1 criterion is not indicated, due to the limited quality of the bioassay data, the lack of mutagenicity data, and the pending assessment by EPA. The Thorpe and Walker (1973) study was used to determine th Tier 2 level of protection because it provides the best quality data for quantitative risk assessment. In this study, 30 CF1 mice of each sex were exposed to 400 ppm lindane in their diet for up to 110 weeks. The pooled control group consisted of 45 animals of each sex. Besides having only one exposure level, the quality of the study was compromised because a low percentage of treated mice survived (3% of females and 17% of males) to the end of the study (EPA 1980; 1985). There was a significant increase in the incidence of hepatic neoplasms in treated male and female mice. Liver neoplasms were found in 27/28 (96%) t r e a t e d male m i c e and 20/21 (95%) treated female m i c e as c o m p a r e d to 11/45 (24%) and 10/44 (23%) liver neoplasms in 98 male and female controls, respectively. The male mouse data were used to calculate a slope factor of 1.3 (mg/kg/d) 1 via the linearized multistage model Global 82. RAD = l x 10~5_________ = 7.6 x 10-6 mg/kg/d = 7.6 ng/kg/d 1.3 (mg/kg/d)-1 Drinking Water Sources: H CV = R AD X Wh________ = 7.6 nq/kq/d x 70 ka__________________ W C + (FC x BAF) 2 1/d + (0.015 kg/d x 1,628 1/kg*) = 20 ng/1 (rounded off to 20 ng/1 (Tier 2)) Nondrinking Water Sources: HC V = RA D x W h ________ = 7.6 nq/kq/d X 70 R q ________ WC + (FC X BAF) 0.01 1/d +(0.015 kg/d x 1,628 1/kg*) = 21.7 ng/1 (rounded off to 20 ng/1 (Tier 2)) *BAF = 1,628, provided by EPA-Duluth and Minnesota PCA. REFERENCES: 99 Agency for Toxic Substances and Disease Registry (ATSDR). 1989. Toxicological Profile for Alpha-, Beta-, Gammaand Delta-Hexachlorocyclohexane. U.S. Public Health Service. Goto, M . , M. Hattori and T. Mizagawa. 1972. Contr i b u t i o ns to ecology. II. Hepatoma development in mi c e after administration of HCH isomers in high dosages. Chemosphere. 1:279-282. Hanada, M . , E. Kawano, S. K awamura and M. Shiro. 1981. Radiation and photo-induced degradation of five isomers of 1,2,3,4,5,6-hexachloro-cyclohexane. Agric. Biol. Chem. 45(3):659-665. International Agency for Research on Cancer (IARC). 1982. IARC Monographs on the Evaluation of the Carcinogenic Risk of chemicals to Humans. Suppl. 4:133-135. National Cancer Institute (NCI). 1977. Bioassay of Lindane for Possible Carcinogenicity. NCI Carcinogenesis Tech. Rep. ser. No. 14. 99 p. NTIS PB-273-480. Thorpe, E. and A.I. Walker. 1973. The toxic o l o g y of dieldrin (HEOD). II. Comparative long-term and toxicity studies in mice with dieldrin, DDT, phenobarbitone, Betz-BHC and gamma-BHC, Food Cosmet. Toxicol. 11:433 442. U.S. Environmental Protection Agency (EPA). 1980. Ambient Water Quality Criteria for Hexachlorocyclohexane. Criteria and Standards Office. Washington, DC. EPA 444/5-80-054. U.S. Environmental Protection Agency (EPA). 1985. Drinking Water Criteria Document for Lindane. Prepared by the Office of Health and Environmental Assessment. Environmental Criteria and Assessment Office, Cincinnati, OH for the Office of Drinking Water, W a s h i n g t o n , D .C . U.S. Environmental Protection Agency (EPA). 1990. Integrated Risk Information System (IRIS database). Chemical file for lindane (58-89-9). Last Revised 8/1/90. U.S. Environmental Protection Agency (EPA). 1991. Health Effects Assessment Summary Tables. 100 October 23, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR MERCURY CAS NO. 7439-97-6 (INCLUDING METHYLMERCURY, CAS NO. 22967-92-6) Tier l Human Noncancer Criterion A review of the available literature on the environmental cycling, fate, and toxicity of mercury and mercury compounds indicates that HNC derivation is most appropriately based upon the human dose-response to methylmercury. Numerous reviews on mercury toxicity (e.g., WHO, 1976; 1990; EPA, 1980; 1984a; 1984b; 1985a) describe the human dose-response relationship resulting from food-borne exposure to methylmercury in Iraq (1971-72), Japan (1940s thru 1960s), and elsewhere. These data are judged to be sufficient for Tier 1 criterion derivation. Studies of widespread human food-borne exposure to m e t h y l m e r c u ry in fish (Minamata and Niigata, Japan) and in seed grain (Iraq) have shown that neurological symptoms of mercury toxicity in adults appear with blood levels of mercury in the range of 200 to 500 ng/ml (Nordberg and Strangert, 1976; Clarkson et al., 1976; WHO, 1976; 1990; 101 EPA, 1980; 1984a; 1984b; 1985a). However, there are a few studies of workers exposed occupationally to mercury via inhalation which suggest that blood mercury levels as low as 10-20 ng/ml may result in the development of signs of renal dysfunction (increased proteinurea and albuminurea) and abnormal psychomotor performance (Roels et al., 1982; Piikivi et al., 1984; Buchet et a l . , 1980). The adult LOAEL of 200 ng/ml in blood has been associated with an intake level of 200-500 u g / d (EPA, 1980; WHO, 1990), a l though the human adult population's variability in mercury elimination rate is significantly bimodal (Clarkson et al, 1976; Nordberg and Strangert, 1976). The human LOAEL of 200 ug/d, or 3 ug/kg/d, for the development of neurological effects forms the basis for the RfD derived by EPA (1985b) and the fish consumption criteria derived by EPA (1980). It has been estimated that less than 5% of the adult population will experi e n c e neurological effects at these levels (WHO, 1990). The risk assessments by EPA (1980) and EPA (1985b) utilized a total uncertainty factor of 10 in conjunction with the LOAEL dose, and both stated that the LOAEL and the risk assessment addressed the sensitivity and the adequate protection of both pre- and postnatal exposures. EPA (1980) justified the 10-fold uncertainty factor as an accounting for "individual differences in habits of fish consumption 102 and in susceptibility to the toxic effects of methylmercury, including prenatal exposures". EPA (1985b) justified the 10-fold u ncertainty factor "to adjust the L O A E L to what is expected to be a NOAEL. Since the effects are seen in sensitive individuals for chronic exposure, no additional factors are deemed necessary". For the derivation of the Tier 1 Human Noncancer Criterion, a total u n c e r t a i n t y factor of 50 will be utilized. This is composed of a 10-fold factor to adjust the adult LOAEL to a presumed adult NOAEL and an additional 5-fold factor to protect CNS development during the sensitive fetal life stages. The use of a 10-fold factor for LOAEL-to-NOAEL conversion is justified by consideration of the severity and irreversibility of the effects at the LOAEL, the long latency of mercury effects, and the occupational studies which suggest that the threshold may be considerably lower than 200 ng Hg/ml blood. An uncertainty factor of 5 is utilized to ensure that the criterion will be protective of the fetal effects of mercury exposure via maternal ingestion of mercury-contaminated fish. The particular sensitivity of the fetus has been recognized in reviews of mercury toxicity (WHO, 1976; 1990; D'ltri, 1978; EPA, 1980; 1984a; 1984b; 1985a). The earliest 103 of these assessments (WHO, 1976) developed a dose-response relationship for the adult which was not presented as being accurate for the more sensitive fetal effects. It was noted that many infant victims reported from Minamata had severe cerebral involvement (palsy and retardation) whereas their mothers had mild or no manifestations of poisoning. Although these observations were qualitatively confirmed by animal studies, quantification of the difference in the degree of sensitivity between human fetuses and adults has been elusive. EPA (1980; 1985b) utilized a total uncertainty factor of 10 and assumed that the resulting risk assessments were adequately protective of fetal effects. However, WHO (1990) reviewed the database on oral methylmercury ingestion, including more recent studies, and made significant advances in delineating quantitatively the greater sensitivity of prenatal exposure relative to adult exposure. Although WHO (1990) did not recommend a particular numeric sensitivity factor for the fetus, their assessment sufficiently demonstrates that an additional uncertainty factor is reasonable and prudent to help ensure adequate protection. They concluded that adult effects occur at a LOAEL (for 5% increased occurrence rate) of 200 ng/ml blood, or at 50 ug/g in hair. Fetal effects on CNS d e v e l o p m e n t occur at a LOAEL (5% increased occurrence rate) of 10-20 ug/g as a peak level in maternal hair. Since the level of mercury in maternal blood correlates to the 104 simultaneous level in new hair growth, the hair serves as a fairly reliable indicator of maternal blood mercury levels during pregnancy. The data suggest that the fetal effects LOAEL may be 2.5 to 5 times lower than the adult effects LOAEL. The HNC is derived from the a d u l t `LOAEL dose of 3 ug/kg/d which is associated with the LOAEL in blood of 200 ng/ml, and an uncert a i n t y factor of 50. The m e t h y l m e r c u r y form is the most significant of the mercury compounds from the standpoint of ambient environmental mercury and human exposures and health impacts. Aqueous concentrations of mercury, and especially methylmercury, may be very low in ambient waters. Other forms of mercury, such as elemental mercury or mercury (I), may be reasonably anticipated to be transformed predominantly to methylmercury in the aquatic environment via oxidation to merc u r y (II) and biomethylation. The biomethylation of inorganic mercury and the very high propensity for methylmercury to bioaccumulate in aquatic organisms result in a high and significant human exposure potential (EPA, 1980; D'ltri, 1990; A n n e t t et al., 1975). The various forms of mercury released to and found in the ambient aquatic environment may be assumed to be converted primarily to methylmercury. Therefore, the HNC is expressed as the total recoverable mercury concentration. 105 ADE = 3 ua/ka/d = 0.06 ug/kg/d 50 Where: Uncertainty Factor = 50, c omposed of: lOx for LOAEL-to-NOAEL conversion 5x for intraspecies variability (protection of fetal CNS development) Drinking Water Sources: H N V = A D E x W h x RSC = 0.06 uq/kq/d x 70 ka x 0.8_________ WC + (FC x BAF) 2 1/d + (0.015 kg/d x 130,440 1/kg*) = 0.0017 ug/1 (rounded off to 2 ng/1 (Tier 1)) Nondrinking Water Sources: HHV - ADE x UH x RSC - 0.06 ug/kg/d x 70 kg x 0.8___________ UC + (FC x BAF) 0.01 l/d + <0.015 kg/d x 130,440 1/kg*) 0.0017 ug/L (ranted off to 2 ng/l (Tier 1)) Where: RSC = 0.8; the substance is persistent and bioaccumulative, *BAF - 130,440, provided by EPA-Duluty and Minnesota PCA. 106 REFERENCES: Annett, C.S. et al. 1975. Mercury in fish and waterfowl .from Ball Lake, Ontario. J. Environ. Qual, 4(2):219- 222 Buchet, J.P., H. Roels, A. Bernard and R. Lauwerys, 1980. Assessment of renal function of workers exposed to inorganic lead, cadmium or m e r c u r y vapor. J. Obcup. Med. 22:741-750. Clarkson, T.W., L. Amin-Zaki and S. K. Al-Tikriti. 1976. An outbreak of methylmercury poisoning due to consumption of contaminated grain. Federation Proceedings. 35(12):2395-2399. D'ltri, P.A. and F.M. D'ltri. 1978. Mercury contamination: a human tragedy. Environmental Management. 2(1):3-16. D'ltri, F.M. 1990. Mercury contamination - what we have learned since Minamata. Environmental Monitoring and Assessment, v. 16. Nordberg, G.F, and P, Strangert. 1976. Estimations of a dose-response curve for long-term exposure to methylmercurie compounds in human beings taking into account variability of critical organ concentration and biological half-time: a preliminary communication. In: Effects and Dose-Response Relationships of Toxic Metals. 1976. Elsevier Scientific Publishing Company. Amsterdam, T h e Netherlands, p. 273-282. Piikivi, L . , H. Hanninien, T. M a r t e l i n et al. 1984. Pyschological performance and long term exposure to m e r c u r y vapors. Scand. J. Work. Environ. H e alth 10:35-41. , Roels, J., R. Lauwerys, J.P. Buchet et al. 1982. Comparison of renal function and psychomoter performance in workers exposed to elemental mercury. Int. Arch. Occup. Environ. Health 50:77-93. U.S. Environmental Protection Agency (EPA). 1980. Ambient Water Quality Criteria Document for Mercury. EPA 440/5-80-058. U.S, Environmental Protection Agency (EPA). 1984a. Mercury Health Effects Update: Health Issue Assessment. OHEA. EPA--600/8-84-019F . 107 U.S. Environmental Protection Agency (EPA). 1984b. Health Effects Assessment for Mercury. EPA/540/1-86/042. NTIS: PB86-134533. U.S. Environmental Protection Agency (EPA). 1985a. Drinking Water Criteria Document for Mercury. Prepared for Office of Drinking Water, by Environmental Criteria and Assessment Office. EPA-600/X-84-178-1. Final Draft. PB86-117827. U.S. Environmental Protection Agency (EPA). 1985b. ' Integrated Risk Information System (IRIS database). Chemical file for methylmercury (22967-97-6). Verification Date 12/2/85. Last Revised 2/1/89. World Health Organization (WHO). 1976. Environmental H ealth Criteria 1: Mercury. WHO, Geneva. World Health Organization (WHO). 1990. Environmental Health Criteria 101: Methylmercury. WHO, Geneva. 