Shaw and colleagues investigated the association between tap water consumption and cardiac defects ( Shaw et al., 1990). They interviewed 145 mothers of children with severe congenital cardiac disease and 176 mothers of children with no such disease. Subjects were asked about their consumption of drinking water from tap and bottled sources during their first trimester of pregnancy. A relationship was found between maternal consumption of 4 or more glasses of tap water per day and increases in cardiac abnormality in the infants ( OR 2.0, 95% CI 1.0 4.0). A follow up study was conducted by Shaw et al. ( 1991) to assess the relationship between water chlorination and cardiac defects. In this study, the source of drinking water during the first trimester was determined for each of the 138 study participants. THM concentrations 36 corresponding to the first trimester of pregnancy were estimated from quarterly utility monitoring records. No relationship between cardiac abnormalities and chlorinated drinking water was identified, and average THM levels during the first trimester of pregnancy for cases ( 64.0 µ g/ L)

was less than for controls ( 74.2 µ g/ L).

Kramer et al. ( 1992) conducted a population based case control study, based on an examination of birth certificates from January 1, 1989 to June 30, 1990 for live infants born to white women in Iowa towns with 1,000 to 5,000 inhabitants. This study reported an increase in intrauterine growth retardation with exposure to chloroform concentrations in drinking water of greater than or equal to 10 µ g/ L ( OR 1.8, 95% CI 1.1   2.9). The authors also found a nonsignificant increase in intrauterine growth retardation associated with exposure to drinking water with BDCM concentrations greater than or equal to 10 µ g/ L compared with drinking water with undetectable BDCM concentrations ( OR1.7, 95% CI 0.9   2.9). THM exposures, including chloroform and BDCM, were estimated from a water supply survey conducted two to three years previously.

King et al. ( 2000a) conducted a retrospective cohort study of 50,000 deliveries and reported that exposure to TTHMs, chloroform and BDCM was associated with an increased incidence of stillbirth. Exposure was estimated by linking the mother's residence at the delivery date to the levels of THMs monitored in public water supplies, using the predicted average exposure levels for the entire duration of the pregnancy. For chloroform, the adjusted odds ratio for stillbirth was increased for exposure  100 µ g./ L ( OR 1.56, 95% CI 1.04 2.34). Risk doubled for women exposed to a BDCM level of greater than or equal to 20 µ g/ L, when compared to women consuming concentrations of less than 5 µ g/ L. When categories of stillbirth ( unexplained 37 deaths and asphyxia related deaths) were examined, relative risk estimates for asphyxia related deaths increased by 32% for each 10  g/ L increase in concentration of BDCM. As indicated by King et al. ( 2002a), an in vitro study ( Alston 1991) suggested that chloroform and related compounds might be linked to the asphyxia related deaths by influencing the methioninehomocysteine metabolic pathway and subsequent abruption of placenta. These data support the asphyxia related deaths observed by King et al. ( 2002a), and indicate that the observed stillbirths were related to BDCM, despite their limitation to one type of stillbirth. A limitation of this study is possible misclassification of exposure as a result of mobility within the study population.

Windham et al. ( 2003) examined menstrual cycle characteristics in relation to the presence of brominated trihalomethanes in tap water in a prospective study of women living in Northern California. Data were also reported for TTHMs and for chloroform. The target population was married women of reproductive age ( 18 39 years old). Participants were selected from among nearly 6500 women using a short screening interview to identify women who were more likely to become pregnant ( i. e., those who reported a menstrual period within the last six weeks, were not surgically sterilized, did not use birth control pills or intra uterine devices, and were noncontracepting for less than 3 months). Out of 1092 eligible women, a total of 403 women finished the study. These participants collected first morning urine samples daily for 2 9 menstrual cycles ( average 5.6 cycles) for measurement of steroid metabolites. The participants filled out a daily diary during the urine collection phase and recorded vaginal bleeding. These measurements ( diary and urinary hormone metabolites) were used to estimate menstrual parameters such as cycle and phase length. Cycle length was calculated from the first day of menses to the day before onset of the next menses. When the available data permitted, the cycle was divided into the follicular 38 phase ( first day of menses through estimated day of ovulation) and the subsequent luteal phase.

Between 1424 and 1714 cycles were available for evaluation of each parameter. Information on water consumption and other variables ( age, race, education, employment, income, pregnancy history, exercise type and frequency, smoking, alcohol and caffeine consumption) was collected in a baseline telephone interview prior to urine collection. Information on the number of showers taken at home per week and their duration was also collected. Exposure to trihalomethanes was estimated from quarterly utility monitoring data and information on drinking water and other tap water use collected during the baseline interview.

A monotonic decrease in mean cycle length was observed with increasing TTHM exposure category. At TTHM concentrations greater than 60 µ g/ L, the adjusted decrement in cycle length was 1.1 day ( 95% C. I.  1.8,  0.40) when compared to TTHM concentrations of 40 µ g/ L or less. The decrease in follicular phase length was similar ( 0.94 day; 95% C. I.  1.6,  0.24).

A unit decrement in mean cycle and follicular phase length of 0.18 days per 10 µ g/ L increase in TTHM concentration ( 95% C. I.  0.29,  0.07) was determined when the cycle specific TTHM level was examined as a continuous variable. Examination of time spent showering did not reveal additional risks with longer showers. Combined with TTHM concentration, decrements in cycle and follicular phase length were seen at the higher TTHM (> 60 µ g/ L) and longer showers (  70 minutes) categories ( 1.2 and   1.6 days respectively). However, the confidence intervals were wide for all duration categories and a clear dose response pattern ( i. e., shorter lengths at higher durations) was not evident.

This study suggests that THM exposure may have effect on menstrual cycle length.

However, examination of time spent showering did not reveal additional risks with longer 39 showers. This is counter to the expected trend, as elevated blood levels of THMs have been documented after showering, due to dermal and inhalation exposure in the shower. However,

information on shower duration was collected by interview and the reported lengths may not have accurately reflected actual shower duration. It would also be useful to have an independent confirmation of these results in another study.

The weight of evidence from epidemiology studies suggests that DBPs are associated with developmental/ reproductive effects under certain exposure conditions. The existing data are relatively sparse, and are insufficient for dose response analysis. There are inconsistencies among the available studies on the association between drinking water disinfection and specific effects,

such as changes in menstrual cycle length, fetal growth, fetal viability, and congenital abnormalities. The studies employing the best exposure assessment found the strongest association between fetal growth and fetal viability and chlorinated drinking water ( Bove et al.,

2002). Evaluation of the studies is made more uncertain by the difficulties of exposure assessment, including the different composition of D/ DBPs at different locations, and the variation in composition over time at a single location. The specific DBP( s) responsible for reproductive and/ or developmental effects is not known. The existing data suggest a relationship between THMs, particularly BDCM, and developmental effects, but this may be related to a tendency among investigators to report concentrations of THMs, and not the concentrations of other DBPs. The available data suggest that additional studies on THMs and other DBPs are needed.

2.1.2. Systemic Effects 40 There are no epidemiology studies addressing whether fetuses or children are more sensitive than adults to systemic effects of chlorinated drinking water.

2.1.3. Carcinogenicity The epidemiologic evidence regarding a relationship between drinking water chlorination and cancer in adults is mixed. An association between chlorinated drinking water and rectal,

colon, kidney, and/ or bladder cancers has been reported based on a number of human studies.

The association is strongest between DBPs and bladder cancer, but causality has not been established. Koivusalo et al. ( 1998) and Cantor et al. ( 1998) both found evidence of increased risk with increasing exposure duration, but the increase was only approximately two fold after a long exposure duration. An intermediate level of evidence supports an association with rectal cancer. Yang et al. ( 1998) and Hildesheim et al. ( 1998) both found associations between chlorinated drinking water exposure and rectal cancer, and the associations had a similar magnitude in both sexes. Hildesheim et al. ( 1998) also found an association in both sexes with lifetime average trihalomethanes ( THMs) concentration. The consistency of the dose response trends, the consistency between sexes, and the apparent control of important potential confounders in this study suggest that the observed associations between the exposures and rectal cancer may be real ( EPA, 2001e). Only one ( King et al., 2000b) of three key recent studies ( Hildesheim et al., 1998; King et al., 2000b; Yang et al., 1998) found an association with colon cancer, and a strong causal association between DBPs and colon cancer is considered unlikely ( EPA, 2001e). The limited data on kidney cancer support the possibility of an association with DBPs. Yang et al. ( 1998) found fairly high standardized rate ratios for kidney cancer, and 41 Koivusalo et al. ( 1998) observed a dose response by mutagenicity tertiles and by duration of exposure. Limited human data ( Cantor et al., 1999) also suggest an association between chlorinated drinking water and gastrointestinal/ urinary tract cancers and brain cancers ( gliomas).