108 N ovember 4, 1991 GREAT LAKES INITIATIVE TIER I HUMAN HEALTH CRITERIA FOR METHYLENE CHLORIDE CAS NO. 75-09-2 Tier 1 TTnmww Noncancer Criterion A review of the literature indicates that hepatic and renal toxicities are characteristic critical effects of methylene chloride subchronic and chronic exposure (EPA, 1989). From animal studies on the chronic toxicity of methylene chloride, the most appropriate basis for HNV derivation is the NOAEL from the chronic oral rat study by the National Coffee Associ a t i o n (NCA, 1982; Serota, et al., 1986a). In this study, F344 rats (85/sex/group) received nominal doses of 0, 5, 50, 125 or 250 m g / kg/day of m e t h y l e n e ch l o r i d e via drinking water exposure for 2 years. An induction of liver toxicity in the females was observed. Treatment-related histological alterations such as increases in hepatocellular foci and fatty changes in the liver were observed in rats of both sexes at nominal doses of 50 mg/kg/day. No treatment-related effects were noted in the rats administered the nominal dose of 5 mg/kg/day. The actual NOAEL doses were 5.85 and 6.47 mg/kg/day for males and females, respectively. In addition to the rat drinking water study, Hazleton Labs conducted a 24-month study with B6C3F1 mice for the National 109 Coffee A s s o c iation (NCA, 1983; Serota et al., 1 9 8 6 b ) . In this study, m i c e were e x p o s e d to nominal doses of 0, 60, 125, 185, and 250 mg/kg/day of methylene chloride in drinking water for up to 24 months. Dose-related histomorphologic changes, such as proliferative hepatic lesions and enhanced amount of Oil Red 0 positive material, were observed in groups exposed to the highest dose of methylene chloride. The study reported a NOAEL of 185 mg/kg/day. Compared to the 5.85-6.47 mg/kg/day NOAEL in rats (NCA, 1982; Serota et al., 1986a), this study demonstrates a wide difference in the interspecies sensitivities to methylene chloride-induced toxic effects. In a 2-year inhalation toxicity and oncogenicity study by Dow Chemical Co. (Nitschke et al., 1988), groups of SpragueD a w l e y rats (90 m a l e and 180 female) were e x p o s e d to 0, 50, 200, or 500 ppm methylene chloride for 6 hours/day, 5 days/week for 2 years. During the course of the study, all rats were monitored after each exposure for signs of toxicity, changes in body weight and food intake. Samples of liver tissue were analyzed for DNA synthesis as indicated by 3 [H]-thymidine uptake. Rats selected for interim n e c r o p s i e s and all the others (at the end of the study) were subjected to extensive gross pathologic, histopathologic, and serum chemistry evaluation. Data on DNA synthesis in the liver, pathology, histopathology, mortality, and other no parameters were evaluated for statistically significant differences between the exposed and controls groups. Pathologic and histopathologic data of the exposed groups indicated that the liver and kidney are the primary targets of methylene chloride toxicity. An increased incidence of hepatocellular vacuolization was observed in male and female rats exposed to 500 ppm of methylene chloride. In addition, elevated numbers of multinucleated hepatocytes were observed in female rats exposed to 500 ppm methylene chloride. The effects of methylene chloride at lower doses (50 and 200 ppm) were comparable with historical controls. Responses to chemical insult leveled off by 12 months in that the responses of female rats exposed to 500 pp m for the first 12 months were comparable to those of female rats exposed to the same concentration for 24 months. Based on these results, the authors concluded that 200 p p m (706.7 m g / m 3 ) is the NOAEL for methylene chloride by the inhalation route. This exposure may be converted to a daily administered dose of approximately 159 mg/kg/day, adjusting for 6 hours/day exposure and assuming that Sprague-Dawley rats breathe a p p r oximately 0.9 m 3 /kg bw/day (EPA, 1988). Burek et al. (1984) reported a 2-year inhalation study of methylene chloride with Sprague-Dawley rats and Golden Syrian Hamsters. In this study, rats and hamsters were exposed to 0, 500, 1500, and 3500 ppm of m e t h y l e n e chloride ill for 6 hours/day, 5 days/week for 2 years. Liver and mammary glands were the principal target tissues of methylene chloride inhalation toxicity in rats. Groups exposed to 500 to 3500 ppm methylene chloride showed increased incidence of hepatocellular vacuolization consistent with fatty changes, and the number of multinucleated hepatocytes in the female rats was elevated. After 18 months of exposure to the regimen, several characteristic lesions of liver and mammary glands were transformed to benign neoplasms (Burek et a l., 1984) . The database is judged to be sufficient for Tier 1 HNC derivation. The k e y study (NCA, 1982; Serota et al. 1986a) provides a chronic oral NOAEL which is supported and supplemented by other oral and inhalation chronic toxicity data. EPA used this key study in the derivation of the oral RfD for risk assessment purposes (EPA, 1 9 8 5 a ) , and to derive lifetime h ealth adviso r i e s for m e t h y l e n e chloride (EPA, 1985b). For the RfD derivation, EPA (1985a) used the male and female rat doses, respectively of 52.58 and 58.32 m<3/kg/day for LOAELs, and 5.85 and 6.47 m g / k g/day for NOAELs for hepatic effects (NCA, 1982; Serota et al., 1986a). The HNC is derived from the h n o a e l from the key study, 6.47 mg/kg/day to female rats. 112 ADE = 6.47 ma/kq/d = 0.065 mg/kg/d 100 Where: Uncertainty factor = 100, composed of: lOx for interspecies extrapolation lOx for intraspecies variability Drinking Water Sources: H N V = ADE x Wh x RSC = 0.065 ma/ka/d x 70 ka____________ WC + (FC x BAF) 2 1/d + (0.015 kg/d x 2.5 1/kg*) = 2.23 mg/1 (rounded off to 2 mg/1 (Tier 1)) Nondrinking Water Sources: H N V = A D E X Wh X RSC = 0.065 mq/ k q / d X 70 ka ______________ WC + (FC x BAF) 0.01 1/d + (0.015 kg/d x 2.5 1/kg*) - 95.8 mg/1 (rounded off to 96 mg/1 (Tier 1)) Where: *BAF =2.5, provided by EPA-Duluth and Minnesota PCA. 113 Note: A r e l ative source c o n t r i b u t i o n (RSC) factor has not been utilized in these draft calculations. REFERENCES: Burek, J.D., K.D. Nitschke, T.J. Bell, D.L. Wackerle, R.C. Childs, J.E. Beyer, D.A. Dittenber, L.W. Rampy, and M.J. McKenna. 1984. Methylene chloride: A two-year inhalation toxicity and oncogenicity study in rats and hamsters. Fundam. Appl. Toxicol. 4:30-47. National Coffee Association (NCA). 1983. Twenty-Four Month Oncogenicity Study of Methylene Chloride in Mice. Prepared by Hazleton Laboratories America, Inc., Vienna, VA. (Unpublished). National Coffee Association (NCA). 1982. 24-Month Chronic Toxicity and Oncogenicity Study of Methylene Chloride in Rats. Final Report. Prepared by Hazleton Laboratories America, I n c . , Vienna, VA. (Unpublished). National Toxicology Program (NTP). 1986. Toxicology and Carcinogenesis Studies of Dichloromethane (Methylene Chloride) in F344/N Rats and B6C3F1 Mice (Inhalation Studies). NTP-TRS-306. Nitschke, K . D . , J.D. Burek, T.J. Bell, R.J. Kociba, L.W. Rampy, and M.J. McKenna. 1988. Methylene chloride: A two-year inhalation toxicity and oncogenicity study in rats. Fundam. Appl. Toxicol. 11:48-59. Serota, D . G . , A.K. Thakur, B.M. Ulland, J.c. Kirschman, N.M. Brown, R.H. Coots and K. Morgareidge. 1986a. A twoyear drinking water study of dichloromethane on rodents. I. Rats. Food Chem. Toxicol. 24:951-958. Serota, D.G,, A.K. Thakur, B.M. Ulland, J.C. Kirschman, N.M. Brown, R.H. Coots and K. Morgareidge. 1986b. A twoyear drinking water study of dichloromethane on rodents. II. Mice. Food Chem. Toxicol. 24:959-964. U.S. Environmental Protection Agency (EPA). 1989. Health Effects Assessment for Methylene Chloride. EPA/600-889-092. Environmental Criteria and Assessment Office (ORD), Cincinnati, OH. PB90-142449. 1X4 U.S. Environmental Protection Agency (EPA). 1988. Recommendations for and Documentation of Biological Values for Use in Risk Assessment. PB88-179874. U.S. Environmental Protection Agency (EPA). 1985a. Integrated Risk Information System (IRIS database). Chemical file for methylene chloride (75-09-2). Verification date 11/6/85. Last reviewed 11/6/85. U.S. Environmental Protection Agency (EPA). 1985b. Healtyh Advisory for Dichloromethane. Prepared by the Office of Drinking Water, Washington, D.C. PB86-118338. Tier 1 BumTM Cancer Criterion M e t h y l e n e chloride is a class B2 c arcinogen (a probable human carcinogen) according to the EPA weight-of-evidence classification of carcinogenic chemicals (EPA, 1 9 8 9 a ) . Th classification rationale is based on sufficient evidence from animal carcinogenicity. Two epidemiological studies on chemical factory workers exposed to methylene chloride (Ott et al., 1983; Friedlander et al., 1978; Hearne et al., 1987) are inconclusive on the human carcinogenicity of methylene chloride. Review of these epidemiological studies and updated evaluation of the cohorts still provide inadequate evidence of human carcinogenicity (EPA, 1 9 8 9 b ) . Experimental carcinogenesis studies indicate that exposure to methylene chloride by the oral route resulted in a significant increase in the incidence of hepatocellular carcinoma and neoplastic nodules in female F344 rats (NCA, 1982; Serota et al., 1986a) and male B6C3F1 mice (NCA, 1983; 115 Serota et al, 1986b). Inhalation studies with methylene chloride produced an increased incidence of mammary tumors in both sexes of Sprague-Dawley (Burek et al., 1980, 1984) and F344 rats (NTP, 1986). The data are judged to be sufficient for Tier 1 HCC derivation. The carcinogenic effects of methylene chloride via the oral route were investigated in two separate 2-year studies sponsored by the National Coffee Association (NCA, 1982, 1983; Serota et al., 1986a, 1986b). In the 1982 study, groups of 85 F344 rats of either sex received nominal doses of 5, 50, 125, or 250 mg / k g / d a y of m e t h ylene c h loride in drinking water. Female rats receiving 50 and 250 mg/kg/day had a significantly increased incidence of combined hepatocellular carcinoma and neoplastic nodules in comparison to matched controls. Male rats, however, did not show an increased incidence of liver tumors. A dosedependent, statistically significant increase in the incidence of salivary gland sarcoma was observed in male rats. A dose-related increase in the average number of benign mammary tumors was observed in female rats. The increased incidence of mammary tumors was observed in male rats, albeit to a lesser degree. The National Coffee Association in its subsequent study (NCA, 1983; Serota et al., 1986b) exposed B6C3F1 mice of 116 e ither sex to 0, 60, 125, 185, or 250 m g / k g / d a y m e t h ylene chloride in drinking water. A statistically significant increase in the incidence of combined hepatocellular carcinoma and neoplastic nodules was observed in male mice exposed to 125 and 185 mg/kg/day. However, only a marginal increase in the incidence of tumorigenesis and reduced average survival time was observed in the 250 mg/kg/day group. Quantitative cancer risk estimates of methylene chloride are based o n NTP inhalation studies in rats and mi c e (NTP, 1986). In this study, groups of 50 male and female F344/N rats and B6C3F1 m i c e were exposed to 0, 1000, 2000, and 4000 ppm (rats), and 0, 2000, and 4000 p p m (mice) for 6 hrs/day, 5 days/week for 102 weeks. Female rats, and to a lesser degree male rats, demonstrated a statistically significant increase in the incidence of mammary gland neoplasms. In mice, methylene chloride elicited an enhanced combined incidence of hepatocellular adenomas and carcinomas in the male (22/50, 24/49, 33/49) and female (3/50, 16/48, and 40/48) mice. Similarly, both male and female mice displayed an increased incidence of alveolar/bronchiolar adenomas and carcinomas (NTP, 1986). In an inhalation study reported by Dow Chemical Company (Burek et al., 1980, 1984), Sprague-Dawley rats and Syrian 117 G o l d e n hamsters of both sexes were exposed to 0, 500, 1500, or 3500 ppm methylene chloride for 6 hrs/day, 5 days/week for 24 months. A statistically significant increased incidence of benign tumors in female hamsters exposed to 3500 ppm was attributed to increased longevity in that group. A statistically significant increase in salivary gland sarcoma was observed in male rats exposed to 3500 ppm methylene chloride. The finding of methylene chlorideinduced salivary gland tumors in male rats is complicated by the observation that these rats had apparently contracted a viral disease, sialodacryoadentitis, in the salivary glands during the earlier phase of the exposure regimen (Burek et al., 1980, 1984). Based on these uncertainties, experimental results from this study were considered inconclusive on the carcinogenicity of methylene chloride. In a subsequent inhalation study by Dow Chemical Company (Nitschke et al., 1982), limited evidence of mammary fibroma/fibrosarcoma was observed in male and female rats exposed to 0, 50, 200, or 500 p p m of m e t h y l e n e c h loride for 2 years. EPA (1989b) derived a recommended oral slope factor from the arithmetic mean of two slope factors derived from the induction of liver tumors in female mice by inhalation (NTP, 1986) and in male mice by drinking water exposure (NCA, 1983; Serota et al., 1986b). These individual slope factors 118 were 2.6 x 10-3 (mg/kg/d)-1 and 1.2 x 10-3 (mg/kg/d)- 1 , r e s pectively (EPA, 1 9 8 9 b ) . This a pproach recom m e n d e d by EPA (1989b) is utilized for Tier l HCC derivation, utilizing an arithmetic mean slope factor of 7.3 x 10-3 (mg/kg/d)- 1 . RAD = 1 x 10-5 = 1 x 10-5______________ q^* 7.3E-3 (mg/kg/day)-1 = 0.00137 mg/kg/day Drinking Water Sources: H C V = RAD x Wh_______ = 0,00137 ma/kcr/dav x 70 ka____________ WC + (FC x BAF) 2 1/day + (0.015 kg/day x 2.5 1/kg*) = 0.047 m g / 1 (rounded off to 0.05 mg/1 (Tier 1)) Nondrinking Water Sources: HCV = RAD x Wh________ = 0.00137 m a / k q / d a v x 70 k g_______ WC + (RC x BAF) 0.01 1/day +(0.015 kg/day x 3.5 1/kg*) = 2.02 mg/1 (rounded off to 2 mg/1 (Tier 1)) 119 Where: PCA. *BAF = 2.5, provided by EPA-Duluth and Minnesota REFERENCES: Burek, J.D., K.D. Nitschke, and T.J. Bell. 1980. Methylene Chloride: A Two-Year Inhalation Toxicity and Oncogenicity Study in Rats and Hamsters. Toxicology Research Laboratory, Health and Environmental Sciences, Dow Chemical Company, Midland, MI. Burek, J.D., K.D. Nitschke, T.J. Bell, D.L. Wackerle, R.C. Childs, J.E. Beyer, D.A. Dittenber, L.W. Rampy, and M.J. McKenna. 1984. Methylene chloride: A two-year inhalation toxicity and oncogenicity study in rats and hamsters. Fundam. Appl. Toxicol. 4:30-47. Friedlander, B . R . , F.T. Hearne, and S. Hall. 1978. Epidemiologic investigation of employees chronically exposed to methylene chloride -- mortality analysis. J. Occup. Med. 20:657-666, Hearne, F.T., F. Grose, J.W. Pifer, B.R. Friedlander, and R.L. Raleigh. 1987. Methylene chloride mortality study: Dose-response characterization and animal model comparison. J. Occup. Med. 29:217-228. National Coffee Association (NCA). 1983. Twenty-Four Month Oncogenicity Study of Methylene Chloride in Mice. Prepared by Hazleton Laboratories America, Inc., Vienna, VA. (Unpublished). National Coffee Association (NCA). 1982. 24-Month Chronic Toxicity and Oncogenicity Study of Methylene Chloride in Rats. Final Report. Prepared by Hazleton Laboratories America, Inc., Vienna, VA. (Unpublished). Nitschke, K.D., J.D. Burek, T.J. Bell, L.W. Rampy, and M.G. McKenna. 1982. Methylene Chloride: A Two-Year Inhalation Toxicity and Oncogenicity Study. Toxicology Research Laboratory, Health and Environmental Sciences, Dow Chemical Company. Midland, MI. (Final Report). National Toxicology Program (NTP). 