However, the data are insufficient for a definitive conclusion, and are insufficient for quantitative cancer risk assessment.

Only one study specifically evaluated the relationship between childhood cancers and chlorinated drinking water. In a population based case control study in Quebec, Infante Rivard et al. ( 2000) examined the possible association between childhood acute lymphoblastic leukemia ( ALL) and THMs. The authors studied 491 cases and 491 controls, matched for age, sex, and region in the province; logistic regression analysis adjusted for maternal age and level of schooling. Individual information was collected on water source, and exposure was estimated from distribution system data for metals, nitrates, and THMs. No association with ALL was found for prenatal or postnatal exposure to total THMs or for specific THMs, and there was no evidence of a dose response. Based on these studies, the data are inadequate to evaluate the potential carcinogenic effects of childhood or prenatal exposure to chlorinated drinking water.

Overall, the epidemiology data are inadequate for a definitive conclusion regarding the carcinogenic potential of chlorinated drinking water for adults or children.

2.2. TRIHALOMETHANES 2.2.1. Chloroform 42 Chloroform is one of the best studied DBPs and has a very extensive toxicological database. Chloroform and its metabolites have been shown to cause liver and kidney toxicity and tumors as their primary adverse effects.

<IP_ADDRESS>. Developmental/ Reproductive Effects In a prospective study, Windham et al. ( 2003) demonstrated that increasing levels of TTHMs were associated with significantly shorter cycles when examined by quartile. Similar decrements were observed in follicular, but not luteal, phase length. However, the effect was attributed to exposure to brominated THMs, and a clear association with reduced cycle length was not observed for chloroform even at the highest quartile (  17 µ g/ L) ( difference  0.3 days;

95% C. I.  1.0, 0.40).

Two epidemiology studies ( Kramer et al., 1992; King et al., 2000a) investigated the relationship between exposure to chloroform in chlorinated drinking water and developmental effects. These studies were discussed in Section 2.1., Chlorinated Drinking Water.

Three developmental toxicity studies ( two in rats and one in rabbits) by the oral route of administration ( Thompson et al., 1974; Ruddick et al., 1983) and three developmental toxicity studies ( two in rats and one in mice) by the inhalation route of administration ( Schwetz et al.,

1974; Murray et al., 1979) were reported.

Thompson et al. ( 1974) studied the effects of chloroform on embryonic and fetal development of Sprague Dawley rats. Groups of 25 pregnant rats ( 181   224 g) were gavaged with chloroform in corn oil at total daily doses of 0, 20, 50, or 126 mg/ kg/ day by oral intubation on days 6   15 of gestation, administered in two doses/ day. Dams receiving 50 or 126 mg/ kg/ day 43 displayed signs of maternal toxicity ( decreased weight gain, mild fatty changes in the liver). There was no evidence of maternal toxicity at 20 mg/ kg/ day, although microscopic examinations were conducted on only 2 dams/ group. Fetuses were removed by caesarean section 1 or 2 days prior to expected parturition and examined for external, skeletal and/ or soft tissue abnormalities. There were no fetal malformations, but some fetal variations were noted as reflected in the statistically significant increase in the incidence of bilateral extra lumbar ribs ( p< 0.05) at the high dose;

however, the increase in affected litters was not statistically significant. Fetal weight was also reduced at the high dose ( p< 0.05). This study identified a maternal NOAEL of 20 mg/ kg/ day and a LOAEL of 50 mg/ kg/ day in rats. For developmental effects, the NOAEL was 50 mg/ kg/ day,

with a LOAEL of 126 mg/ kg/ day.

In the same study, Thompson et al. ( 1974) administered chloroform ( in corn oil) to Dutch Belted rabbits. In a preliminary range finding study, doses of 0, 25, 63, 100, 159, 251, or 398 mg/ kg/ day were administered to pregnant rabbits on days 6   18 of gestation. High levels of maternal death ( 60   100%) were observed at doses of 100 mg/ kg/ day and above. Adverse effects at 63 mg/ kg/ day included anorexia, weight loss, diarrhea, abortion, and one maternal death. No overt signs of toxicity other than mild diarrhea and intermittent anorexia were observed in dams dosed with 25 mg/ kg/ day. In the main study, groups of 15 pregnant dams ( 1.7   2.2 kg) were dosed by oral intubation with chloroform at 0, 20, 35, or 50 mg/ kg/ day on days 6   18 of gestation.

Decreased maternal weight gain was observed in dams given 50 mg/ kg/ day. Four high dose dams died from hepatotoxicity, but no evidence of hepatotoxicity was observed in surviving rabbits.

Four high dose dams aborted, but this was not considered to be a treatment related effect as three control animals also aborted. Histopathology examinations revealed no evidence of maternal 44 toxicity at 35 mg/ kg/ day. Small reductions in body weights ( 7.5% and 12%, respectively) were observed in fetuses from dams administered 20 or 50 mg/ kg/ day ( p< 0.05), whereas only a 5.5%

decrease in fetal weight was observed at 35 mg/ kg/ day. At least some of the decrease in fetal weight at the high dose may be attributable to the larger litter size at the high dose ( 7.4, vs. 6.4 in the controls), although the mean litter size at the mid dose was only 4.5. An increased incidence of fetuses with incompletely ossified skull bones ( usually parietals) was observed at 20 and 35 mg/ kg/ day ( p< 0.05); a smaller increase at the high dose was not statistically significant, and the results were not significant when the litter was used as the statistical unit of comparison. This study is limited by the high incidence of abortions and mortality in the control group, but the study authors did not consider the observed effects to be evidence of teratogenicity or fetotoxicity. This study identified a maternal NOAEL of 35 mg/ kg/ day and a maternal LOAEL of 50 mg/ kg/ day,

based on hepatotoxicity; there were no developmental effects related to chloroform treatment.

Ruddick et al. ( 1983) investigated the developmental toxicity of chloroform in groups of 15 mated Sprague Dawley rats. Pregnant dams ( 8   14 animals per dose group) were given 0, 100,

200, or 400 mg/ kg chloroform in corn oil on days 6   15 of gestation. Maternal weight gain was depressed by at least 20% at all dose levels. In addition, all dose levels of chloroform produced maternal liver enlargement, decreased hemoglobin, and decreased hematocrit. Levels of serum inorganic phosphorus and cholesterol were elevated in the dams at the highest exposure level.

Fetal weight was decreased by about 19% at the highest dose level. There were no fetal malformations, but sternebra aberrations were observed with a dose dependent incidence at 200 mg/ kg/ day and 400 mg/ kg/ day. Interparietal deviations also occurred at the high dose. There 45 was a clear increase in the incidence of these variations indicating a potential developmental effect.

In an inhalation study, Schwetz et al. ( 1974) exposed pregnant female Sprague Dawley rats to chloroform at target concentrations of 0, 30, 100 or 300 ppm ( actual concentrations of 0,

30, 95, or 291 ppm; 0, 146, 464 or 1,420 mg/ m3) for 7 hours/ day from gestation days 6 through 15. Because marked anorexia was observed in an earlier experiment in dams exposed to 300 ppm chloroform, an additional control group was starved; that is, allowed only 3.7 grams of food per day. The numbers of pregnant rats exposed in each group were 68, 22, 23, and 3, respectively.

The low percent pregnancy observed at the high concentration was not considered to be treatment related, due to the time of exposure; however, the use of such a small number of animals in the 300 ppm group decreased the statistical sensitivity of any adverse effects observed in this group. The dams were sacrificed on gestation day 21, and fetuses were removed by caesarian section. Food consumption and body weight gain exhibited concentration related decreases in all exposure groups. A significant increase in relative liver weights in dams exposed to 100 or 300 ppm was observed at study termination, with a significant decrease in absolute liver weight at 300 ppm. However, there was no effect on serum glutamate pyruvate transaminase activity at any concentration. At 300 ppm, 61% of the implantations were resorbed, a statistically significant increase. This high resorption rate was not observed in the " starved" control group,

suggesting that weight loss cannot account for the observed effect, although the starved control group was provided more food than was consumed by the 300 ppm group. Fetal body weights were significantly decreased ( 40%) at 300 ppm, and fetal crown rump lengths were slightly, but significantly, decreased ( 2%) at 30 ppm and significantly decreased ( 15%) at 300 ppm. The 46 frequencies of litters with acaudia or imperforate anus were significantly increased at 100 ppm.