1986. Toxicology and Carcinogenesis Studies of Dichloromethane (Methylene Chloride) in F344/N Rats and B6C3F1 Mice (Inhalation Studies). NTP-TRS-306. 120 Ott, M.G. L.K. Skory, B.B. Holder, J.M. Bronson and p.R. Williams. 1983. Health evaluation of employees occupationally exposed to methylene chloride -- mortality. Scanc. J. Work Environ. Health. 9:8-16. Serota, D.G., A.K. Thakur, B.M. Ulland, J.C. Kirschman, N.M. Brown, R.H. Coots and K. Morgareidge. 1986a. A twoyear drinking water study of dichloromethane on rodents. I. Rats. Food Chem. Toxicol. 24:951-958. Serota, D.G., A.K. Thakur, B.M. Ulland, J.C. Kirschman, N.M. Brown, R.H. Coots and K. Morgareidge. 1986b. A twoyear drinking water study of dichloromethane on rodents. II. Mice. Food Chem. Toxicol. 24:959-964. U.S. Environmental Protection Agency (EPA). 1985. Addendum to the Health Assessment Document for Dichloromethane (Methylene Chloride). Updated Carcinogenicity Assessment. Prepared by the Carcinogen Assessment Group, OHLA, Washington, DC. EPA 600/8-B2/Q04FF. U.S. Environmental Protection Agency (EPA). 1989a. Risk A s s e s s m e n t Guidance For Superfund, Vol 1., H u m a n Health Evaluation Manual (Part A). EPA/540/1-89/002. Office of Emergency and Remedial Responses (Superfund), Washington, D.C. U.S. Environmental Protection Agency (EPA). 1989b. Integrated Risk Information System (IRIS database). Chemical file for dichloromethane (75-09-2). Verification Date 4/6/89. Last Reviewed 4/6/89. 121 122 September 6, 1991 GREAT LAKES INITIATIVE HUMAN HEALTH CRITERIA FOR POLYCHLORINATED BIPHENYLS (PCBS) CAS NO. 1336-36-3 Tier 2 Human Noncancer Level of Protection Studies of low-level oral PCB exposure in several species have demonstrated effects on serum chemistry, liver toxicity, reproductive capability and other endpoints at doses of less than 5 mg/kg bw/day. The most appropriate data for use in HNV development are the rhesus monkey data due to the high sensitivity of the species and the relative wealth of the database including studies on reproduction and development. Also, as a nonhuman primate the rhesus monkey may serve as the most appropriate model species for potential human effects. Adult male and female rhesus monkeys were administered Aroclor 1248 in the diet at levels of 2.5 and 5.0 ppm for up to 18 months (Allen, 1975; Allen and Barsotti, 1976; Barsotti et al., 1976; Allen et al., 1980). Assuming that rhesus monkeys consume daily an amount of food equivalent to 4% of their body weight, these exposure levels of 2.5 and 5.0 ppm are equivalent to 0.1 and 0.2 mg/kg bw/day, r e s p e c t i v e l y (EPA, 1985). After six m o n t h s exposure, the females were bred with unexposed males. Conception occurred in 12/12, 8/8 and 6/8 of the adult females re c e i v i n g 0, 2.5 Preceding page blank 123 and 5.0 ppm, respectively. The number of live infants born at 0, 2.5 and 5.0 ppm was 12, 5 and 1, respectively. Exposure of the females continued until three months after parturition. Infants were allowed to remain with their mothers and nurse for a minimum of four months, resulting in transmammary as well as transplacental exposure. At both dose levels (dose-specific responses not distinguished), adult females developed acne, alopecia (hair loss), erythema, swelling of the eyelids, abnormal menstrual cycles, and abnormal serum chemistries. Two treated females that d i e d after 173 days (2.5 p p m group) or 310 days (5 ppm group) were found at necropsy to have signs of liver toxicity including focal areas of necrosis. Treated animals were noted to appear more susceptible to the opportunistic intestinal pathogen Shigella flexneri type IV. Offspring of the exposed groups (combined) demonstrated decreased birth weights (399 22 g. vs. 507 59 g. in controls) a nd hyperpigmentation of the skin. Three of the 6 infants died, which was attributed to PCB toxicity. These included the only infant from the 5 ppm group and two of the five infants from the 2.5 ppm group. Necropsies of these infants showed rudimentary thymuses, small spleens, underdeveloped splenic lymph nodes, hypocellularity of the bone marrow and fatty infiltration of liver cells, among other effects. In a continuation of these studies, Barsotti (1980) and Allen et al. (1980) reported that in subsequent breeding trials 12 4 during the recovery period for these same adult female monkeys, effects on reproduction and offspring development were still apparent for greater than one year. Barsotti (1980) and Barsotti and VanMiller (1984) administered diets containing 0, 0.25 or 1.0 p pm Aroclor 1016 to groups of 8 adult female rhesus monkeys for seven months prior to breeding, through gestation and a 4-month nursing period. The total exposure period was 87 9 weeks. Exposure did not result in signs of overt toxicity in adult females. All treated females conceived, carried their fetuses to term and delivered viable offspring. The birth weights from the control, 0.25 and 1.0 ppm groups were 512 64, 491 24 and 422 29 g . , respectively. The newborn weights in the 1.0 ppm group were significantly less than the controls (p less than 0.01). Experimental groups of infants gained weight consistently, and the infant weights among the 1.0 ppm group were not significantly lower than the control group at weaning (864 97g vs. 896 9 0 g ) . The authors determined that during the pre-breeding exposure period, the 0.25 and 1.0 ppm treated females consumed 1.7 0.3 mg/kg bw/7 months and 6.1 0.9 mg/kg bw/7 months, or approximately 0.008 and 0.03 mg/kg bw/day, respectively. Bowman et al. (1981) reported that offspring of female rhesus monkeys fed diets containing 0.5 or 1.0 ppm Aroclor 125 1248 three days per week or 2.5 ppm daily (estimated average doses were 0.006, 0.013 and'0.085 mg/kg bw/day, respectively) displayed greater locomotor activity than controls. However, g roup sizes w e r e small (n = 3-7), the quantitative differences in activity were not dose-related, and the variability within each group was noted to be substantial. Becker et al. (1979) found that groups of 1-2 rhesus m o nkeys fed diets c o n t a i n i n g 3, 10, 30 or 100 ppm of PCBs as Aroclor 1242 for several months had dose-dependent findings of gastric lesions, lack of body weight gain, - reduced hemoglobin, persistent leukocytosis, and early mortality. These effects occurred even in the animal receiving the lowest dose of 3 ppm (0.12 mg/kg bw/day), which expired after 245 days of dosing. The substantial studies of Aroclor 1248 in rhesus monkeys establish a subchronic LOAEL for marked systemic toxicity, reproductive and developmental effects at 2.5 ppm (0.1 mg/kg bw/day). One study utilizing Aroclor 1016 indicates a LOAEL for neonatal weight depression at 1.0 ppm (0.03 mg/kg bw/day) with a NOAEL for this effect at 0.25 ppm (0.008 mg/kg bw/day) (Barsotti, 1980; Barsotti and VanMiller, 1984) . None of these studies were chronic in duration, generally spanning less than 10% of the expected lifespan of about 20 years (Gold et al., 1984). The composition of Aroclor 1016 is primarily di-, tri-, and tetrachloro isomers 126 of biphenyl, w i t h an average chlorine p e r c e n t a g e (41%) that is very similar to Aroclor 1242 (EPA, 1980). Other Aroclor mixtures composed of more highly chlorinated congeners have inadequate data to identify the approximate threshold level for the sensitive systemic and reproductive/developmental effects, EPA (1985) considered the poor metabolism of PCBs and their bioaccumulation tendency, and the severity of effects seen at the subchronic LOAEL of 0,1 mg Aroclor 1248/kg bw/day, and declined recommendation of an Acceptable Daily Intake (ADI). The database is judged insufficient for Tier 1 Human Noncancer Criterion development. A Tier 2 level of protection is derived from the Aroclor 1016 NOAEL in rhesus monkeys at 0.25 ppm (approximately 0.008 mg/kg bw/day). The Tier 2 level of protection is intended to be applicable to all PCB isomers and Aroclor mixtures (i.e., total PCBs) until a more appropriate methodology may be developed. In particular, the dose-response for chronic and reproductive/ developmental effects of the more highly chlorinated PCBs needs to be investigated and characterized. A total uncertainty factor of 1,000 is used in the calculation: ADE = 0.008 ma/kcr/d = 8 x 10~6 mg/kg/d = 8 ng/kg/d 1,000 127 Where: U ncertainty factor = 1,000 composed of: lOx for intraspecies variability lOx for interspecies extrapolation lOx for subchronic exposure duration Drinking Water Sources: HNV - APE Uh RSC - 8 rw/ko/d k 70 fca x 0.B _________________ WC * (FC x BAF) 2 l/d + (0.015 kg/d x 1,776,860 l/kg*) 0.017 ng/L (rowded off to 20 pg/l (Tier 2)) Nondrinking Water Sources: HMV APE x Mh x RSC 8 ng/ko/d x 70 ka > 0.8________________ WC + (FC x BAF) 0.01 L/d (0.015 kg/d x 1,776,860 l/kg*> s 0.017 ng/L Crcunded off to 20 pg/l (Tier 2)) Where: RSC = 0.8; the substance is persistent and bioaccumulative. *BAF = 1,776,860, provided by EPA-Duluth and Minnesota PCA. References: Allen, J.R. 1975. Response of the non-human primate to p o l y c h l o r i n at e d biphenyl exposure. Fed. Proc. 34: 1675-1679. Allen, J.R. and D.A. Barsotti. 1976. The effects of transplacental and mammary movement of PCBs on infant rhesus monkeys. Toxicology. 6:331-340. 128 Allen, J.R., D.A. Barsotti and L.A. Carstens. 1980. Residual effectsof polychlorinated biphenyls on adult n o n human primat e s and their offspring. J. Toxicol. Environ. Health. 6(1):55-66. Barsotti, D.A., R.J. Marlar and J.R. Allen. 1976. Reproductive dysfunction in rhesus monkeys exposed to low levels of polychlorinated biphenyls (Aroclor 1248). Fd. Cosmet. Toxicol. 14:99-103. Barsotti, D.A. 1980. Gross, clinical and reproductive effects of polychlorinated biphenyls (PCBs) in the rhesus monkey. Diss. Abstr. Int. 41(10):3744-5. Barsotti, D.A. and J.P. VanMiller. 1984. Accumulation of a commercial polychlorinated biphenyl mixture (Aroclor 1016) in adult rhesus monkeys and their nursing infants. Toxicology. 30(1):31-44. Becker, G.M., W.P. M c N u l t y and M. Bell. 1979. P o lychlorinated biphenyl induced m o r p h o l o g i c changes in the gastric mucosa of the rhesus monkey. Lab. Invest. 40(3):373-383. Bowman, R.E., M.P. Heironimus and D.A. Barsotti. 1981. Locomotorhyperactivity in PCB-exposed rhesus monkeys. Neurotoxicology. 2(2):251-68. Gold, L.S., et al. 1984. A carcinogenic potency database of the standardized results of animal bioassays. Environmental Health Perspectives. 58:9-319. U.S. Environmental Protection Agency (EPA). 1985. Drinking WaterCriteria Document for Polychlorinated Biphenyls (PCBs). Environmental Criteria and Assessment Office. EPA-600/X-84-198-1. PB-86-118312, Tier 1 Human Cancer Criterion PCBs (as a class) have sufficient carcinogenicity weight-of- evidence for a B2 classification (probable human carcinogen) based on the induction of hepatocellular carcinomas in three strains of rats and two strains of mice and inadequate yet suggestive evidence of excess risk of liver cancer in humans (EPA, 1987). The data are judged sufficient for Ti e r 1 HCC 124 derivation. Although animal feeding studies demonstrate the carcinogenicity of commercial PCB preparations, it is not known which of the PCB congeners in such mixtures are responsible for these effects. EPA (1987) developed a carcinogenicity risk assessment for PCBs with a slope factor derived from Aroclor 1260 data, clearly stating the intent that the assessment be considered representative for all PCB mixtures. The application of this approach to regulatory programs is a prudent approach to ensure adeguate protection of public health. A review of the available carcinogenicity data indicates that the most appropriate studies for quantitative cancer risk assessment are the bioassays of Kimbrough et al (1975) and Norback and Weltman (1985). These studies utilized d i f f erent rat strains -- Sherman rats in the Ki m b r o u g h et al (1975) study, Sprague-Dawley rats in the Norback and Weltman (1985) study -- but otherwise had several similarities. Both utilized large numbers of animals in chronic Aroclor 1260 feeding studies with only one exposure group. Dosed groups received 100 ppm for 630 days in the bioassay by Kimbrough et al. (1975), while Norback and Weltman (1985) administered 100 ppm for 16 months followed by a 50 ppm diet for an additional 8 months, then a basal diet for 5 months. The predominant neoplastic effect in each study was the 130 increased incidence of hepatocellular neoplasms in female rats. Using the linearized multistage procedure, EPA (1987) estimated slope factors of 7.7 (mg/kg/d) 1 and 3.9 (mg/kg/d) 1 from the data of Norback and Weltman (1985) and K i m b rough et al. (1975), respectively. The larger of these slope factors, 7.7 (mg/kg/d) , was selected by EPA (1987) as the preferred slope factor estimate. Although the Norback and Weltman (1985) study included a test protocol of partially hepatectomizing some of the animals, EPA (1987) noted that the study had favorable qualities. The rat strain used (Sprague-Dawley) is known to have a low incidence of spontaneous hepatocellular neoplasms, the study duration spanned the natural life of the animal, and concurrent morphologic liver studies showed the sequential progression of liver lesions to hepatocellular carcinomas. Extrapolation modeling utilized a female rat liver tumor incidence rate of 45/47 in the dosed group. This includes 7 animals which had earlier undergone partial hepatectomy, and the liver tumor incidence for this subgroup was unreported. Exclusion of this group would have very little impact on the resulting slope factor, and the tumor promoting effect of the partial hepatectomization should be minimal (Hiremath, 1991). The 131 Tier 1 Human Cancer Criterion for PCBs is based on the slope factor of 7.7 (mg/kg/d) 1 derived from the rat bioassay of Norback and Weltman (1985). RAD = 1 xj r * _____ _ 1.3 x 10 6 mg/kg/d 7.7 (mg/kg/d) 1 = 1.3 ng/kg/d Drinking Water Sources: HCV = RAP x Uh______ 1.3 rw/ka/d x 70 ka__________________ WC (FC x BAF) 2 l/d + (0.015 kg/d x 1,776,860 L/kg*> 3.4 x 103 ng/t (rounded off to 3 pg/l (Tier 1)) Nondrinking Water Sources: HCV = RAD x Uh 1.3 nq/fco/d x 70 ko___________________ WC (FC x BAF) 0.01 L/d (0.015 kg/d x 1,776,860 L/kg*) => 3.4 x 10-3 ng/l (rouided off to 3 pg/l (Tier 1)) Where: *BAF = 1,776,860, provided by EPA-Duluth and Minnesota PCA. References: Hiremath, C. 1991. Toxicologist, U.S. EPA O f f i c e of Research and Development. Personal communication with R. Sills, M i c h i g a n Department of Natural Resources. 