Malformations were not observed at 300 ppm, but there were only three litters at this concentration. The frequency of litters with delayed ossification was elevated in all exposure groups. In addition, there were statistically significant increases in wavy ribs at 30 ppm, and in missing ribs and subcutaneous edema at 100 ppm. The authors concluded that chloroform exposures of 100 and 300 ppm were highly embryotoxic and fetotoxic, with embryolethality a significant effect at 300 ppm.

Murray et al. ( 1979) found that 100 ppm chloroform was teratogenic in CF 1 mice exposed on gestation days 8   15, but was fetotoxic in mice exposed on gestation days 1   7 or 6   15. Groups of 34   40 pregnant females ( as determined by vaginal plug) were exposed to 0 or 100 ppm ( 0 or 490 mg/ m3) for 7 hours/ day on gestation days 1   7, 6   15, or 8   15, and sacrificed on gestation day 18. The ability of the mice to maintain pregnancy was significantly decreased in the groups exposed on gestation days 1   7 or 6   15, and there was a slight ( but not statistically significant) decrease in pregnancies in the group exposed on gestation days 8   15. Statistically significant decreases in fetal weight and fetal length were observed in the groups exposed on gestation days 1   7 and 8   15 but not on days 6   15. Cleft palate was observed at a statistically significant increased incidence in litters of mice exposed on gestation days 8   15, but not in the other groups. Cleft palate was seen predominantly in fetuses with retarded growth. No other malformations were significantly increased in any group, although increased incidences of two skeletal variations were observed. Delayed ossification of skull bones was significantly increased in all exposed groups, and delayed ossification of sternebrae was significantly increased in the groups exposed on gestation days 1   7 and 8   15, but not 6   15. The study authors suggested that 47 the lack of malformations in the group exposed on gestation days 6   15 may have resulted from the lethality to the early embryo obscuring other effects. Maternal toxicity was evident as increased liver weight and increased serum glutamate pyruvate transaminase activity.

Based on the findings of animal studies discussed above, developmental effects have been found after chloroform exposure ( e. g., pup weight reduction, skeletal variations). These prenatal effects, however, were typically associated with exposures causing maternal toxicity, and occurred at oral doses above those causing hepatotoxicity. The oral NOAEL for developmental toxicity is in the range of 35   50 mg/ kg/ day, and the oral LOAEL for hepatotoxicity is 12.9 mg/ kg/ day for chloroform.

A multigeneration reproductive assay was conducted with chloroform in CD 1 mice ( NTP, 1988). This assay evaluated reproductive effects in two successive generations, as well as systemic effects in the second generation ( i. e., F 1 animals). In the first phase, mice were administered chloroform by gavage in corn oil at 6.6, 16, or 41 mg/ kg/ day, 7 days/ week for 18 weeks. In the second phase, the last litter of the control and of the high dose groups were retained. After weaning, the mice were administered the same chloroform dose as their parents,

and dosing continued through mating and parturition, when the study was terminated. No adverse effects on fertility or reproduction of the F 1 generation were observed, although increased liver weight and liver lesions ( degeneration of centrilobular hepatocytes, accompanied by occasional single cell necrosis) were observed in all females exposed to the single dose tested.

The degeneration was characterized as minimal in 2/ 20, mild in 9/ 20 and moderate in 9/ 20 animals. Thus, a dose of 41 mg/ kg/ day caused mild to moderate liver histopathology in F 1 females. No NOAEL can be identified for this effect, because the low and mid dose groups were 48 not evaluated histopathologically. However, no adverse effects on fertility or reproduction were found.

<IP_ADDRESS>. Systemic Effects Numerous animal studies in several species ( rats, mice and dogs) have shown that liver and kidney toxicity are primarily target sites for the systemic effects of chloroform. Nasal toxicity is also found in the rat following inhalation exposure. Organ toxicity ( including liver and kidney tumor response) following chloroform treatment vary with the exposure route, vehicle of administration and strain of rat or mouse. The sensitivity to the organ toxicity induced by chloroform is associated with oxidative metabolism. These results were summarized in two EPA documents ( EPA, 1994b, 1998b). Organ toxicity that results from chloroform is considered to be part of the continuum that leads to tumor development. The organ toxicity is thus discussed below in the context of the mode of carcinogenic action for chloroform.

<IP_ADDRESS>. Carcinogenicity Chloroform has been found to cause liver and kidney tumors in rodents ( discussed in EPA,

1994a, 1998b). A substantial body of data indicates that chloroform is not a DNA reactive mutagen. Thus, mutagenicity is not the key influence of chloroform on the carcinogenic process.

Chloroform induces liver and kidney tumors at doses that cause cell injury or organ toxicity.

Numerous studies have shown that organ toxicity and regenerative proliferation are associated with tumorigenicity of chloroform, and thus are key steps in its carcinogenic mode of action ( EPA, 1998b; ILSI, 1997). To explore the issue of whether fetuses or children are at increased 49 cancer risk compared with adults, the mode of carcinogenic action of chloroform in children versus adults was examined. To address this question, the Agency evaluated the available data for children and adults on chloroform metabolism and age related differences in the rate of cell proliferation ( EPA, 1998b).

Organ toxicity from chloroform is dependent on oxidative metabolism primarily by cytochrome P450 CYP2E1 ( as discussed in EPA, 1998b). The oxidative metabolism of chloroform generates highly tissue reactive metabolites. One metabolite is phosgene, a highly reactive dihalocarbonyl, and the other is the strong acid HCl. The very high reactivity of phosgene prevents it from entering the nucleus and forming adducts with DNA. The chloroform metabolites produce cytotoxicity ( cell death) and regenerative hyperplasia ( EPA, 1998b; ILSI,

1997). This process may lead to tumor development if sustained. Metabolism of chloroform via a reductive pathway, if it occurs, could lead to the formation of free radicals and tissue damage,

but the reductive pathway is absent or minor under normal physiological conditions ( EPA,

2001a).

Given that oxidative metabolism is key to the carcinogenic potential of chloroform, studies on CYP2E1 in fetal and adult tissues were evaluated ( EPA, 1998b). Studies in humans have not shown consistent results; however, in those studies showing expression of CYP2E1, levels in fetuses were lower than those in adults ( Boutelet Bochan et al., 1997; Carpenter et al., 1996;

Hakkola et al., 1998a; Vieira et al., 1996). Vieira et al. ( 1996) suggested that CYP2E1 activity increases rapidly in the 24 hours after birth, and that activity level in children aged 1 to 10 years is comparable to that of adults. 50 Animal studies of CYP2E1 provide evidence of rapid induction of this gene soon after birth ( Song et al., 1986; Umeno et al., 1988; Schenkman et al., 1989; Ueno and Gonzalez, 1990).

The study by Schenkman et al. ( 1989) indicated that CYP2E1 protein is present in low levels in CD rat neonates, rises to a peak level at age 2 weeks and subsequently decreases to adult levels by puberty. Analysis of protein levels quantified from western blots showed a maximum at 2 weeks with decreasing levels at 4 and 12 weeks. The protein level at 12 weeks was approximately 50% of the level at 2 weeks. The authors did not provide a statistical analysis of this result, but it appears from the error bars that the 2 week and 12 week levels ( but not 4 week levels) were significantly different. Song et al. ( 1986) conducted a similar analysis in Sprague

Dawley rats and reported a rapid transcriptional induction of CYP2E1 ( P450) within 1 week following birth that remained elevated throughout 12 weeks. Enzyme activity followed a similar pattern. Ueno and Gonzalez ( 1990) showed that extracts from 3 day old and 12 week old rat liver, but not those from fetal or newborn rat liver, were able to generate significant CYP2E1 transcription in vitro. The ability of the extract to result in transcription of CYP2E1 was slightly greater at 12 weeks.

Taken together, the animal studies do not provide conclusive evidence of an early period of increased enzymatic activity. If, however, the twofold increase in CYP2E1 induction in animals were seen in humans, its importance in terms of chloroform toxicity would depend on the dose. Under low dose conditions ( e. g., much lower than the Km) it is possible that an increase in the level of enzyme would not have any effect on active metabolite formation, because the amount of chloroform, and not CYP2E1, would control the rate of phosgene production. On the other hand, under saturating doses of chloroform, all the available enzyme would be active; thus a 51 twofold increase in CYP2E1 could result in greater activation of the compound. Although the animal data remain unclear regarding the potential for a neonatal period of increased CYP2E1 activity above that in the adult, the data in humans show a rapid induction after birth, gradually increasing over the first year to reach adult levels during years 1 10 ( Vieira et al., 1996).