132 Kimbrough, R.D. et al. 1975. Induction of liver tumors in Sherman strain female rats by Aro c l o r 1260. J. National Cancer Institute. 55(6):1453. Norback. D. and R.H. Weltman. 1985. P o l y c h l o r i n at e d biphenyl induction of hepatocellular carcinomas in the Sprague-Dawley rat. Env. Health Persp. 60:97-105. U.S. Environmental Protection Agency (EPA). 1987. Integrated Risk Information System (IRIS database). Chemical file for polychlorinated biphenyls (PCBs) (1336-36-3). Verification Date 4/22/87. Last Revised 1/1/90. 133 134 October 31, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR PENTACHLOROPHENOL CAS NO. 87-86-5 Tier 1 Hinnjin Noncancer Criterion A review of the available literature indicates that HNC derivation for pentachlorophenol (PeCP) is most appropriately based on the chronic oral rat study by Schwetz et al. (1978). Twenty-five rats/sex were administered PeCP in the diet for two years at levels resulting in doses of l, 3, io or 30 mg PeCP/kg bw/day. The test substance was representative of Dowicide EC-7, a commercially available and purified grade of PeCP. An accumulation of pigment in the liver and kidneys was observed in females r e c e iving 10 or 30 mg/kg/day and in males receiving 30 mg/kg/day. The NOAEL was 3 mg/kg/day. The chronic rat NOAEL at 3 mg/kg/day is supported by several subchronic and reproduction/development studies (EPA, 1985a; 1985b). J o h n s o n et al. (1973) adm i n i s t e r e d 3, 10 or 30 mg/kg bw/day purified PeCP to rats for 90 days via feed. Increased liver weights were observed at 10 or 30 mg/kg/d, and increased kidney weights occurred at 30 mg/kg/d. The NOAEL was 3 mg/kg/day. Kimbrough and Linder (1978) administered purified PeCP to rats via diet a r y levels of 20, 100 or 500 ppm for eight months. Only the highest exposure Preceding page blank 135 of 500 ppm (approximately 25 mg/kg bw/day) resulted in hepatocellular changes. Kidney weights were elevated over controls at all dose levels including 20 ppm (approximately 1 mg/kg bw/day), but without a dose-related increase. Goldstein et al. (1977) reported a NOAEL at 5 mg/kg bw/day (100 ppm in feed) to female rats over eight months, with hepatic and body weight effects at 25 mg/kg bw/day (500 ppm in feed). Reproductive/developmental studies have reported a very low tendency for placental transfer (Larsen et al., 1975), a lack of tera t o g e n i c effects, and fetotoxicity at 15 or 30 mg/kg bw/day (Schwetz et al., 1974; 1978). A statistically significant increase in delayed skull ossification has been reported at doses as low as 5 mg/kg bw/day (Schwetz et al., 1974), but no effects on reproduction, neonatal growth, survival, or development occurred at 3 mg/kg bw/day (Schwetz et al., 1978). The EPA (1985b) concluded that 3 mg/kg bw/day could be considered a NOEL for PeCP's fetotoxicity. The quality of the key study and supporting database are judged to be sufficient for Tier 1 HNC development. The derivation of the HNC from the chronic rat NOAEL at 3 mg/kg/day, with lOx uncertainty factors for inter- and intraspecies extrapolation, is consistent with the RfD development by EPA (1985a). 136 ADE = 3 ma/ka/d = 0.03 mg/kg/d 100 Where: Uncertainty Factor = 100, composed of: lOx for intraspecies variability lOx for interspecies extrapolation ' Drinking Water Sources: H N V = A D E x Wh x R S C = 0.03 m a / k a / d x 70 k a _____________ WC + (FC x BAF) 2 1/d + (0.015 kg/d x 650 1/kg*) = 0.179 m g / 1 (rounded off to 0.2 mg/1 (Tier 1)) Nondrinking Water Sources: H N V = A P E x W h x R S C = 0.03 ma/ k a / d x 70 ka________________ WC + (RC x BAF) 0.01 1/d + (0.015 kg/d x 650 1/kg*) = 0.215 mg/1 (rounded off to 0.2 m g / 1 (Tier 1)) Where: PCA. * BAF = 650, provided by EPA-Duluth and Minnesota 137 NOTE: A R e l ative Source C o n t ribution (RSC) factor has not been utilized in these draft calculations. REFERENCES : Goldstein, J.A. et al. 1977. Effects of pentachlorophenol on hepatic drug-metabolizing enzymes and porphyria related to contamination with chlorinated dibenzo-pdioxins and dibenzofurans. Biochem. Pharmacol. 26:1549-1557. Johnson, R.L., et al. 1973. Chlorinated dibenzodioxins and pentachlorophenol. Environ. Health Perspect. 5:171 175. Kimbrough, R.D. and R.E. Linder. 1978. The effect of technical and purified pentachlorophenol on the rat liver. Toxicol. Appl. Pharmacol. 46:151-162. Larsen, R.V. et al. 1975. Placental transfer and teratology of pentachlorophenol in rats. Environmental Letters. 10(2): 121-128. Schwetz, B.A. et al. 1974. The effect of purified and commercial grade pentachlorophenol on rat embryonal and fetal development. Toxicol. Appl. Pharmacol. 28(1): 151-161. Schwetz, B.A. et al. 1978. Results of two-year toxicity and reproduction studies on pentachlorophenol in rats. In: Rao, K.R. (Ed.). 1978. Pentachlorophenol Chemistry, Pharmacology, and Environmental Toxicology. Plenum Press, New York, NY. pp. 301-309. U.S. Environmental Protection Agency (EFA). 1985a. Integrated Risk Information System (IRIS database). Chemical file for Pentachlorophenol (87-86-5). Verification Date 5/20/85. Last Revised 5/7/91. U.S. Environmental Protection Agency (EPA). 1985b. Drinking Water Criteria Document for Pentachlorophenol. Prepared for the Office of Drinking Water, by the Environmental Criteria and Assessment Office. EPA600/X-84-177-1. PB-86-118015. 138 Tier 1 Human Cancer Criterion A review of the available literature on the carcinogenicity of pentachlorphenol (PeCP) indicates a lack of evidence of human or animal carcinogenicity prior to NTP (1989). As pointed out in NTP (1989), prior studies were generally limited by study design flaws. EPA has concluded that the evidence provided by NTP (1989) would support classification of PeCP into Group B2; probable human carcinogen (EPA, 1989/ 1990) . The conclusions of NTP (1989) are that under the conditions of the 2-year feeding studies, there was clear evidence of carcinogenic activity for male and female B6C3F1 mice exposed to PeCP as Dowicide EC-7, and for male mice exposed to PeCP as a technical grade composite. Additionally, there was some evidence of carcinogenic activity for female mice exposed to the PeCP technical-grade composite. The database is judged to be sufficient for Tier 1 HCC derivation. The study design of NTP (1989) involved exposure via feed for 103 weeks to groups of 50 mice/sex, w i t h g r oups of 35/sex serving as controls for each grade of PeCP studied. The grades of PeCP selected for the chronic bioassays were commercial samples (Dowicide EC-7 and a technical-grade composite), considered to be representative of PeCP forms 139 which humans are exposed to. Exposure levels in feed were 0, 100 or 200 p p m te c h n i c al-grade PeCP or 0, 100, 200 or 600 ppm Dowicide EC-7. There was clear evidence that administration of the technical-grade composite resulted in increased incidences of hepatocellular and adrenal medullary neoplasms in male mice, and some evidence of induction of hepatocellular tumors and hemangiosarcomas in female mice. There was clear evidence that the administration of Dowicide EC-7 resulted in hepatocellular tumors and adrenal medullary pheochromocytomas in both sexes and hemangiosarcomas in female mice. Among the three types of tumors induced by pentachlorophenol, the EPA Science Advisory Board considered the hemangiosarcomas to be the tumor of greatest concern (EPA, 1990). To give preference to the hemangiosarcoma data and because some male groups experienced significant early loss, only the female mice are used in the quantitative risk assessment. Pooled tumor incidence was utilized, with exclusion of animals which died prior to the first tumor observation. The slope factor of 1.2 E-l per (mg/kg)/day is calculated as the geometric mean of the slope factors for each pentachlorophenol preparation. This approach is consistent with EPA (1990). R A D = 1 x 10~5_____________ = 8.3 3 x 10" 5 mg/ k g / d 140 1.2 E-l (mg/kg/d)-l = 8 3 . 3 ng/kg/d Drinking Water Sources: HCV = RAD x Wh________ = 83.3 na/ka/d x 70 ka_____________ WC + (FC X BAF) 2 1/d + (0.015 kg/d X 650 1/kg*) = 496 ng/1 (rounded off to 0.5 ug/1 (Tier 1)) Nondrinking Water Sources: H C V = RAD x Wh________ = 83.3 na/ka/d x 70 kg_________________ W C + (RC x BAF) 0.01 1/d + (0.015 kg/d x 650 1/kg*) = 597 ng/1 (rounded off to 0.6 ug/1 (Tier 1)) Where: PCA. *BAF = 650, provided by EPA-Duluth and Minnesota REFERENCES: NTP. 1989. Toxicology and Carcinogenesis Studies of Two Pentachlorophenol Technical-Grade Mixtures in B6C3F1 Mice (Feed Studies). U.S. DHHS. Technical Report Series No. 349. 141 U.S. Environmental Protection Agency (EPA). 1989. 54 FR 22062-22160. National Primary and Secondary Drinking W a t e r Regulations; Proposed Rule. M ay 22, 1989. U.S. Environmental Protection Agency (EPA). 1990. Integrated Risk Information System (IRIS database), chemical file for Pentachlorophenol (87-86-5). Verification Date 8/2/90. Last Revised 5/7/91. 142 Oct o b e r 23, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR 2 , 3 , 7 , 8 -TETRACHLORODIBENZO-P-DIOXIN (2,3,7,8- T C D D ). CAS NO. 1746-01-6 Tier 1 Human Noncancer Criterion Of the many subacute and chronic studies available for 2,3,7,8-TCDD, a few stand out as supporting Tier 1 criterion derivation. In a two-year toxicity and oncogenicity study, rats were a d m inistered doses of 0, 0.001, 0.01 and 0.1 ug/kg bw/day of 2,3,7,8-TCDD via diet (Kociba et al., 1978). Animals given the high dose exhibited increased mortality, decreased weight gain, slight depression of erythroid parameters, increased urinary excretion of porphyrins and delta-aminolevulinic acid and increased serum levels of certain enzymes. Histopathologic or gross effects were seen in liver, lymphoid, lung and vascular tissues. An increased tumor incidence was also seen. Similar effects, but to a lesser degree, were seen in mid-dose animals. A NOAEL of 0.001 ug/kg / d a y (1 ng/kg/day) was r eported in this study. A NOAEL of 0.001 ug/kg bw/day via feed exposure was also reported in a three-generation rat reproduction study (Murray et al., 1979). At 0.1 ug/kg/day, decreases in FQ genera t i o n fertility and F^^ generation litter size were reported. At 0.01 ug/kg/day, significant decreases in fertility were seen in the F^ and F 2 generations; other 14 3 effects included decreased litter size at birth, decreased gestational survival and decreased neonatal growth and survival. The reproductive capacity of the low dose rats did not appear to be significantly affected in any generation. However, a rvaluation of these data using different statistical methods indicated that both lower dose levels resulted in significant reductions in offspring survival indices, increases in liver and kidney weight of pups, decreased thymus weight of pups, decreased neonatal weights and increased incidence of dilated renal pelvis (Nisbet and Paxton, 1982). Nisbet and Paxton (1982) c o n c luded that 0.001 u g / k g / d a y (1 ng/kg/day) was no t a N O E L in the Murray et al. (1979) study. Kimmel (1988) considered the data of Murray et al. (1979) to be suggestive of a pattern of decreased offspring survival and increased offspring renal pathology even at 0.001 ug/kg/day, although the pooling of data from different generations by Nisbet and Paxton (1982) was considered biologically inappropriate. Studies by Schantz et al. (1979) and Allen et al. (1979) suggest that rhesus monkeys are more sensitive to 2,3,7,8TCDD than rats. When monkeys were administered 50 ppt 2,3,7,8-TCDD in feed for 7 to 20 months, decreases in fertility, increases in abortions and other toxic effects (alopecia, hyperkeratosis, weight loss, decreased hematocrit and white blood cell count and increased serum levels of 144 SGPT) were noted. The 50 ppt dietary residue level corresponds to a daily dose of 1.5 ng/kg bw/day (EPA, 1984). Therefore, 1.5 ng/kg/day can be considered a LOAEL for rhesus monkeys from these studies. In a continuation of the rhesus monkey studies by Schantz et al. (1979) and Allen et al. (1979), Bowman et al. (1989a, 1989b) have evaluated the effects of 5 and 25 ppt 2,3,7,8TCDD in feed on reproduction and on behavior, respectively. Breeding of the animals after 7 and 24 m o n t h s of exposure resulted in impaired reproductive success at 25 ppt but not at 5 ppt (approximately 0.67 and 0.13 ng/kg bw/day, respectively). The exposures were discontinued after 4 years, and a third breeding ten months post-exposure did not indicate reproductive impairment (Bowman et al., 1989a). The offspring from these breeding experiments were evaluated for development and behavioral effects utilizing several testing methods (Bowman et al., 1989b). Although there were no significant effects of TCDD exposure on birth weight, growth, or physical appearance of the offspring, some behavioral test results were interpreted to be indicative of TCDD effects. These included alterations in the social behavior between the mothers and their infants and of peer groups of the offspring after weaning. However, the study groups were very limited in size and the statistical and biological significance of the findings are unclear. This 145 study may be interpreted to provide only suggestive evidence of possible behavioral effects. The reproduction study of Bowman et al. (1989a) provides much clearer evidence of a LOAEL at 25 ppt (0.67 ng/kg/day) and a NOAEL at 5 ppt (0.13 ng/kg/day). The EPA has used the equivocal evidence for a rat LOAEL at 1 ng/kg/day, supported by an unequivocal rhesus monkey LOAEL at 1.5 ng/kg/day, in the development of an Acceptable Daily Intake (ADI) (EPA, 1984; 1985a) and Drinking Water Equivalent Level (DWEL) (EPA, 1985b; 1990). In light of the more recent rhesus monkey study of Bowman et al. (1989a), there is improved resolution of the threshold for the sensitive effect of reproductive impairment in this species. The Human Noncancer Criterion is based on the NOAEL of 0.13 ng/kg/day for reproductive effects from this study. The entirety of the rhesus monkey studies, supported by the evidence in rats cited above, is judged sufficient for Tier 1 criterion development. A D E = 0.13 n a /ka/d = 1.3 x 10" 3 ng/kg/d = 1.3 pg/ k g / d 100 Where: Uncertainty Factor = 100, composed of: lOx for intraspecies variability lOx for interspecies extrapolation 146 Drinking Water Sources: H N V = APE x Wh x RSC = 1.3 pg/ka/d x 70 kg x 0.8____________ WC + (FC X BAF) 2 1/d + (0.015 kg/d X 50,000 1/kg*) = 0.097 pg/1 (rounded off to 0.10 pg/1 (Tier 1)) Nondrinking Water Sources: HNV = APE x Wh x RSC = 1.3 /kq/d x 70 kg x 0.8__________ VC + (FC x BAF) 0.01 t/d + (0.015 kg/d x 50,000 l/kg*) = 0.097 pg/l (roLnded off to 0.10 pg/l (Tier 1)) Where: RSC = 0.8; the substance is persistent and bioaccumulative. *BAF = 50,000, provided by EPA-Duluth and Minnesota PCA. REFERENCES: Allen, J.R. et al. 1979. Reproductive effects of halogenated aromatic hydrocarbons on nonhuman primates. Ann. NY Acad. Sci. 320:419-425. Bowman, R.E., et al. 1989a. Chro n i c d i e t a r y intake of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) at 5 or 25 parts per trillion in the monkey: TCDD kinetics and dose-effect estimate of reproductive toxicity. Chemosphere. 