Therefore, although children may have the capacity to metabolize chloroform, data on CYP2E1 activity provide no evidence to suggest that children have an increased susceptibility to chloroform toxicity compared with adults ( EPA, 1998b). Furthermore, the data from Schenkman et al. ( 1989) indicate that levels of CYP2E1 are approximately 2 fold higher in rats at 2 weeks than at 12 weeks of age. The 2 fold magnitude of the difference in CYP2E1 levels is within the default UF value of 3 used to account for intraspecies variability in toxicokinetics ( the CYP2E1 studies do not address the contribution of toxicodynamics to the total default UF of 10). Based on this analysis, differences in age dependent metabolism of chloroform by CYP2E1 are adequately accounted for in the existing UF for intraspecies variability.

Thus, the human data are inconclusive on age related differences in CYP2E1 activity, and it is not possible to state whether children may have an increased susceptibility to chloroform toxicity and carcinogenicity as compared to adults. However, based on animal data, the degree of age dependent differences in CYP2E1 expression ( if any) are likely well accounted for by default assumptions about the magnitude of intraspecies variability.

The next issue to examine is whether the developing fetus may be more sensitive to the toxicity of chloroform because of its greater rate of cell proliferation. There are very few data on prenatal and postnatal exposures to chloroform and resultant organ toxicity. Liver toxicity was found in a multigeneration reproductive assay in CD 1 mice ( NTP, 1988). The periods of 52 exposure included prenatal, postnatal and adult stages. The liver toxicity from this multigeneration reproductive study was compared with that of a comparable 90 day study in adult B6C3F1 mice for liver toxicity ( Bull et al., 1986). The similarity of effects at comparable doses from these two studies suggests that there is no increased susceptibility to chloroform that results from prenatal or postnatal exposures ( EPA, 1998b). It should be noted that there are limitations in this comparison; different strains of mice were used, and only LOAELs were identified in these two studies ( EPA, 1998b).

Under the 1986 U. S. EPA Guidelines for Carcinogen Risk Assessment, chloroform has been classified as Group B2, probable human carcinogen, based on sufficient evidence of carcinogenicity in animals ( EPA, 1998a). Under the Draft Guidelines for Carcinogen Risk Assessment ( EPA 1999), chloroform is likely to be carcinogenic to humans by all routes of exposure under high exposure conditions that lead to cytotoxicity and regenerative hyperplasia in susceptible tissues ( EPA, 2001a). Chloroform is not likely to be carcinogenic to humans by any route of exposure under exposure conditions that do not cause cytotoxicity and cell regeneration.

This weight of evidence conclusion is based on several lines of evidence. 1) Observations in animals exposed by both oral and inhalation pathways indicate that sustained or repeated cytotoxicity with secondary regenerative hyperplasia precedes, and is probably required for,

hepatic and renal neoplasia. 2) There are no epidemiological data specific to chloroform. The epidemiological data relating cancer to exposure to chlorinated drinking water are mixed, and any observed effects cannot necessarily be attributed to chloroform amongst multiple other disinfection byproducts. 3) Genotoxicity data on chloroform are essentially negative; there are a few scattered positive results that generally have limitations such as excessively high doses or 53 confounding factors. Thus, the weight of evidence of the genotoxicity data on chloroform supports a conclusion that chloroform is not strongly mutagenic, and that genotoxicity is not likely to be the predominant mode of action underlying the carcinogenic potential of chloroform ( EPA, 2001a).

Because a substantial database indicates that tumor development for chloroform is secondary to organ toxicity and regenerative proliferation, a nonlinear dose response approach for tumorigenicity is viewed as appropriate. Use of a low dose linear approach is considered overly conservative for extrapolating cancer risk associated with chloroform exposure ( EPA, 1998a, b,

2001a; ILSI, 1997). EPA ( 2001a) determined a point of departure ( POD) of 23 mg/ kg/ day,

based on increased kidney tumors in Osborne Mendel rats ( Jorgensen et al., 1985). The Agency compared this POD to the RfD of 0.01 mg/ kg/ day calculated as described in the next paragraph,

and concluded that the margin of exposure ( MOE) of 2000 was adequately protective of public health for cancer effects. Sustained tissue toxicity, which is a key event in the cancer mode of action for chloroform, will not occur at doses below the RfD.

<IP_ADDRESS>. Basis for RfD and MCLG The Agency used an oral study in dogs ( Heywood et al., 1979) to derive the chloroform RfD. The study identified a LOAEL of 15 mg/ kg/ day for hepatotoxicity ( fatty cysts and an increase of serum glutamic pyruvic transaminase). A BMDL 10 ( the 95% lower bound confidence limit on the dose associated with a 10% extra risk) of 1.2 mg/ kg/ day was also calculated, based on the prevalence of animals with moderate to marked fatty cysts in liver in the same study.

Applying a factor of 6/ 7 to account for exposure 6 days/ week, the LOAEL was converted to 12.9 54 mg/ kg/ day, and the BMDL 10 was converted to 1.0 mg/ kg/ day. The RfD was based on both the benchmark dose ( BMD) and traditional NOAEL/ LOAEL approach, which coincidentally result in the same RfD. An RfD of 0.01 mg/ kg/ day was calculated using the BMD approach by applying an overall UF of 100 ( 10 for interspecies extrapolation and 10 for protection of sensitive individuals) to the BMDL 10 of 1.0 mg/ kg/ day. Using the traditional approach, a composite UF of 1,000 was applied ( 100 was used to account for inter and intraspecies differences and a factor of 10 for use of a LOAEL), resulting in an RfD of 0.01 mg/ kg/ day. This RfD corresponds to a Drinking Water Equivalent Level ( DWEL) of 0.35 mg/ L, assuming an adult tap water consumption of 2 L/ day for a 70 kg adult.

Because a nonlinear approach was used for chloroform carcinogenicity, the MCLG is based on liver toxicity as the most sensitive effect for chloroform ( as it is the lowest possible LOAEL for any organ) and as a precursor response to a key step to its carcinogenicity. This approach is considered equally protective of both adults and children because the database on chloroform does not indicate that children are more sensitive than adults to liver toxicity. The mode of action by which chloroform produces organ toxicity and carcinogenicity is considered to be the same for children and adults.

The MCLG of 0.07 mg/ L was calculated from the RfD by assuming an adult tap water consumption of 2 L per day for a 70 kg adult, and by applying a relative source contribution ( RSC) of 20%. The RSC is based on data indicating that exposure to chloroform by other routes and sources of exposure may potentially contribute a substantial percentage of the overall exposure to chloroform ( EPA, 2001d). Based on average daily doses for each source and route of exposure under specific conditions, EPA estimated that for the median individual, ingestion of 55 total tap water ( including water based drinks) can contribute roughly 28% of the total dose of chloroform, while the dose from showering ( inhalation and dermal exposure) was estimated to contribute approximately 14% of the total dose ( EPA, 2001d). There is, however, considerable uncertainty in these exposure estimates. For example, because chloroform is so volatile, most of the chloroform would evaporate from ingested hot liquids such as coffee or tea, so the value of 28% from ingestion of total tap water may be an over estimate. In addition, the proportion of intake from different sources varies widely across the population. Therefore, the default RSC of 20% is used, a value that is consistent with the percent of the total dose from ingestion of total tap water. Thus, the MCLG is 0.07 mg/ L: ( 0.01 mg/ kg/ day x 70 kg x 0.2)/ ( 2/ L/ day) = 0.07 mg/ L.

<IP_ADDRESS>. Children's Risk in Relation to the MCLG The MCLG derived for chloroform is considered protective of both adults and children,

given that developmental effects occurred at doses above those causing hepatotoxicity. In addition, fetal effects were only seen at levels at which maternal toxicity was noted. Also, the mode of action data indicates that children are not uniquely sensitive to the organ toxicity caused by high doses of chloroform and there is no evidence from the available studies to suggest that children or fetuses would be qualitatively more sensitive to its effects than adults. In addition, the developing fetus would not be expected to be particularly sensitive to a cytotoxic agent, such as chloroform at low levels, because cell division occurs at a rapid pace and the cellular repair capacity is high ( EPA, 2001a). 56 The MCLG is also considered protective of carcinogenic effects in children and fetuses.

EPA believes that using a nonlinear dose response approach for setting the MCLG is protective for children and fetuses for the following reasons: 1) the reactive metabolite inside the cell should have similar effects resulting from its reacting with and disrupting macromolecules in the cells of fetuses, children, and adults; 2) cell necrosis and reparative replication are not likely to be qualitatively different in various life stages; 3) cancer risk to the fetus or children would be a function of cytotoxicity, as in adults, and protecting fetuses and children from sufficient levels of the chemical that would cause cytotoxicity should protect against cancer risk ( EPA, 2001a); 4) a comparison of liver toxicity between a multigeneration reproductive study and a comparable 90

day study suggests that there is no increased susceptibility to chloroform effects on the liver that results from prenatal or postnatal exposure, indicating that children are not more sensitive to the liver cytotoxicity that is a precursor to chloroform carcinogenicity ( EPA, 1998b). This final factor also supports the conclusion that children are not more sensitive than adults to the noncancer liver effects of chloroform.

2.2.2. Brominated Trihalomethanes There is sufficient evidence for carcinogenicity via ingestion of bromoform and bromodichloromethane ( BDCM) to consider them probable human carcinogens. The evidence is limited for dibromochloromethane ( DBCM). Based on the available data, a mechanism of action involving mutagenicity was postulated for the brominated THMs, indicating that linear low dose extrapolation for BDCM and bromoform is appropriate. The proposed mechanism of carcinogenicity for these compounds was examined to determine if this would provide any reason 57 for concern that children or fetuses may be more susceptible to development of cancer following exposure. If carcinogenicity is the result of mutations by either the parent compound or a metabolite, children or the developing fetus could be more susceptible to the carcinogenicity of brominated THMs due to a higher rate of cell proliferation in the target organs. An increased risk of this type would be true for all genotoxic carcinogens and not specific to brominated THMs.

There are no data currently available for brominated THMs to permit quantification of a possible increase in risk to the developing fetus or children.

A number of epidemiology studies have reported an association between exposure to THMs and developmental/ reproductive effects ( Bove et al., 1995; Dodds et al., 1999; Gallagher et al., 1998; King et al., 2000a; Klotz and Pyrch, 1999; Kramer et al., 1992; Savitz et al., 1995;

Waller et al., 1998, Windham et al., 2003). For all of these studies, because the subjects were exposed to other contaminants and disinfection byproducts in the drinking water, correlation of the effects directly to individual brominated THM exposure is difficult. Further details about these studies were presented in Section 2.1, Chlorinated Drinking Water.

Although the mechanism of brominated trihalomethane toxicity is not known with certainty, data indicate that the adverse effects of this group of chemicals are secondary to metabolism. Brominated THMs are extensively metabolized via oxidative ( using NADPH and oxygen) and reductive ( using NADPH or NADH and inhibited by oxygen) pathways in humans and animals, primarily in the liver, but also in the kidney ( EPA, 2002c). Recent data suggest that bioactivation of brominated trihalomethanes to mutagenic species is also mediated by one or more glutathione S transferase mediated conjugation pathways. 58 Both oxidative and reductive reactions are mediated by cytochrome P450s. Oxidative metabolism results in the production of a dihalocarbonyl ( CX 2 O) intermediate, which may undergo a variety of reactions, such as adduct formation with cellular nucleophiles, hydrolysis to yield carbon dioxide or glutathione dependent reduction to yield carbon monoxide. Reductive metabolism results in the production of free radical species such as the dihalomethyl radical ( CHX 2 .). In vitro and in vivo data suggest that metabolism via the reductive pathway occurs more readily for the brominated trihalomethanes than it does for chloroform. Gao and Pegram ( 1992) reported that binding of reactive intermediates to rat hepatic microsomal lipid and protein under reductive ( anaerobic) conditions was more than twice as high for bromodichloromethane as for chloroform. Tomasi et al. ( 1985) used electron spin resonance ( ESR) spectroscopy to measure the production of a free radical intermediates ( a product of the reductive pathway) in vitro using rat hepatocytes isolated from phenobarbital induced male Wistar rats. The intensity of the ESR signal was greatest for bromoform, followed by BDCM and then chloroform. The largest ESR signal was detected when hepatocytes were incubated under anaerobic conditions.

Tomasi et al. ( 1985) also used ESR to evaluate free radical production in vivo in rats given intraperitoneal injections of chloroform, BDCM, or bromoform. The intensity of the ESR signal followed a ranking similar to that observed in in vitro experiments ( bromoform > BDCM>

chloroform), confirming that the reductive formation of free radicals is greater for brominated trihalomethanes than for chloroform. Together, these data indicate that reductive metabolism is a more important pathway for metabolism of brominated trihalomethanes than for chloroform. The relative importance of the oxidative and reductive pathways for the brominated THMs in vivo has not been determined, but it is of note that oxygen partial pressure in the kidney and liver is low 59 under in vivo conditions. Both dihalocarbonyls ( produced by the oxidative pathway) and dihalomethyl radicals ( produced by the reductive pathway) are reactive species and may form covalent adducts with a variety of cellular components ( EPA, 2002c).

Metabolism of the brominated THMs involves a number of enzymes, including CYP2E1,

CYP2B1/ 2 and CYP1A ( EPA, 2002c). The roles of the different P450 isozymes in brominated trihalomethane metabolism ( i. e., oxidative versus reductive pathways) have not been definitively identified, but CYP2E1, CYP2B1/ 2 and CYP1A have been implicated in the oxidative pathway.

The toxicity of BDCM and DBCM has been shown to be at least partly related to bioactivation by the enzyme CYP2E1 ( e. g., Thornton Manning et al., 1994). Thus, a higher level of CYP2E1 activity in children, as compared to adults, would mean that children could be at greater risk for carcinogenic effects from these compounds.

Studies in humans have not shown consistent results; however, in those studies showing expression of CYP2E1, levels in fetuses were lower than those in adults ( Boutelet Bochan et al.,

1997; Carpenter et al., 1996; Hakkola et al., 1998a; Vieira et al., 1996). Vieira et al. ( 1996)

suggested that CYP2E1 activity increases rapidly in the 24 hours after birth, and that activity level in children aged 1 to 10 years is comparable to that of adults.

Animal studies of CYP2E1 provide evidence of rapid induction of this gene soon after birth ( Song et al., 1986; Umeno et al., 1988; Schenkman et al., 1989; Ueno and Gonzalez, 1990).

The study by Schenkman et al. ( 1989) indicated that CYP2E1 protein is present in low levels in CD rat neonates, rises to a peak level at age 2 weeks and subsequently decreases to adult levels by puberty. Analysis of protein levels quantified from western blots showed a maximum at 2 weeks with decreasing levels at 4 and 12 weeks. The protein level at 12 weeks was 60 approximately 50% of the level at 2 weeks. The authors did not provide a statistical analysis of this result, but it appears from the error bars that the 2 week and 12 week levels ( but not 4 week levels) were significantly different. Song et al. ( 1986) conducted a similar analysis in Sprague Dawley rats and reported a rapid transcriptional induction of CYP2E1 ( P450) within one week following birth that remained elevated throughout 12 weeks. Enzyme activity followed a similar pattern. Ueno and Gonzalez ( 1990) showed that extracts from 3 day old and 12

weekold rat liver, but not those from fetal or newborn rat liver, were able to generate significant CYP2E1 transcription in vitro. The ability of the extract to result in transcription of CYP2E1 was slightly greater at 12 weeks.

Taken together, the animal studies do not provide conclusive evidence of an early period of increased enzymatic activity. If, however, the 2 fold increase in CYP2E1 induction in animals were seen in humans, its importance in terms of brominated THM toxicity would depend on the dose. Under low dose conditions ( e. g., much lower than the Km) it is possible that an increase in the level of enzyme would not have any effect on active metabolite formation, because the amount of brominated THMs, and not CYP2E1, would control the rate of the enzymatic metabolism. On the other hand, under saturating doses of brominated THMs, all the available enzyme would be active; thus a 2 fold increase in CYP2E1 could result in greater activation of the compound.

Although the animal data remain unclear regarding the potential for a neonatal period of increased CYP2E1 activity above that in the adult, the data in humans show a rapid induction after birth,

gradually increasing over the first year to reach adult levels during years 1 10 ( Vieira et al., 1996).

Therefore, although children may have the capacity to metabolize brominated THMs, data on CYP2E1 activity provide no evidence to suggest that children have an increased susceptibility to 61 brominated THM toxicity compared with adults ( EPA, 2002c). Furthermore, the data from Schenkman et al. ( 1989) indicate that levels of CYP2E1 are approximately 2 fold higher in rats at 2 weeks versus 12 weeks of age. The 2 fold magnitude of the difference in CYP2E1 levels is within the default UF value of 3 used to account for intraspecies variability in toxicokinetics ( the CYP2E1 studies do not address the contribution of toxicodynamics to the total default UF of 10).