1 8 (1 - 6): 243-252. 147 Bowman, R.E., et al. 1989b. Behavioral effects in monkeys exposed to 2,3,7,8-TCDD transmitted maternally during gestation and for four months of nursing. Chemosphere. 18(1-6) :235-242. Rimmel, G.L. 1988 . A p p e n d i x C. Rep r o d u c t i v e and DevelopmentalToxicity of 2,3,7,8-TCDD. Reproductive Effects Assess m e n t Group, OHEA/ORD, EPA. In: EPA. 1988. A Cancer Risk-Specific Dose Estimate for 2,3,7,8-TCDD. Appendices A-F. Review Draft. E P A / 6 0 0 / 6 - 8 8 / 0 0 7 Ab. ` Kociba, R. J. et al. 1978. Results of a two - y e a r chro n i c toxicity and oncogenicity study of 2,3,7,8tetrachlorodibenzo-p-dioxin in rats. Toxicol. Applied Pharmacol. 46:279-303. Murray, F. J. et al. 1979. T h r e e - g e n e r a t i o n r e p r o duction study of rats given 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the diet. Toxicol. Applied Pharmacol. 50:241-252. Nisbet, I.C.T. and M.B. Paxton. 1982. Statistical aspects of three-generation studies of the reproductive toxicity of TCDD and 2,4,5-T. The American Statistician. 36(3): 290-298. Schantz, S. L. et al. 1979. Toxic o l o g i c al effects p r oduced in nonhuman primates chronically exposed to 50 ppt TCDD. Toxicol. Applied Pharmacol. 48:A180. (Abstract No. 360). U.S. Environmental Protection Agency (EPA). 1984. Ambient Water Quality Criteria for 2,3,7,8-Tetrachlorodibenzop-dioxin. Office of Water Regulations and Standards. EPA 440/5-84-007. U.S. Environmental Protection Agency (EPA). 1985a. Health Assessment Document for Polychlorinated Dibenzo-p- dioxins. Office of Health and Environmental Assessment. EPA/600/8-84/014F. U.S. Environmental Protection Agency (EPA). 1985b. Drinking Water Criteria Document for 2,3,7,8Tetrachlorodibenzo-p-dioxin. ECAO/ODW. EPA-600/X-84194-1. PB 86-117983. 148 U.S. Environmental Protection Agency (EPA). 1990. 55 Federal Register No. 143. Wednesday, July 25, 1990. National Primary and Secondary Drinking Water Regulations; Synthetic Organic Chemicals and Inorganic Chemicals. Proposed rule. Tier l winnaTi cancer criterion The EPA (1984) evaluated the available epidemiological and animal bioassay data on the potential carcinogenicity of 2.3.7.8- TCDD. They determined that some case-control studies provide limited evidence for the human carcinogenicity of phenoxy acids and/or chlorophenols, which contain impurities including 2,3,7,8-TCDD. They concluded that the evidence for the human carcinogenicity of 2,3,7,8TCDD based on the epidemiologic studies is only suggestive due to the difficulty of evaluating the risk of 2,3,7,8-TCDD exposure in the presence of the confounding effects of phenoxy acids and/or chlorophenol. Recently published epidemiology studies may be interpreted to provide suggestive evidence of carcinogenicity (Zober et al., 1990; Fingerhut et al., 1991). The potential use of these new studies for quantitative risk assessment has not yet been fully explored. With regard to animal bioassays, the EPA (1984) concluded that several rodent studies establish that 2.3.7.8- TCDD is an animal carcinogen in multiple species and organs and is probably carcinogenic in humans. The weight of evidence of carcin o genicity is s ufficient for Group B2 classification (probable human carcinogen), and satisfies the database requirements for Tier 1 criterion derivation. 149 Among the carcinogenicity bioassays, NTP conducted bioassays with both Osborne-Mendel rats and B6C3F1 mice (NTP, 1 9 8 2 a ) , Groups of 50 mice and 50 rats of each sex were given 2,3,7,8-TCDD in corn oil-acetone by gavage twice per week for 104 weeks. Doses of 0, 0.01, 0.05 or 0.5 u g/kg/week were administered to rats and male mice while female mice r e c eived 0, 0.04, 0.2 or 2.0 ug/kg/week. Controls consisted of 75 rats and 75 mice of each sex. Animals were killed at weeks 105-107. 2,3,7,8-TCDD caused an increased, doserelated incidence of follicular-cell adenomas or carcinomas of the thyroid in male rats. A significant increase in subcutaneous tissue fibromas was also seen in high-dose males. High-dose female rats exhibited increased incidences of hepatocellular carcinomas and neoplastic nodules, subcutaneous tissue fibrosarcomas and adrenal cortical adenomas. In male and female mice, 2,3,7,8-TCDD induced an increased dose-related incidence of hepatocellular carcinomas. High-dose female mice also exhibited increased incidences of thyroid follicular-cell adenomas. In a dermal study also conducted under contract for NTP (NTP, 1 982b), 30 male and 30 female Swiss Webster mice were treated with 2,3,7,8-TCDD in acetone for 3 days/week for 104 weeks. Doses of 0.005 ug and 0.001 ug 2,3,7,8-TCDD were administered to the clipped backs of males and females, respectively. A similar group was pretreated with one ISO application of 50 ug dimethylbenzanthracene (DMBA) one week before 2,3,7,8-TCDD administration. 2,3,7,8-TCDD induced a statistically significant increase of fibrosarcomas in the integumentary system of females given both 2,3,7,8-TCDD alone and following a single application of DMBA. Van M iller et al. (1977) administered diets c ontaining 0, 0.001, 0.005, 0.05, 1, 50, 500 and 1000 p p b 2 , 3 , 7 , 8-TCDD to groups of 10 male Sprague-Dawley rats. Animals received the diets for 78 weeks and were then placed on control feed until they were kille d at w e e k 95. Al l rats fed the higher concentrations (1-1,000 ppb) died early. A variety of tumors were produced and the total number of animals with tumors generally increased, but the small number of animals limits the value of the data. Kociba et al. (1978) administered 2,3,7,8-TCDD via the diet to groups of 50 male and 50 female Sprague-Dawley rats for 2 years. Control groups consisted of 86 animals of each sex. The doses administered were 0, 0.001, 0.01 and 0.1 ug/kg/day. 2,3,7,8-TCDD induced an increased incidence of hepatocellular carcinomas and hepatocellular hyperplastic (neoplastic) nodules in female rats at the two highest dose levels. The highest dose of 2,3,7,8-TCDD also induced an increase in the incidence of stratified squamous cell carcinomas of the hard palate and/or nasal turbi n a t e s in 151 both males and females, squamous cell carcinomas of the t ongue in males and squamous cell c arcinomas of the lungs in females. Kociba et al. (1978) is chosen as the basis for quantitative cancer risk assessment. The Kociba study found that the principal target organ for 2,3,7,8-TCDD-induced tumors was the liver in female rats, demonstrating a dose-related statistically significant increase of hepatocellular carcinomas and hyperplastic (neoplastic) nodules. For quantitative risk assessment, the data were adjusted for early mortality by eliminating those animals that died during the first year of the study. Also, in the mid-dos,e group, two of the reported 20 females with tumors had both nodules and carcinomas; 18 aff e c t e d animals w e r e u s e d as the input for the dose group. Using the linearized multistage model, the resulting slope factor for 2,3,7,8-TCDD is 1.51 x 105 (mg/kg/day)_ 1 . However, an independent pathologist (Squire) was engaged by EPA to reevaluate the h i s t o p a t h o l og i c slides from the Kociba study (EPA, 1984). Squire reported higher tumor incidences than Kociba, generating a slightly higher slope factor of 1.61 x 105 (mg/kg/day)" 1 . EPA (1984) used an average of the two slope factors, 1.56 x 10s (mg/kg/day)"1 , to generate surface water criteria. 152 In March 1990 a panel of seven independent pathologists referred to as the Pathology Work i n g Group (PWG) blindly reevaluated the female rat liver slides from K o c i b a et al. (1978). Liver lesions were classified according to the National Toxicology Program's 1986 liver tumor classification scheme (Sauer, 1990). Using the linearized multistage model, the liver tumor incidence rates reported by the PWG result in a slope factor of 5.1 x 104 (mg/kg/day)-1 for liver tumors only, and a slope factor of 7.5 x 104 (mg/kg/day)"1 for pooled significantly increased tumors of the liver, lung or nasal turbinates/hard palate. The latter method avoids double-counting of tumor-bearing animals (Bayard, 1990). The Human Cancer Criterion is based on the pooled significant tumors in female rats of Kociba et al. (1978) with the liver tumor rvaluation of the Pathology Working Group (Sauer, 1990). The linearized multistage model generates a slope factor of 7.5 x 104 (mg/kg/day)"1 from these data. RAD = 1 X 10~5_______________ = 1.33 x 1 0 " 10 m g / k g / d 7.5 x 104 (mg/kg/d)"1 = 0.133 pg/kg/d 153 Drinking Water Sources: H C V = RAD x Wh_______ = 0.133 pg/ka/d x 70 k:cr_______________ WC + (FC x BAF) 2 1/d + {0.015 kg/d x 50,000 1/kg*) = 1.2 x 10"2 pg/1 (rounded off to 0.01 pg/1 (Tier l)) NondrinJcing Water Sources: HCV = RAD x Uh 0.133 pg/lcg/d x 70 kg______________ WC + (FC x BAF) 0.01 l/d + <0.015 kfl/d x 50,000 l/kg*) 1.2 x 10'* ps/t (rouxled o f f t o 0.01 pg/l (Tfer 1)) Where: PCA. *BAF = 50,000, provided by EPA-Duluth and Minnesota REFERENCES: Bayard, S. 1990. T o x i c o l o g i s t/ S t a t i s t i c i a n wi t h the U.S. EPA Office of Research and Development, Human Health A s s e s s m e n t Group. Personal c o m munication w i t h R. Sills, Michigan Department of Natural Resources. Fingerhut, M. et al. Cancer mortality in workers exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin. The New England Journal of Medicine. 234(4):212-218. Kociba, R.J. et al. 1978. Results of a two-year chronic toxicity and oncogenicity study of 2,3,7,8tetrachlorodibenzo-p-dioxin in rats. Toxicol. Applied Pharmacol. 46:279-303. National Toxicology Program (NTP). 1982a. Bioassay of 154 2.3.7.8- tetrachlorodibenzo-p-dioxin in Osborne-Mendel Rats and B6C3F1 Mice (Gavage Study). NTP-TR-209. National Toxicology Program, U.S. DHHS, Research Triangle Park, NC. National Toxicology Program (NTP). 1982b. Carcinogenesis Bioassay of 2,3,7,8-tetrachlorodibenzo-p-dioxin in Swiss-Webster Mice (Dermal Study). NTP-TR-201. National Toxicology Program, U.S. DHHS, Research Triangle Park, NC. Sauer, R.M. 1990. Pathology Working Group: 2,3,7,8Tetrachlorodibenzo-p-dioxin in Sprague-Dawley Rats. Pathco, Inc. Submitted to the Maine Scientific Advisory Panel. U.S. Environmental Protection Agency (EPA). 1984. Ambient Water Quality Criteria for 2,3,7,8-Tetrachlorodibenzo-pdioxin. EPA 440/5-84-007. Van Miller, J.P. et al. 1977. Increased incidence of neoplasms in rats exposed to low levels of 2,3,7,8tetrachlorodibenzo-p-dioxin. Chemosphere 6(10):625-632. Zober, A., P. M e s serer and P. Huber. 1990. T h i r ty-fouryear mortality follow-up of BASF employees exposed to 2.3.7.8- TCDD after the 1953 accident. Int. Arch. Occup. Environ. Health. 62(2):139-157. 155 156 October 31, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR TOLUENE CAS NO. 108-88-3 Tier 1 Human Noncancer Criterion A review of the available literature indicates inadequate human data for quantitative risk assessment of toluene. Occupational exposure, cigarette smoking and deliberate inhalation of solvents found in various preparations ("glue sniffing") are common means of human exposure to toluene (NTP, 1990). Chronic exposure to toluene vapors at levels of approximately 200-800 ppm have been associated primarily with CNS effects, possibly peripheral nervous system effects, hepatomegaly and hepatic function changes, and renal function affects (EPA, 1987). Although these findings provide qualitative evidence of the human toxicity of toluene, specific exposure levels were not provided and the se data do not provide a do s e - r e s p o n se r e l a t i o n s h i p (EPA, 1987; EPA, 1990; NTP, 1990). The majority of the subchronic-chronic studies on toluene are inhalation studies determining behavioral or histopathologic effects o f t o l u e n e exposure. 