Based on this analysis, differences in age dependent metabolism of brominated THMs by CYP2E1 are adequately accounted for in the existing UF for intraspecies variability.

Thus, the human data are inconclusive on age related differences in CYP2E1 activity and it is not possible to state whether children may have an increased susceptibility to brominated THM toxicity as compared to adults. However, based on animal data, the degree of agedependent differences in CYP2E1 expression ( if any) are likely well accounted for by default assumptions about the magnitude of intraspecies variability.

Information on age related differences in the activity of the other human cytochrome P450s involved in brominated THM metabolism was reviewed in a recent summary of the current literature ( EPA, 2000d). For CYP1A1, constitutive fetal mRNA levels were measurable in some studies, with expression levels continuing to increase through gestation ( Oesterheld, 1998).

However, a decline in CYP1A1 activity was reported following birth ( Rendic and Di Carlo,

1997). No data on age dependent differences in inducible levels of CYP1A1 were presented in the review. For CYP1A2, low fetal mRNA levels and low enzyme activity in the neonate were reported. However, there were similar levels of CYP1A2 enzyme activity in adults and in infants,

and activity was greater in young children than in adults ( Hakkola et al., 1998a; 1998b). Based on patterns of theophylline clearance as a measure of CYP1A2 activity, CYP1A2 is minimally 62 active in fetuses, reaches maximum activity during early childhood and decreases thereafter ( EPA,

2000d).

The human CYP isoforms CYP2A6, CYP2D6, and CYP3A4 are potential candidates for metabolism of brominated THMs, based on overlapping catalytic activity with rodent CYP2B1/ 2 ( WHO, 2000), but the identity of the human CYP isoforms, other than CYP2E1, capable of metabolizing brominated THMs is not known. In light of this knowledge gap, the toxicological consequences of their developmental expression patterns are not clear. However, except for CYP1A2, fetuses and children generally have lower expression than adults of the CYPs that may be involved in brominated THM metabolism ( Hakkola et al., 1998a, 1998b; Tanaka, 1998;

Oesterheld, 1998).

Higher enzyme activity will not necessarily result in higher tissue doses of metabolite,

since at low doses ( i. e., much lower than the Km), the amount of metabolism would be controlled by the amount of substrate, not the amount of enzyme. In addition, if metabolism of the brominated THMs by CYP1A, CYP2B1/ 2, or their analogues represents a low affinity route of metabolism ( as it does for chloroform), the impact of age dependent expression of these enzymes at environmental exposure levels may be less important than for CYP2E1. No data were identified on the relative affinity of the brominated THMs for the different CYP isoforms.

Overall, the P 450 enzyme expression and activity data do not suggest that children would be more susceptible than adults to brominated THM.

Only minimal data are available regarding the enzymes involved in the reductive metabolism of brominated trihalomethanes. While experimental evidence indicates that CYP2E1 and CYPB1/ 2 catalyze the oxidative pathway, the identities of the cytochrome P450 isoforms that 63 catalyze the reductive pathway have not been established. In general, CYP2E1 protein can catalyze reductive as well as oxidative reactions ( Lieber, 1997) and this isoform has been implicated in the production of trichloromethyl radicals from carbon tetrachloride ( see Lieber et al. 1997). However, evidence for a dual role of either CYP2E1 or CYP2B1/ 2 in catalyzing the oxidative and reductive pathways for trihalomethane metabolism has been contradictory, perhaps as a result of the different concentrations of chloroform used in different experiments ( summarized in Testai et al., 1996). To address the issue of concentration, Testai et al. ( 1996)

studied the role of different isoforms in chloroform metabolism, and they found that the cytochrome P450 isoforms involved in oxidative metabolism of brominated trihalomethanes do not participate in the reductive pathway. Thus, no conclusion can be made on whether children are more sensitive than adults based on the limited data on the reductive pathway.

The information on age related differences in the production of metabolites via the glutathione conjugation pathway is also very limited. Studies in Salmonella typhimurium strains engineered to express the rat glutathione S transferase theta 1 1 ( GSTT1 1) gene indicate that metabolism of brominated THMs to mutagens may also be catalyzed by glutathione S transferase ( DeMarini et al., 1997; Landi et al., 1999). Data on the GST theta genes are currently quite limited; however, one study reported that theta class GSTs were expressed in human adult liver,

but not fetal liver ( Mera et al., 1994). These results suggest that the fetus does not experience increased risk from GST theta mediated mutagenicity. The occurrence of increased risk in children cannot be evaluated, since the age at which expression of GST theta begins is unknown.

<IP_ADDRESS>. Bromodichloromethane 64 Developmental/ Reproductive Effects Developmental and reproductive toxicity data are available for bromodichloromethane ( BDCM) and were considered in the derivation of the MCLG. Five epidemiologic studies have reported a relationship specifically between developmental/ reproductive toxicity and BDCM.

These studies were summarized in detail in Section 2.1.1., but key conclusions are highlighted here. In a prospective study, Windham et al. ( 2003) demonstrated that increasing levels of individual brominated trihalomethanes or total brominated trihalomethanes in the drinking water were associated with significantly shorter cycles when examined by quartile. Similar decrements were observed in follicular, but not luteal, phase length. For BDCM, the adjusted decrements were 0.74 days ( 95% C. I.  1.5,  0.02) for mean cycle length and 0.8 days ( 95% C. I.  1.5,  0.08)

for mean follicular phase length at the highest quartile (  16  g/ L). The strongest association for an individual compound was observed for DBCM. Menses length was slightly increased at the highest quartile for BDCM exposure.

A population based case control study ( Kramer et al., 1992) reported an increased risk ( not statistically significant) of intrauterine growth retardation with exposure to BDCM concentrations in drinking water of greater than or equal to 10 µ g/ L, compared with drinking water with undetectable BDCM concentrations ( OR = 1.7, 95% CI = 0.9   2.9). As previously discussed, Waller et al. ( 1998) reported that pregnant women exposed to TTHMs in drinking water at levels of 75  g/ L or higher had an increased risk of spontaneous abortion. This study also reported that consumption of five or more glasses of cold water with a BDCM concentration of at least 18 µ g/ L was associated with an increased risk of spontaneous abortion. After adjustment for exposure to other THMs, the adjusted OR was 3.0 ( CI = 1.4 6.6). 65 King et al. ( 2000a) reported that exposures to TTHMs, chloroform and BDCM were associated with an increased risk of stillbirth. Risk doubled for women exposed to a BDCM level of greater than or equal to 20  g/ L, when compared to women consuming concentrations of less than 5  g/ L. In a retrospective cohort study among 49,842 residents of Nova Scotia ( Dodds and King, 2001), exposure to BDCM at concentrations  20  g/ L was associated with an increased risk of neural tube defects ( adjusted relative risk = 2.5, 95% CI = 1.2  5.1), but there was no evidence of a dose response trend.

Seven oral studies in laboratory animals evaluated the developmental toxicity of BDCM ( Bielmeier et al., 2001; CCC, 2000a, b, c; Narotsky et al., 1997a; NTP, 1998a; Ruddick et al.,

1983). To determine the effect of vehicle on BDCM toxicity, Narotsky et al. ( 1997a)

administered BDCM by gavage in either corn oil or an aqueous vehicle with Emulphor ® to pregnant F344 rats ( 12   14/ group) at dose levels of 0, 25, 50, or 75 mg/ kg/ day during gestation days 6   15. Decreased maternal weight gain and full litter resorption were observed at 50 and 75 mg/ kg/ day. The incidence of full litter resorption was significantly higher in the corn oil vehicle ( 83%) compared with the aqueous vehicle ( 21%) at the high dose. Accordingly, the NOAEL for developmental toxicity ( full litter resorptions) was 25 mg/ kg/ day, with a LOAEL of 50 mg/ kg/ day. The LOAELs for maternal effects ( based on decreased maternal weight gain during GD 6 8) in aqueous and corn oil vehicles were 25 and 50 mg/ kg/ day, respectively. No maternal NOAEL for the aqueous study could be identified.

Bielmeier et al. ( 2001) conducted a series of experiments on the effect of BDCM on pregnancy loss, characterized as full litter resorptions, in female F344 and Sprague Dawley rats.