157 Preceding p a g e blank The most appropriate basis for HNV derivation for toluene is the NOAEL from a 13-week rat gavage study (NTP, 1990). In this study, toluene in corn oil was administered by gavage to groups of weanling F344/N rats and B6C3F1 mice (10/sex/group) at dose levels of 0, 312, 625, 1250, 2500 or 5000 mg/kg, 5 days per week for 13 weeks. All rats at 5000 mg/kg died during the first week, and 8/10 rats at 2500 mg/kg died, two of which were due to gavage errors. Histopathologic changes were observed in the liver, kidney, brain and urinary bladder at > 1250 mg/kg. The NOAEL for the rats is 312 mg/kg/day with a LOAEL based on liver and kidney weight changes in male rats at 625 mg/kg. Because the exposure was for 5 days/week, these doses are converted to time-weighted-average doses of 223 and 446 mg/kg/day, respectively (EPA, 1990). As described above, NTP (1990) also conducted a 13-week gavage study in B6C3F1 mice. All mice that received 5000 mg/kg died during the first week, and 40% of those that received 2500 mg/kg died before the end of the 13-week gavage study. Clinical signs observed in mice at 2500 mg/kg included sub-convulsive jerking, prostration, impaired grasping reflex, bradypnea, hypothermia, hypoactivity and ataxia. The final mean body weight of males at 2500 mg/kg was lower than that of vehicle controls. At 1250 mg/kg, relative liver weights were increased for mice, but this 158 increase was not statistically significant. The NOAEL for this study was 1250 mg/kg and the LOAEL was 2500 mg/kg. Adjusting the doses for 5 days/week exposure provides timeweighted-average NOAEL and LOAEL doses of 893 and 1786 mg/kg/day, respectively. Another subchronic oral toxicity study was conducted by Wolf et al. (1956), in w h i c h female W i s t a r rats we r e administered toluene by stomach tube at doses of 0, 118, 354 and 590 mg/kg/day, 5 days/week for a total of 138 doses (converted to t i m e - w e i ghted-avera g e doses of a p p roximately 0, 18, 253, and 422 mg/kg/day per EPA, 1990). No adverse effects were observed at any dose level in any of the parameters monitored: growth, appearance and behavior, mortality, organ/body weight, blood urea nitrogen levels, bone marrow counts, peripheral blood counts or morphology of major organs. The HNOAEL for this study was 422 mg/kg/day as the time-weighted-average dose. NYLAR mice were exposed pre- and post-natally to toluene p r o vided in the drinki n g w ater at c o n c e ntrations of 0, 16, 80 and 400 ppm (Kostas and Hotchin, 1981). Rotorod performance was measured at 45 and 55 days of age. An inverse dose-response relationship in the effects was noted in which rotorod performance was improved with increasing dose. No effects were observed for the following 159 reproductive measurements: maternal fluid consumption, offspring mortality rate, development of eye or ear openings, or surface-righting response, resulting in a HNOAEL of 400 ppm. Assuming mice consume water at a p p oximately 0.24 1/kg bw/day (EPA, 1988), the H N O A E L was approximately 96 mg/kg/day. Nawrot and Staples (1979) administered 0.3, 0.5 or 1.0 ml/kg bw toluene, 3 times/day (equivalent to approximately 780, 1300 and 2600 mg/kg/day, per EPA, 1990) to pregnant CD-I mice from days 6-15 of gestation. No method of exposure was mentioned in this limited information abstract. Teratogenic effects were reported at 2600 mg/kg/day and increased embryolethality was reported at 780 mg/kg/day, therefore the LOAEL for this study was 780 mg/kg/day. No other subchronic-chronic oral toxicity studies were identified in the available literature for toluene. Chronic inhalation studies include NTP (1990), in which F344/N rats and B6C3F1 mice (60/sex/dose) were exposed to vapors of toluene, 6.5 hours/day, 5 days/week for 2 years. Dose levels w e r e 0, 120 (mice only), 600 or 1200 ppm. T en animals/group (except male mice) were removed at 15 months for toxicologic evaluation. At 15 months, there was an increased incidence and severity in the erosion of olfactory epithelium, and degeneration of respiratory epithelium was 160 increased in the exposed rats, At the end of the study inflammation of nasal mucosa and metaplasia of olfactory epithelium were increased in exposed females rats. Nephropathy was seen in almost all rats with a severity somewhat increased in exposed rats. For mice, no biologically relevant increase in any nonneoplastic lesion was observed. Chronic inhalation of toluene was studied in F344 rats exposed to 30, 100 or 300 p p m (113, 377 or 1130 m g / m 3 ) toluene 6 hours/day, 5 days/week for 24 months (Gibson and Hardisty, 1983; C U T , 1980, as cited in NTP, 1990; EPA, 1990; and EPA, 1987). A dose-related reduction in hematocrit values was reported in female rats exposed to 100 and 300 ppm. Increased corpuscular hemoglobin concentration was reported in females at 300 ppm. In a perinatal study with CD-l mice (Courtney et al., 1986), toluene was administered by inhalation at 200 and 400 ppm (750 and 1500 mg/m3 , respectively) to pregnant female CD-l mice 7 hours/day from days 6-16 of gestation. Fetotoxicity was observed at 400 ppm with a significant shift in the fetal rib profile. An increased body weight in the neonates on day 1 postpartum was observed at 400 ppm. At 200 ppm, there was an increase in dilated renal pelves which the authors concluded might reflect desynchronization of 161 maturation with respect to development and growth. Assuming that m i c e breathe appoximately 1.7 m 3 /kg/day (EPA, 1988), the 7 hours/day 200 ppm LOAEL and 400 ppm FEL convert to approximately 370 and 740 mg/kg/day, respectively, as daily administered doses. The Tier 1 HNC is derived from the NOAEL dose of 312 mg/kg (converted to 223 mg/kg/day for 5 days/week administration) from the 13-week oral rat study by NTP (1990) with an uncertainty factor of 1000. The uncertainty factor accounts for inter- and intraspecies extrapolation, and for subchronic-to-chronic extrapolation. This approach should be protective of developmental effects, as suggested by the limited developmental toxicity data. This approach is consistent with the risk assessment of toluene for the oral RfD and the drinking water health advisory developed by EPA (1990; 1987). ADE = 223 ma/kq/d = 0.223 mg/kg/day 1,000 Where: Uncertainty Factor = 1,000, composed of: lOx for intraspecies variability lOx for interspecies extrapolation lOx for subchronic exposure duration 162 Drinking Water Sources: H NV = ADE X Wh X RSC = 0.223 mq/ka/d X 70 kq_______ WC + (FC x BAF) 2 1/d + (0.015 kg/d x 40 1/kg*) = 6.00 mg/1 (rounded off to 6 mg/1 (Tier 1)) Nondrinking Water Sources: HNV = A P E x W h x RSC = 0.223 m g / k a / d x 70 ko________ WC + (FC x BAF) 0.01 1/d + (0.015 kg/d X 40 1/kg*) = 25.59 mg/1 (rounded off to 26 mg/1 (Tier 1)) Where: *BAF = 40, pro v i d e d by E PA-Duluth and M i n n e s o t a PCA. NOTE: A Relative Source Contribution (RSC) factor ha s not been utilized in these draft calculations. REFERENCES: Chemical Industry Institute of Technology ( C U T ) . 1980. A 24-Month Inhalation Toxicology Study in Fischer-344 Rats Exposed to Atmospheric Toluene. C U T , Research Triangle Park, NC. As cited in EPA, 1987; EPA, 1990; NTP, 1990. Courtney, K.D., J.E. Andrews, J. Springer, M. Mnache, T. Williams, L. Dailey and J.A. Graham. 1986. A 163 perinatal study of toluene in CD-I mice. Fundam. Appi. Toxicol. 6:145-154. Gibson, J.E. and J.F. Hardisty. 1983. Chronic toxicity and oncogenicity bioassay of inhaled toluene in Fischer-344 rats. Fundam. Appl. Toxicol. 3:315-319. Kostas, J. and J. Hotchin. 1981. Behavioral effects of low-level perinatal exposure to toluene in mice. Neurobehav. Toxicol. Teratol. 3:467-469. National Toxicology Program (NTP). 1990. Toxicology and Carcinogenesis Studies of Toluene (CAS No. 108-88-3) in F344/N Rats and B6C3F1 Mice (Inhalation Studies). NTP- TR. No. 371, NIH Publication No. 90-2826. Nawrot, P.S. and R.E. Staples. 1979. Embryo-fetal toxicity and teratogenicity of benzene and toluene in the mouse. Teratology. 19:41A (abstract). U.S. Environmental Protection Agency (EPA). 1990. Integrated Risk Information System (IRIS database). Chemical file for toluene (108-88-3). Verification Date 6/20/90. Last Reviewed 6/20/90. U.S. Environmental Protection Agency (EPA). 1988. Recommendations for and Documentation of Biological Values for Use in Risk Assessment. PB88-179874. U.S. Environmental Protection Agency (EPA). 1987. Drinking Water Criteria Document for Toluene. Prepared by the Office of Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH for the Office of Drinking Water, Washington, DC. ECAO-CIN-408. Wolf, M . A . , V.K. Rowe, D.D. McCollister, R.L. Hollingsworth and F. Oyen. 1956. To x i c o l o g i c al studies of certain alkylated benzenes and benzene. A.M.A. Arch. Ind. Health. 14:387-398. 164 October 28, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR TOXAPHENE CAS NO. 8001-35-2 Tier i Human Noncancer criterion A review of the available literature indicates that liver pathology and immunological effects are the critical noncancer endpoints for chronic exposure to toxaphene. In a study by Lehman (1952, as cited by EPA, 1985; 1987), rats were e x p osed for a lifetime to 0, 25 or 100 p pm toxaphene in their diet. Fatty degeneration was found in the livers of rats fed 100 ppm, but no adverse effects were observed in animals fed 25 ppm. The dose of 25 ppm (1.25 mg/kg/d) was considered the NOEL for this study. However, in a study conducted with rats by Fitzhugh and Nelson (1951, as cited by EPA, 1985; 1987) there was an increase in the weight of the liver with minimal cell enlargement at dietary levels of 25 ppm. The dose of 25 ppm (1.25 mg/kg/d) was considered the LOEL for this study. Concentrations in food in these two studies were converted to mg/kg bw/d by assuming the rats consumed 5% of their body weight in food per day (EPA, 1987). An unpublished two year study in monkeys conducted by Hercules, Inc. found no clinical or histological effects at dietary exposure levels of toxaphene ranging from 0.64 to 0.78 m g / k g bw/d (EPA, 1980). Details of this study were not available for development of this criterion. 165 The effect of toxaphene on the immune system of SwissW e bster mice was examined by A l l e n et al. (1983). In the first part of the study, female w eanling m i c e were fed 10, 100, or 200 ppm toxaphene for eight weeks. According to EPA (1987) these values are equivalent to 1.5, 15 and 30 mg/kg/d, respectively. The formation of IgG antibody was depressed in animals receiving 100 and 200 ppm toxaphene. Liver/body weight ratios were increased in females at the 100 and 200 ppm level. In the second part of the study, mature female mice were fed the same amounts of toxaphene used in the first part of the study. These animals were mated three weeks after feeding began and were maintained on the treatments until three weeks after parturition at which time the pups were weaned onto the control diet. The results of the assays which were performed on the offspring eight weeks after birth showed suppressed antibody formation in the 100 ppm toxaphene group and enhanced antibody formation in the 200 ppm group. Phagocytic ability of macrophages was significantly reduced in the offspring at all doses. The reduction in phagocytic ability of macrophages was greatest at 100 ppm. The LOAEL for immunological effects in the offspring occurred when dams were fed 10 ppm (1.5 mg/kg/d) toxaphene. However, the actual dose received by the offspring cannot be estimated accurately because the offspring received toxaphene 16 transplacentally, in the milk and p o s s i b l y in the feed (EPA, 1985). Several studies have examined the reproductive and/or developmental effects of toxaphene. In a one generation reproduction and development study by Chu et al. (1988), groups of 30 female and 15 male Sprague-Dawley rats were e x posed to 0, 4, 20, 100 or 500 ppm toxaphene in the feed and m a t e d after 13 and 20 weeks of exposure. Dams were exposed to the treatment diet throughout mating and pregnancy. Fla pups were separated from the dams three weeks after birth and placed on the same treatment as their dams and FQ and Fla animals were exposed to toxaphene for a total of 26 weeks. Flb pups were killed 21 days post partum. Toxaphene had no effect on reproduction, litter size, pup weight, fertility, gestation or survival. Adult F0 female rats exhibited a significant increase in liver weight at 20, 100 and 500 ppm whereas F Q males and Fla males and females exhibited a significant increase in liver weights only at 500 ppm. Histological changes were observed in the liver, kidney and thyroid of the Fq and Fla male and female rats at 4 ppm but these changes did not appear to be b i o l o g i c a l l y significant. The changes in these organs at 20 ppm and above were more extensive and prevalent and were considered significant by the researchers. The NOAEL for male and female Sprague-Dawley rats in this study was 167 therefore 4.0 ppm. The geometric mean of the doses given in the study for males and females of the F0 and Fla groups at 4 ppm was calculated to be 0.35 mg/kg/d. In a three genera t i o n reproduction study by Kenn e d y et al. (1973), S prague-Dawle y rats were exposed to 25 or 100 'ppm toxaphene in the diet. The two test groups and two control groups each consisted of eight males and 16 females. Adult rats exposed to 100 ppm had slight cytoplasmic fatty vacuolization in their livers. No effects were observed at the 25 ppm dose level. Toxaphene had no effect on reproduction, litter size, pup survival or weanling body weights. No evidence of teratogenicity was found in this study. The NOAEL for systemic effects was 25 ppm, or 1.8 mg/kg bw/d assuming the rats consumed 7% of their body weight in food per day (EPA, 1988). Chernoff and Carver (1976) administered relatively high doses of toxaphene to pregnant CD rats in order to examine the effect of the chemical on fetal development. Toxaphene was administered in corn oil by gastric intubation on days 7 t h r o u g h 16 of gestation at doses of 15, 25 or 35 mg/kg/d. There was a reduction in weight gain of dams at all dose levels and in fetuses at 25 mg/kg/d, but there were no doserelated changes in fetal mortality or the occurrence of anomalies. In a study conducted by Kavlock et al. (1985), 168 supernumerary ribs were found in the neonates of pregnant mice treated with 100 and 120 mg/kg/d toxaphene on day 8 of gestation. However, this anomaly occurred at doses which were toxic to dams. The results of these studies suggest that teratogenic and fetotoxic effects occur only at doses which cause maternal toxicity. According to EPA (1985), "toxaphene appears to have a low teratogenic potential unless doses are large enough to induce maternal toxicity." The studies cited thus far suggest that toxaphene has little effect on neonate development. However, toxaphene has been found to affect the behavior of neonates at very low doses. In a study by Olson et al. (1980), pregnant albino rats were exposed to 50 ug/kg bw/d toxaphene via the feed beginning on day 5 of gestation. The neonates were exposed to the same treatment levels until the study was terminated at 3 months postpartum. Early development of the pups was tested on p o s t natal days 7 t h ro u g h 17 by m e a s u r i n g s w i m m i n g ability and righting reflexes. Motivational, learning and retention tests were run on postnatal days 70 through 90 using a symmetrical maze. Early in development, the treated animals exhibited retarded maturation as measured by their swimming ability, but by day 16 all groups displayed normal swimming ability. Toxaphene-fed animals exhibited a retarded righting reflex in tests conducted on postnatal days 7 16 9 through 17. Toxaphene treatment had no effect on the results of motivational tests. The Chu et al. (1988) study was considered the key study for risk assessment. This study used a sufficient number of doses to establish a NOEL for liver toxicity. The primary weakness of this study was that it was subchronic in duration. However, the results found in this study were consistent with the results of several chronic studies in rats. These chronic studies we r e not us e d for risk, assessment because they either established a NOEL which was higher than the LOEL determined in the Chu study (Lehman, 1952; Kennedy et al., 1973) or a NOEL was not established (Fitzhugh and Nelson, 1951). The Chu et al. (1988) study was also better suited for risk assessment because it used four dose levels whereas the chronic studies used only two dose levels. Although the subchronic monkey HNOAEL of 0.78 mg/kg/d (Hercules, Inc. as cited by EPA, 1980) was higher than the subchronic rat NOAEL OF 0.35 mg/kg/d (Chu et al., 1988), the latter study was preferred for risk assessment due to the greater extent of the dose-response data in rats. The lowest dose of toxaphene which was found to produce an adverse effect in rats was 50 ug/kg (Olson et al., 1980). This dose level was not used in the HNV derivation for toxaphene because the effects measured at this dose appeared 170 to be transient and the actual dose received by the neonates w a s uncertain. EPA (1985) did not use the O lson et al. (1980) study to derive health advisories for toxaphene for the same reasons cited above. The quality of the Chu et al. (1988) study was deemed sufficient to derive a Tier 1 HNC. The HNV was derived from the NOEL (0.35 mg/kg/d) determined for Sprague-Dawley rats using an uncertainty factor of 100 to account for intraspecies variability and interspecies extrapolation. An additional 10-fold uncertainty factor was not used for subchronic to chronic extrapolation because there were several chronic studies which supported the results of the Chu et al. (1988) study. ADE = 0.35 mg/ka/d = 0.0035 mg/kg/d = 3.5 ug/kg/d 100 Where: Uncertainty factor = 100, composed of: lOx for intraspecies variability lOx for interspecies extrapolation Drinking Water Sources: HNV = ADE x Uh x RSC 3.5 ua/lco/d x 70 ka x 0,8__________ VC + (FC x BAF) 2 l/d (0.015 kg/d x 2,117,450 l/kg") 171  0.0062 ug/l (rounded off to 6 ng/l (Tier 1)) Nondrinking Water Sources: HNV = APE Uti X RSC * 3.5 ug/kg/d x 70 kg x 0.B____________ WC (FC x 8AF) 0.01 t/d + (0.015 kg/d x 2,117,450 t/kg*) 0.0062 ug/l (rotnded off to 6 ng/L (Tier 1)) *BAF = 2,117,450 provided by EPA-Duluth and Minnesota PCA. REFERENCES: Allen, A.L., L.D. Roller and G.A. Pollock. 