In one experiment, doses of 0, 75, or 100 mg/ kg/ day were administered via gavage to pregnant 66 rats on gestation day 9 in 10% Emulphor. The dose related incidence of full litter resorptions was 0% ( 0/ 8), 64% ( 7/ 11), and 90% ( 9/ 10), respectively, identifying a LOAEL of 75 mg/ kg/ day. This effect was seen in F344 rats, but not in Sprague Dawley rats. Dosing during different portions of pregnancy showed that the critical period for induction of full litter resorption was limited to the luteinizing hormone ( LH) dependent phase of pregnancy, suggesting that BDCM may disrupt pregnancy via a LH mediated mode of action. Measurement of serum LH and progesterone levels indicated that full litter resorption was accompanied by a marked reduction in progesterone concentration without a corresponding drop in LH levels. The failure of BDCM to exert adverse effects after the LH dependent window, the reduction in serum progesterone level, and the unchanged serum LH levels led this group to conclude that the target of toxicity was the ovary and that the mode of action was a reduced sensitivity of the corpus luteum to LH. Although there are significant differences between rats and humans in the hormonal maintenance of pregnancy,

the authors did consider their findings possibly relevant to humans. Additional research, some of which is currently ongoing, is needed to evaluate the relevance to humans of these findings in rats.

Ruddick et al. ( 1983) investigated developmental toxicity in pregnant Sprague Dawley rats ( 9   15/ group) administered BDCM by gavage in corn oil at dose levels of 0, 50, 100, or 200 mg/ kg/ day from gestation days 6   15. Maternal weight gain was significantly depressed in the high dose group, and reduced in the low and mid dose groups; the decreases were not statistically significant. There were no fetal malformations, but a dose dependent increased incidence of sternebra aberrations was observed in all dose groups. Although there was a clear increase in the incidence of these variations, no statistical analysis was performed by the authors.

However, a statistical analysis ( using the Fisher Exact Test) conducted on the published data 67 found that none of these increases differed significantly from controls. A trend test showed a statistically significant dose related trend ( p= 0.03), and stepwise analysis indicated that the trend became significant only when the high dose ( 200 mg/ kg/ day) was included in the analysis. These findings suggest that the NOAEL and LOAEL were 100 mg/ kg/ day and 200 mg/ kg/ day,

respectively. However, due to the small sample sizes, the statistical power of the experiment to detect effects at lower doses is limited. The NOAEL and LOAEL for maternal effects were 100 mg/ kg/ day and 200 mg/ kg/ day, respectively, based on significantly decreased maternal weight gain ( EPA 2002c).

NTP ( 1998a) conducted a short term study screening study in Sprague Dawley rats investigating both reproductive and developmental toxicity from BDCM administered in drinking water. Two groups of male rats and three groups of female rats were treated with BDCM at concentrations of 0, 100, 700, or 1300 mg/ L. Based on measured water consumption, the authors estimated dose levels for the treated males to be 0, 8, 41, or 68 mg/ kg/ day, and for the treated females to be 0, 14, 72, or 116 mg/ kg/ day ( groups A and C) and 0, 13, 54, or 90 mg/ kg/ day ( group B). The rats were exposed for 25 to 30 days, with the exception of group B females, which were exposed from gestation day 6 to evidence of littering/ birth ( approximately 15 days). BDCM exposure did not affect any reproductive parameter investigated in males or females, with the exception of a non dose related increase in the number of live fetuses per birth at the 14 mg/ kg/ day dose in Group C females and a slight decrease in the number of live fetuses per birth at the 72 mg/ kg/ day dose in Group A females. On the basis of these results, NTP ( 1998a) concluded that BDCM was not a short term reproductive or developmental toxicant at doses up to approximately 68 and 116 mg/ kg/ day in male and female rats, respectively. The study 68 sensitivity was decreased by the use of a relatively small number of animals per group ( 5

13/ sex/ dose) and the lack of microscopic examination of the pups.

The Chlorine Chemistry Council sponsored a range finding reproductive/ developmental toxicity study of BDCM in Sprague Dawley rats ( CCC, 2000a; Christian et al., 2001b). BDCM was administered to parental rats ( P generation, 10/ sex/ group) in drinking water at concentrations of 0, 50, 150, 450, or 1350 ppm. Exposure began 14 days before cohabitation and continued until the day of sacrifice. Lactation was extended for one week ( LD 22 29) beyond the normal 3

week period because F 1 pup body weights in the three highest dose groups were significantly reduced on LD 21 relative to control values ( results are described below). On LD 29, two F 1 pups per sex were selected from each litter for an additional week of postweaning observation,

provided ad libitum access to water containing the same concentration of BDCM administered to their parents ( P generation), and sacrificed on Day 8 postweaning.

Exposure dependent reductions in both absolute ( g/ day) and relative ( g/ kg body weightday water consumption were observed in all rats of both sexes and were attributed to taste aversion. Treatment related clinical signs ( e. g., dehydration, emaciation, chromorhinorrhea) were observed in both sexes in the 1350 ppm exposure groups and were considered to be generally associated with reduced water consumption. The only other observed effect was a concentrationdependent reduction in F 1 pup body weights and weight gain in the 150, 450, and 1350 ppm exposure groups, both prior to, and after weaning. Based on decreased pup weight and pup weight gain, the LOAEL for developmental toxicity is 150 ppm, and the corresponding NOAEL is 50 ppm. Although the effect of reduced water consumption on the decreases in feed consumption, body weight gain, and body weight observed in the P generation adults is unclear, 69 the LOAEL for parental toxicity is considered to be 150 ppm and the NOAEL is 50 ppm. Due to the marked changes in drinking water consumption by P generation female rats during different physiological stages ( pre mating, mating, gestation, and lactation), it is not possible to convert the administered drinking water concentrations into a single biologically meaningful average daily dose; instead, dose in mg/ kg/ day was reported separately for the premating, gestation, lactation,

and post weaning phases of the study.

In the full study, the Chlorine Chemistry Council ( CCC, 2000b; summarized in Christian et al., 2001a) examined the developmental effects of BDCM in Sprague Dawley rats. Pregnant rats were given BDCM in drinking water at concentrations of 0, 50, 150, 450, or 900 mg/ L ( 0, 2.2,

18.4, 45.0, or 82.0 mg/ kg/ day) on gestation days 6 21. No treatment related clinical signs or necropsy results were observed. Significantly reduced water consumption was observed in all dose groups and was attributed to taste aversion. Decreases in absolute and relative feed consumption were observed in the three highest dose groups; associated decreases in maternal body weight gain were attributed to taste aversion. The effect on maternal body weight gain was persistent at the two highest doses, but was transient at the lower doses. No effects on early or late resorptions, fetal body weight, liver litter size, or other developmental parameters were observed from BDCM exposure. There were no instances of full litter resorption and no dead fetuses. The only statistically significant changes in the occurrence of skeletal variations were reversible delays in ossification at the high dose, including an increased fetal incidence of wavy ribs and a decreased number of ossification sites per fetus per litter for the forelimb phalanges and the hindlimb metatarsals and phalanges. The study authors did not consider the increased fetal incidence of wavy ribs to be related to BDCM exposure, because the more relevant measure of 70 litter incidence did not differ significantly from the controls and was within the historical range for this alteration at the test facility. The maternal NOAEL and LOAEL were 18.4 mg/ kg/ day and 45.0 mg/ kg/ day, respectively, based on significant reductions in maternal body weight and body weight gains, while the developmental NOAEL and LOAEL were 45.0 mg/ kg/ day and 82.0 mg/ kg/ day, respectively, based on a significant decrease in the number of ossification sites per fetus for the forelimb phalanges and the hindlimb metatarsals and phalanges.

The Chlorine Chemistry Council also examined the developmental effects of BDCM in New Zealand white rabbits ( CCC, 2000c). Pregnant rabbits were given BDCM in drinking water at concentrations of 0, 15, 150, 450, or 900 mg/ L ( 0, 1.4, 13.4, 35.6, or 55.3 mg/ kg/ day) on gestation days 6 29. No treatment related clinical signs or necropsy results were observed.

Significantly reduced feed and water consumption and body weight gains were observed in the 35.6 and 55.3 mg/ kg/ day dose groups. No observable effects on fetal body weight, live litter size,

and a number of other developmental parameters were noted from BDCM exposure. Statistically significant increases in the number of fused sterna centra were observed in the 13.4 and 35.6 mg/ kg/ day dose groups; however, this effect was not dose related, and the observed incidences were within the historical range for the testing facility. The maternal NOAEL and LOAEL identified in this study were 13.4 and 35.6 mg/ kg/ day, respectively, based on decreased body weight gain, while the developmental NOAEL was 55.3 mg/ kg/ day, based on an absence of statistically significant, dose related effects at any dose tested.