1983. Effect of toxaphene exposure on immune responses in mice. J. Toxicol. Environ. Hlth. 11:61-69. Chernoff, N. and B.D. Carver. 1976. Fetal t o xicity of toxaphene in rats and mice. Bull. Environ. Contam. Toxicol. 15:660-664. Chu, I., V. Secours, D.C. Villeneuve, V.E. Valli, A. Nakamura, D. Colin, D.J. Clegg and E.P. Arnold. R e p r o d u c t i o n study of t o x a phene in the rat. J. Environ. Sci. Hlth. B23(2): 101-126. 1988. Fitzhugh, O.G. and A.A. Nelson. 1951. Comparison of chronic effects produced in rats by several chlorinated hydrocarbon insecticides. Fed. Proc. 10:295. Kavlock. R . J . , N. Chernoff and E.H. Rogers. 1985. The effect of acute maternal toxicity on fetal development in the mouse. Terat. Careinog. Mutag. 5:3-13. Kennedy, G.L. Jr., J.P. Frawley, and J.C. Calandra. Multigeneration reproductive effects of three pesticides in rats. Toxicol. Appl. Pharmacol. 596. 1973. 25:589 172 Lehman, A.J. 1952. Oral toxicity of toxaphene. Q. Bull. Assoc. Food Drug Off. U.S. 16:47. Olson, K.L., F. Matsumura and G.M. Boush. 1980. Behavioral effects on juvenile rats from perinatal exposure to low levels of toxaphene, and its toxic components, toxicant A and toxicant B. Arch. Environ. Contam. Toxicol. 9:247-257. U.S. Environmental Protection Agency (EPA). 1980. Ambient Water Quality Criteria for Toxaphene. EPA 440/5-80076. U.S. Environmental Protection Agency (EPA). 1985. Drinking Water Criteria Document for Toxaphene. Environmental Criteria and Assessment Office. Cincinnati, OH. PB 86-118049. U.S. Environmental Protection Agency (EPA). 1987. Health Advisories for 16 Pesticides. Office of Drinking Water, Washington, D.C. NTIS No. PB 87-200176, 264 pp. U.S. Environmental Protection Agency (EPA). 1988. Recommendations for the Documentation of Biological Values for Use in Risk Assessment. NTIS No. PB 88 179874 . Tier l winn^Ti c a n c e r C r i t erion A c c o r d i n g to EPA (1985,- 1987), there are i nadequate data available to ascertain whether toxaphene is a human carcinogen. However, two chronic studies have shown that toxaphene induces the formation of liver tumors in B6C3F1 mice. One of these studies also found that toxaphene induces the formation of thyroid tumors in Osborne-Mendel rats. Toxaphene was mutagenic for Salmonella typhimurium strains TA98 and TA100 (Hill, 1977 as cited by EPA, 1985) and was also found to induce the formation of sister chromatid exchanges in Chinese hamster lung (DON) cells (Steinel et al., 1990). However, toxaphene produced a 17 3 negative response in a modified dominant lethal assay which used male ICR/Ha Swiss mice (Epstein et al., 1972). According to EPA (1985; 1987), the weight-of-evidence for toxaphene carcinogenicity is sufficient for B2 classification (probable human carcinogen). The data are sufficient to derive a Tier 1 HCC. In a study conducted by NCI (1979), 50 Osborne-Mendel rats/sex/group and 50 B6C3F1 mice/sex/group were administered toxaphene in their diets for 80 weeks. Male rats recei v e d t i m e - we i g h t ed average (TWA) doses of 112 and 556 ppm, while females received TWA doses of 540 and 1080 ppm. Both male and female mice received TWA doses of 99 and 198 ppm. Both rats and mice had 10 matched controls/sex w ith an additional 45 pooled controls/sex for rats and 40 additional pooled controls/sex for mice. Male and female rats exhibited a statistically significant dose-related increased incidence of thyroid tumors (adenomas and carcinomas), whereas treated mice exhibited a statistically significant dose-related increased incidence of liver tumors (NCI, 1979). In a study conducted by Litton Bionetics (1978, as cited by EPA, 1987), toxaphene was a d m i nistered to B6C3F1 m i c e (54 mice/sex/group) in the diet for 18 months at doses of 0, 7, 20 and 50 ppm. Animals were observed for a period of 6 174 months after treatment. An increased incidence of hepatocellular tumors (adenomas and carcinomas) was seen in both sexes and was statistically significant in males administered 50 ppm. This study, rather than the NCI study, was used for cancer risk assessment because it utilized more dose levels and a lower range of doses while still eliciting a tumorigenic response in the liver. EPA (1987) recommends the same key study and slope factor (1.1 (mg/kg/d)- 1 ) as utilized for HCC derivation. RAD = l x 10~5 = 9 x 10-6 mg/kg/d = 9 ng/kg/d -1 1.1 (mg/kg/d) Drinking Water Sources: HCV - RAD x Wh = 9 rw/kq/d x 70 lea_______________ UC + (FC * BAF) 2 l/d + (0.015 kg/d * 2,117,450 l/kg*) = 0.0198 ng/l (rounded off to 0.02 ng/l (Tier 1)) Nondrinking Water Sources: HCV s RAD x Wh = 9 ng/kg/d x 70 kg___________________ WC (FC x BAF) 0.01 l/d + (0.015 kg/d x 2,117,450 l/fcg*) = 0.0198 ng/l (romded off to 0.02 ng/l (Tier 1)) *BAF = 2,117,450, provided by EPA-Duluth and Minnesota PCA. 175 REFERENCES: Epstein, S.S., E. Arnold, J. Andrea, W. Bass and Y. Bishop. 1972. Detection of chemical mutagens by the dominant lethal assay in the mouse. Toxicol. Appl. Pharmacol. 23:288-325. Hill, R.N. 1977. Memorandum to Fred Hagemen. Off. Spec. Pestic. Rev., U.S. EPA. December 15. As cited in: EPA, 1984. Litton Bionetics. 1978. Carcinogenic evaluation in mice: Toxaphene. Prepared by Litton Bionetics, Inc., Kensington, MD for Hercules, I n c . , Wilmington, DE. National Cancer Institute (NCI). 1979. Bioassay of Toxaphene for Possible Carcinogenicity, Carcinogenesis Testing Program. Division of Cancer Cause and Prevention. NCI, National Institute of Health, Bethesda, Maryland, 20014. U.S. Department of Health, E d u c ation and Welfare. DHEW P u b l i c a t i o n No. (NIH) 7 9 837. Steinel, H.H., A. Ar l a u s k a s and R. S. Baker. 1990. SCE induction and cell-cycle delay by toxaphene. Mutat. Res. 230:29-33. U.S. Environmental Protection Agency (EPA). 1985. Drinking Water Criteria Document for Toxaphene. Environmental Criteria and Assessment Office. Cincinnati, OH. PB 86-118049. U.S. Environmental Protection Agency (EPA). 1987. Integrated Risk Information System (IRIS database). Chemical file for toxaphene (8001-35-2). Verification Date 3/5/87. Last Revised 1/1/91. 176 N o v e m b e r 4, 1991 GREAT LAKES INITIATIVE TIER 1 HUMAN HEALTH CRITERIA FOR TRICHLOROETHYLENE CAS NO. 79-01-6 Tier 1 Wiiman Noncancer Criterion A review of the available literature indicates inadequate human data for quantitative oral risk assessment of trichloroethylene (TCE). Humans exposed occupationally to TCE in air have been reported to experience central nervous system disorders such as dizziness, headaches, nausea, euphoria, palpitations, disturbances of vision, fatigue, and irritability,* disturbances in cardiac rhythm, gastrointestinal disorders, toxic effects on the liver, and cutaneous reactions. (NIOSH, 1973; Grandjean et al., 1955; IARC, 1979). A review of animal studies indicates that the most appropriate basis for HNV derivation for TCE is the NOAEL from a subchronic drinking water study conducted by Tucker et al. (1982). In this study, male and female CO-1 mice were administered TCE in drinking water ad libitum for 4 or 6 months. The concentrations used w e r e 0, 0.1, 1.0, 2.5, and 5.0 mg/mL, corresponding to time-weighted average daily 177 doses of 0, 18, 217, 393, and 660 m g / kg/day for male mice and 0, 18, 193, 437 and 793 mg / k g / d a y for female mice, respectively, according to the authors. Following 4 months of exposure, enlarged livers were observed in males at the 1, 2.5 and 5.0 mg/1 doses, and in females at the 5.0 mg/1 dose. A significant increase in kidney weight was seen at the highest dose in males at 6 months and in females at 4 and 6 months of exposure. The results of this study indicate a LOAEL value of 216 mg/kg/day and a NOAEL of 18 mg/kg/day based on liver effects in male CD-I mice. Sanders et al. (1982) investigated the immune status of CD-I mice using the same animals and bioassay as described above (Tucker et al., 1982). A s t a tistically significant (p < 0.05) finding of depressed cell-mediated immunity was observed in female mice at all four TCE concentrations after 4 months of exposure. However, this response at 4 months did not exhibit a clear dose-response relationship, and after 6 months of exposure the effect was statistically significant at only the highest dose. In assessing bone marrow function, the ability of bone marrow stem cells from the TCE exposed mice to form colonies in a semi-solid medium was found to be statistically significantly inhibited at all dose levels after 4 months and 6 months of exposure. While a statistically significant dose-response was observed, the biological significance of the in vitro observation is 178 questionable, This study may be interpreted to provide only suggestive evidence of a L O A E L at 18 m g / k g / d a y for immune system effects. The observations reported in the key study are supported by evidence of hepatotoxicity of TCE from gavage studies (Tucker et al., 1982; Buben and O'Flaherty, 1985; Zenick et al., 1984) and inhalation studies (Kjellstrand et al., 1983; 1982; 1981; Kimmerle and Eben, 1973). In a 14-day gavage study, Tucker et al. (1982) found a statistically significant increase in liver weights among male CD-l mice receiving TCE at 240 mg/kg, but not among mice receiving a dose of 24 mg/kg. In a subchronic study performed by Buben and 0'Flaherty (1985), TCE was administered by gavage in corn oil to male Swiss-Cox mice 5 days a w e e k for 6 weeks. D oses were 0, 100, 200, 400, 800, 1600, 2400, and 3200 mg/kg body weight. Doses of 100 mg/kg or more resulted in a statistically significant increase in the liver to body weight ratio when compared to controls. In addition, the authors noted a dose-response relationship in liver triglyceride levels and glucose-6-phosphatase activity. The increase in liver size was associated with hypertrophy of liver cells and a decrease in the DNA concentration in the liver. 179 In another subchronic study, Zenick et al, (1984) administered TCE in corn oil gavage to Long-Evans hooded male rats for 5 days per week for 6 weeks. Doses administered w e r e 0, 10, 100, and 1000 mg/kg/day. Elevated liver weight to body weight ratios were observed in the 100 and the 1000 mg/kg groups, and a reduced body weight gain was observed in the 1000 mg/kg group. Results of a 2-year gavage study (NTP, 1988) are not useful for HNV derivation because of the high doses used. Male and female ACI, August, Marshall, and Osborne-Mendel rats (50/sex/strain/group) were administered TCE in a corn oil v e hicle at doses of 0, 500, a nd 1000 mg/kg, 5 days per w e e k for 103 weeks. Renal tubular cell cytomegaly was observed in 82% to 100% of all dosed animals, and TCE produced toxic nephropathy in 17% to 80% of the dosed animals, but in none of the untreated or vehicle controls. No liver effects were reported. Reduced survival relative to vehicle controls was observed in 7/16 dosed groups, and nephrotoxicity, central nervous system toxicity characterized by sedation, loss of consciousness, tremors, and convulsions were observed in all dosed groups. EPA (1987a) developed a lifetime health advisory for TCE based on a rat inhalation study (Kimmerle and Eben, 1973). Subchronic exposure to TCE by adult rats at 55 ppm (300 180 m g / m 3 ) , 8 hours/day, 5 days per week for 14 w e e k s was followed by the investigation of indices of toxicity including hematological investigations, liver and renal function tests, blood glucose, and organ/body weight ratios. Liver weights were found to be elevated, while other test values were not different from controls. The elevated liver weights were attributed to hydropic changes or fatty accumulation. The LOAEL exposure is estimated to be equivalent to 270 mg/kg/day. The r e f e rence dose (RfD) determined by EPA (1987a) on the basis of this study was 0.00735 mg/kg/day. Other EPA documents that evaluate TCE t o x i c i t y (e.g. EPA, 1985; 1987b; 1988a) p r e s e n t cancer risk characterization, but do not recommend a key study or critical effect for noncancer assessment. A number of other inhalation studies corroborate the results of the Kimmerle and Eben (1973) study. Kjellstrand et al., (1983), exposed male and female NHRI mice (10/sex/group) to TCE at concentrations of 0, 37, 75, 150, a nd 300 p p m (0, 199, 403, 806, and 1612 mg/m3 ) for 30 days. The critical toxic effect observed was a change in liver weight, which was statistically greater in all exposed groups than in controls. Although a NOAEL was not identified by the authors, the LOAEL was determined to be 37 ppm (199 m g / m 3 ) for liver effects of TCE. This may be converted to an approximate administered dose of 338 mg/kg/day assuming that 181 mice breathe approximately 1.7 m 3/kg bw/day (EPA, 1 9 8 8 b ) . Kjellstrand et al. (1982) exposed NMRI mice, Sprague Dawley rats and mongolian gerbils to 150 ppm TCE for 30 days. A marked increase in liver weight was found in all three species. Using the same species and exposure, Kjellstrand et al. (1981) also observed increased kidney weights among TCE exposed gerbils, and to a lesser extent in mice and rats exposed to 150 ppm TCE for 30 days. The data are judged to be sufficient for Tier 1 HNC derivation. The HNC is based on the mouse N O A E L of 18 mg/kg/day determined in the key study (Tucker et al., 1982). There exists only suggestive evidence of effects on immune parameters at this dose (Sanders et al., 1982). Subchronic inhalation studies (Kimmerle and Eben, 1973; Kjellstrand et al., 1983) corroborate the sensitivity of liver effects, but involve shorter exposure durations than Tucker et al. (1982), fail to provide a NOAEL, and present uncertainties in route-to-route extrapolation for oral risk assessment. According to EPA (1985), TCE does not produce significant signs of developmental toxicity except at levels which affect maternal well being. Also no clinical evidence of fetotoxicity or teratogenicity from TCE exposure has been reported. Therefore, any potential for TCE to cause reproductive/developmental effects should be adequately protected by this approach to HNC development. 182 A D E = N O A E L = 18 mg/k q / d a v = 0.018 m g / k g / d a y UF 1,000 Where: Uncertainty factor = 1,000 composed of lOx for intraspecies variation lOx for interspecies extrapolation lOx for subchronic to chronic extrapolation Drinking Water Sources: HNV = APE x Wh x RSC = 0.018 m a / k g / d a y x 70 ka_________ WC + (FC x BAF) 2 1/d + (0.015 kg/d x 18 l/kg*> = 0.56 mg/1 (rounded off to 0.6 mg/1 (Tier 1)) Nondrinking Water Sources: H N V = APE x W h x RSC = 0.018 m a / k a / d a v x 70 ka____________ W C + (FC x BAF) 0.01 1/d + (0,015 k g / d x 18 l/kg*> = 4.5 mg/1 (rounded off to 5 mg/1 (Tier 1)) Where: *BAF = 18 provided by EPA-P u l u t h and M i n n e s o t a PCA. ia 3 NOTE: A R e l a t i v e Source C o n t r ibution (RSC) factor has not been utilized in these draft calculations. REFERENCES: Buben, J. and E. O'Flaherty. 1985. D e l i n e a t i o n of the role of metabolism in the hepatotoxicity of trichloroethylene and perchloroethylene: A dose effect study. Toxicol. Appl. Pharmacol. 78:105-122. Grandjean, E . , et al. 1955. Investigations into the effects of exposure to trichloroethylene in mechanical engineering. Br. J. Indust. Med. 12:131-142. International Agency for Research on Cancer (IARC). 1979. IARC Monographs on the Evaluation of Carcinogenic Risk of Chemicals to Man. WHO Publ. Centre, USA. Volume 20. Albany, NY. Kimmerle, G. and A. Eben. 1973. Metabolism, e x c r etion and t o x i c o l o g y of t r i c h l o r o e t hy l e n e after inhalation. 1. Experimental exposure on rats. Arch. Toxicol. 30:115. Kjellstrand, P., B. Holmquist, P. Aim, M. Knaje, S. Romare, I. Jonsson, L. Mansson, and M. Bjerkemo. 19B3. Trichloroethylene: Further studies of the effects on body and organ weights and plasma butyrylcholinesterase activity in mice. Acta Pharmacol, et Toxicol. 53:375 384. Kjellstrand, P . , A. Edstrom, M. Bjerkemo, and B. Holmquist. 1982. Effects of trichloroethylene inhalation on acid phosphatase in rodent brain. Toxic. Let. 10:1-5. Kjellstrand, P . , M. Kanje, L. Mansson, M. Bjerkemo, I. Mortensen, J. Lanke, and B. Holmquist. 1981. Trichloroethylene: Effects on body and organ weights mice, rats and gerbils. Toxicol. 21:105-115. in National Institute of Occupational Safety and Health (NIOSH). 1973. Criteria for a Recommended Standard for O c c u p a t i o n a l Exposure to Trichloroethylene. R e p o r t No. NIOSH-TR-043-73. DHEW. Public Health Serv. Center for D i s e a s e Control. Cincinnati, OH. 184 National Toxicology Program (NTP). 1988- Toxicology and Carcinogenesis Studies of Trichloroethylene in Four Strains of Rats. National Toxicology Program. Technical Report Series No. 273. U.S. Department of Health and Human Services. Public Health Service, National Institutes of Health. Sanders, V., A. Tucker, K. White, B. Kauffmann, P. Hallett, R. Carchman, J. Borzelleca, and A. Munson. 1982. Humoral and cell-mediated immune status in mice exposed to trichloroethylene in drinking water. Toxicol. Appl. Pharmacol. 62:358-368. Tucker, A.N., V. Sanders, D. Barnes, T. Bradshaw, K. White, L. Sain, J. Borzelleca, and A. Munson. 1982. Toxicology of trichloroethylene in the mouse. Toxicol. Appl. Pharmacol. 62:351-357. U.S. Environmental Protection Agency (EPA). 1988a. Health Effects Assessment for Trichloroethylene. EPA/600/889/097 ._ Washington, D.C. U.S. Environmental Protection Agency (EPA). 1988b. Recommendations for and Documentation of Biological Values for use in Risk Assessment. PB88-179874. U.S. Environmental Protection Agency (EPA). 1987a. Trichloroethylene Health Advisory. Office of Drinking Water. U.S. Environmental Protection Agency (EPA). 1987b. Addendum to the Health Assessment Document for Trichloroethylene: Updated Carcinogenicity Assessment for Trichloroethylene. EPA/600/8-82/006. Washington, D.C. U.S. Environmental Protection Agency (EPA). 1985. Health Assessment Document for Trichloroethylene. EPA/600/882/006F Washington, D.C. Zenick, H., K. Blackburn, E. Hope, N. Richdale, M. Smith. 1984. Effects of trichloroethylene exposure on male reprod u c t i v e function in rats. Toxicol. 31: 237-250. Tier 1 m m " Cancer Criterion Six epidemiologic studies have been performed to investigate the carci n o g e n i ci t y of tr i c h l o r o e t hy l e n e (TCE) in exposed ias workers (Axelson et al., 1978; Hardell et al., 1981; Malek et al., 1979; Novotna et al., 1979; Paddle, 1983; Tola et al, 1980). Results of those studies we r e inadequate to attribute cancer incidence to TOE exposure. However, because they suffer from various limitations and deficiencies, they- also fail to provide adequate evidence that TCE is not a human c arcinogen (EPA, 1985). Based on weight of evidence, EPA (1985, 1987, 1988) classified TCE in Group B2- Probable Human Carcinogen. The evidence reviewed by EPA (1985) for carcinogenicity of TCE in experimental animals includes increased incidence of hepatocellular carcinomas in male and female B6C3F1 mice (NCI, 1976; NTP, 1982, 1986) by gavage, m a l i g n a n t lymphomas in female Han:NMRI mice by inhalation (Henschler et al., 1980); and renal adenocarcinomas in male Fischer 344 rats by g a v a g e (NTP, 1982, 1986). Evidence presented in EPA (1987) markedly strengthened the B2 classification by showing that inhalation is a second exposure route that results in carcinogenic activity in rats and mice, and by identifying d i v e r s e tumor sites (EPA, 1987). EPA (1985) developed a quantitative cancer risk assessment based on four sets of gavage bioassay data that show h e p a t o c e l l u la r carcinomas in male and female mi c e (NTP, 186 1982; NCI, 1976), The NCI bioassay involved exposure by gavage to B6C3F1 mice. Although rats were also tested, excessive mortality in all groups cast doubt on the adequacy of those results. Mice were dosed in groups of 50 animals per sex, 5 days/week for 78 weeks. Surviving animals were sacrificed at 90 weeks and subjected to complete necropsy and histopathological examination. The time-weightedaverage (TWA) doses for male mice w e r e 1,169 and 2,339 mg/kg, and for the female mice they were 869 and 1,739 mg/kg. The study included 20 matched vehicle control animals of each sex. It was concluded that TCE induced a s t a t i stically significant (p < 0.05) increase in the incidence of hepatocellular carcinoma in both male and female B6C3F1 mice. A reduction in the time-to-tumor response was also reported among male mice at the high dose level. The presence of the trace contaminant epichlorohydrin (0.09%) in the test material for this bioassay could be a cause for concern. However, it has been determined that any potential contribution of epichlorohydrin to the overall carcinogenic potency of TCE in the b i o assay was negligible (EPA, 1985). NTP (1982) conducted a carcinogenicity bioassay on TCE in B6C3F1 mice and F344/N rats. The rats experienced reduced survival when compared to controls, and the results were therefore invalidated. Male and female mice were dosed by 187 gavage at 1,000 mg/kg, 5 days/week for 103 weeks. Survival was significantly lower (p < 0.004) in treated males whereas survival in treated females was lower after 95 weeks, but the overall difference between vehicle controls and treated females was not significant. Male and female mice had a statistically significant increase in the incidence of hepat o c e l l u la r carcinoma (p < 0.002) and h e p a t o c e l l u la r adenoma (p < 0.05) over c o r responding vehicle controls. The TCE test material for that bioassay was not contaminated with detectable amounts of epichlorohydrin. The potency of TCE with regard to the induction of hepatocellular carcinomas in mice has been determined to be very similar in the NTP (1982) bioas s a y and t he NCI (1976) bio a s s a y (EPA, 1985). Additional studies were reviewed by EPA (1987) identifying positive findings by inhalation exposure in rats and mice. Maltoni et al. (1986) conducted bioassays of Sprague-Dawley rats e x posed to 0, 100, 300 and 600 p p m of TCE 7 hours/day, 5 days/week for 104 weeks. Necropsy was performed on all animals. Male rats demonstrated increased incidences of renal tubuli megalonucleocytosis and renal adenocarcinomas, and a slight increase in leukemias, particularly immunoblastic lymphosarcomas. Maltoni et al. (1986) also c o n d ucted bioassays on Swiss mice and B6C3F1 m i c e (90 mice/strain/sex/group) exposed to 0, 100, 300, and 600 ppm 188 TCE for 78 weeks. Statistically significant increases in hepatomas were noted among male Swiss mice at the high concentration, and significant increases in pulmonary tumors were observed among male Swiss mice at high and medium exposures. Among the B6C3F1 mice, there were increases in hepatomas in males and females, pulmonary tumors in females, and in the total number of tumors among females at all concentrations. Fukuda et al., (1983) reported the results of bioassays with female ICR mice and Sprague-Dawley rats (49-50 per group) e x posed to airborne concentrations of 0, 50, 150, and 450 ppm of TCE for 7 hours/day, 5 days/week for 107 weeks. There were no statistically significant increases in tumors among rats, however a statistically significant increase in lung adenocarcinomas was found among the mice. Using the mice liver tumor data sets from NTP (1982) and NCI (1976), EPA (1985) calculated human slope estimates of 1.9 xlO- 2 , 8.0 x 10"3 , 1.8 x 10"2 , and 5.8 x 10"3 per mg/kg/day. Because the slope estimates from these four data sets were found to be comparable, their geometric mean was used to derive the recommended slope factor of 1.1 x 10~2/ (mg/kg/day). EPA (1987) also developed slope factors from the inhalation studies of Maltoni et al. (1986) and Fukuda et al.(1983). These slope factors were found to be 189 comparable to the studies (EPA, 1987). developed earlier from the gavage RAD = 1 x 1 0 ~ 5________________ = 9.091 x 10" 4 mg/ k g / d 1.1 x 10"2 (mg/kg/d)"1 Drinking Water Sources: HCV = RAD x Wh_______ = 9.091 x 10~4 ma/ k a / d x 70 ka WC + (FC x BAF) 2 1/d + (0.015 kg/d x 18 1/kg*) = 0.028 mg/1 (rounded off to 0.03 m g / 1 (Tier 1)) Nondrinking Water Sources: H C V = R AD X Wh________ = 9.091 x _1CT4 m q / k q / d X 7Q_Jc_q_______ _ WC + (FC X BAF) 0.01 1/d + (0.015 kg/d x 18 1/kg*) = 0.227 mg/1 (rounded off to 0.2 mg/1 (Tier 1)) Where: *BAF = 18 Provided by EPA-Duluth and M i n n e s o t a PCA. 19 0 REFERENCES: Axelson, 0. et al. 1978. A cohort study on trichloroethylene exposure and cancer mortality. Occup. Med. 20:194-196. J. Fukuda, K . , K. Takemoto, H. Tsuruta. 1983. Inhalation carcinogenicity of trichloroethylene in mice and rats. Ind. Health 21:243-254. Hardell, L . , et al. 1981. Malignant lymphomas and exposure to chemicals, especially organic solvents, chlorophenols, and phenoxy acids: a case-control study. Br. J. Cancer. 43:169-176. Henshler, L. et al. 1980. Carc i n o g e n i ci t y study of trichloroethylene by long-term inhalation in the animal species. Arch. Toxicol. 43:237-248. Malek, B., B. Kromarova, and 0. Rodova. 1979. An epidemiological study of hepatic tumor incidence in subjects working with trichloroethylene. II. Negative results of retrospective investigations in drycleaners. Prakov. Lek. 31: 124-126. As cited in EPA (1985). Maltoni, C., G. Lefemine, and C. Cotti. 1986. Experimental research on trichl o r o e t hy l e n e carcinogenesis. In: Maltoni, C. M. M e l h a m eds. Arc h i v e s of R esearch on Industrial Carcinogenesis. Vol. V. Pr i n c e t o n NJ. Princeton Scientific Publishing Co. National Cancer Institute (NCI). 1976. Carcinogenesis Bioassay of Trichloroethylene. CAS No. 79-01-6. NCI- CG-TR-2. National Toxicology Program (NTP). 1982. Carcinogenesis Bioassay of Trichloroethylene. Cas No 79-01-6. NTP 81-84. NIH Publication No. 82-1799. National Toxicology Program (NTP). 1986. Toxicology and Carcinogenesis Studies of Trichloroethylene in F344/N Rats and B6C3F1 Mice. NTP TR 243. U.S. Department of Health and Human Services. National Institutes of Health. Bethesda, MD. Novotna, E., A. David, and B, Malek. 1979. An epidemiological study on hepatic tumor incidence in subjects w o r k i n g with trichloroethylene: I. Negative results of retrospective investigations in subjects 191 with primary liver carcinoma. Pracovni Lekartsvi. 31(4): 121-123. As cited in EPA (1985). Paddle, G. 1983. Incidence of liver cancer and trichloroethylene manufacture: joint study by industry and cancer registry. British Medical Journal. 286:846 Tola, s., R. Vilhnuer, E. Jaruinene, and M. Korkale. 1980. A cohort study of workers exposed to trichloroethylene J. Occup. Med. 22:737-740. U.S. Environmental Protection Agency (EPA). 1988. Health Effects Assessment for Trichloroethylene. EPA/600/889/097. Washington, D.c. U.S. Environmental Protection Agency (EPA). 1987. Addendum to the Health Assessment Document for Trichloroethylene: Updated Carcinogenicity Assessment for Trichloroethylene. EPA/600/8-82/006. Washington, D.C. U.S. Environmental Protection Agency (EPA). 1985. Health Assessment Document for Trichloroethylene. EPA/600/882/006F Washington, D.C. 192