Christian et al. ( 2002) conduced a standard two generation reproductive toxicity study in which BDCM was continuously provided to Sprague Dawley rats in the drinking water at concentrations of 0, 50, 150, or 450 ppm. Average daily doses estimated by the study authors 71 were 4.1 to 12.6, 11.6 to 40.2, and 29.5 to 109 mg/ kg day, respectively. Exposure of the parental generation ( P) was initiated when the test animals were approximately 43 days of age and continued through a 70 day pre mating period and a cohabitation period. P generation males were exposed for approximately 106 days prior to sacrifice. Exposure of P generation female rats continued through gestation and lactation. F 1 generation rats were exposed to BDCM in utero and by consumption of the dam's drinking water during the lactation period. At weaning, F 1 rats ( 30/ sex/ concentration) were selected for a postweaning/ premating exposure period, followed by a cohabitation period, and the exposure continued through gestation and lactation. F 1 generation females delivered litters and the F 2 litters were sacrificed on lactation day 22.

Deaths at 150 and 450 ppm were associated with reduced water consumption, weight loss and/ or adverse clinical signs and may have been compound related. Adverse clinical signs occurred at 150 and 450 ppm were attributed to reduced water consumption. Body weight and body weight gain were significantly reduced in the 450 ppm P generation males and females and 150 and 450 ppm F 1 generation males and females, and was associated with decreased food consumption. Rats at 450 ppm also had decreased absolute and relative organ weights. Water consumption was significantly reduced in P and F 1 generation males and females at all concentrations of BDCM, and was reduced by 10 20% at 150 and 450 ppm. There were no gross pathological or histopathological indications of compound related toxicity. Most indicators of reproductive or developmental toxicity examined were not significantly affected by BDCM treatment. However, decreased pup body weight was observed at weaning and/ or during lactation in the F 1 and F 2 generations in the 150 and 450 ppm groups. Small, but statistically significant, delays in F 1 generation sexual maturation occurred at 150 ( males) and 450 ppm 72 ( males and females) as determined by timing of vaginal patency or preputial separation, but no statistically significant delays were seen when body weight at weaning was included as a covariate in the analysis. Delayed estrus was observed in F 1 females in the 450 ppm exposure group. The study authors also considered this effect to be a secondary response associated with reduced pup weights. The results of this study appear to identify NOAEL and LOAEL values for reproductive effects of 50 ppm ( 4.1 to 12.6 mg/ kg day) and 150 ppm ( 11.6 to 40.2 mg/ kg day), respectively,

based on delayed sexual maturation. However, the study authors have questioned whether delayed sexual maturation in F 1 males associated with reduced body weight should be treated as reproductive toxicity or general toxicity, since the root cause appears to be dehydration brought about by taste aversion to the compound. The parental NOAEL and LOAEL are also 50 and 150 ppm, respectively, based on reduced body weight and body weight gain in F 1 males and females.

Klinefelter et al. ( 1995) evaluated the effects of BDCM exposure on male reproduction at an interim sacrifice as part of a two year bioassay, in which F344 rats were administered BDCM in drinking water at concentrations of 0, 330 mg/ L, or 620 mg/ L. The authors estimated the doses to be 0, 22, and 39 mg/ kg/ day. At 52 weeks, the authors conducted an interim sacrifice,

which included an evaluation of epididymal sperm motion parameters and histopathology of the testes and epididymides. No histologic alterations were observed in any reproductive tissue.

Sperm velocities ( mean straight line, average path, and curvilinear), however, were significantly decreased at 39 mg/ kg/ day. No effect on sperm motility was observed at 22 mg/ kg/ day. The NOAEL and LOAEL for reproductive effects are thus 22 and 39 mg/ kg/ day, respectively.

Several studies on the developmental toxicity of BDCM gave negative results at doses up to 116 mg/ kg/ day in Sprague Dawley rats ( NTP, 1998a) and 55.3 mg/ kg/ day in rabbits ( CCC, 73 2000c) when administered in drinking water. However, in other studies ( CCC, 2000b; Ruddick et al., 1983), slightly decreased numbers of ossification sites in the hindlimb and forelimb were observed in fetuses of Sprague Dawley rats administered 82 mg/ kg/ day in the drinking water on gestation days 6 to 21 ( CCC, 2000b) and sternebral aberrations were observed in the offspring of Sprague Dawley rats administered 200 mg/ kg/ day by gavage in corn oil on gestation days 6 to 15 ( Ruddick et al., 1983). Reductions in mean pup weight gain and pup weight were observed when parental Sprague Dawley rats were administered BDCM in the drinking water at concentrations of 150 ppm and above ( a biologically single meaningful estimate of intake on a mg/ kg/ day basis could not be calculated for this study) ( CCC, 2000a). Full litter resorption has been noted in F344 rats ( Narotsky et al., 1997a; Bielmeier et al., 2001), but not Sprague Dawley rats, treated with BDCM at doses of 50 to 100 mg/ kg/ day during gestation days 6 to 10 or with a single dose of 75 mg/ kg/ day on gestation day 9. Based on these observations, it appears that BDCM administered in the drinking water can induced various developmental effects, and the types of these effects are species and strain dependent.

Systemic Effects BDCM causes decreased weight gain and various adverse effects in the nervous and immune systems, thyroid, kidney, and liver, but the predominant systemic effects from acute and chronic exposure to BDCM are on the liver and kidney. NTP ( 1987) administered BDCM to rats by gavage in corn oil at doses of 0, 50 or 100 mg/ kg/ day for 102 weeks. Histologic alterations in the liver and kidney were observed at 50 mg/ kg/ day and higher. In a similar study in mice ( NTP,

1987), histologic alterations in the liver, kidney and thyroid of male mice were noted at 25 74 mg/ kg/ day and higher doses. Aida et al. ( 1992) administered microencapsulated BDCM to rats in the diet at dose levels ranging from 6 to 168 mg/ kg/ day for 24 months. At 6 mg/ kg/ day, liver fatty degeneration and granuloma were observed.

Carcinogenicity BDCM has been found to cause tumors in multiple target organs of multiple species.

Tumors were observed in the large intestine and kidneys of male and female rats, kidneys of male mice and livers of female mice when these rodents were treated with BDCM in corn oil via gavage in a 2 year bioassay ( NTP, 1987). The NTP concluded that under the conditions of these 2 year gavage studies, clear evidence of carcinogenic activity existed for male and female rats and mice. A study investigating the relationship between liver toxicity and tumorigenicity of BDCM concluded that the shape of the dose response curve was different for these two effects ( Melnick et al., 1998). The authors concluded that there does not appear to be a causal relationship between liver toxicity and tumor development for BDCM.

In another rat study ( George et al., 2002), BDCM induced heptocellular adenomas and carcinomas in male rats exposed to the compound via drinking water for 104 weeks. However, a biphasic dose response was observed; the significant increase in hepatocellular neoplasmas only occurred in the low dose group ( 8/ 45 vs. 2/ 45 in the control group) and not in the mid and highdose groups ( 7/ 48 and 4/ 49, respectively). The underlying basis for the biphasic response is unknown, but this pattern of response might be explained by inhibition of the hepatic metabolism of BDCM by the compound itself. The same authors ( George et al., 2002) also reported that BDCM was not carcinogenic to male mice, but there is no evidence that an adequately high dose 75 was tested in this study. DeAngelo ( 2002) reported mixed results on BDCM induction of aberrant crypt foci ( ACF), putative early preneoplastic lesions of the colon. The authors observed increased ACF in a 13 week drinking water rat study, but not in 13 or 30 week mouse studies.

The study authors concluded that ACF induced by BDCM do not progress to neoplasia, as judged by the absence of colon neoplasms in the two year cancer study conducted by George et al.

( 2002).

BDCM has shown mixed results in genotoxicity assays, with both positive ( Fujie et al.,

1990; Simmon and Tardiff, 1978) and negative results ( Hayashi et al., 1988; NTP, 1987; Varma et al., 1988) from in vitro and in vivo studies. Synthesis of the overall weight of evidence from these studies is complicated by the use of a variety of testing protocols, different strains of test organisms, different activating systems, different dose levels, different exposure methods ( gas versus liquid) and, in some cases, problems due to evaporation of the test chemical. Although data are mixed, the weight of evidence indicates that BDCM is genotoxic. Recent studies in Salmonella strains engineered to contain rat GST theta suggest that the mutagenicity of the brominated THMs may be mediated by glutathione conjugation ( DeMarini et al., 1997; Landi et al., 1999). Furthermore, a recent study in female B6C3F1 mice ( Melnick, et al., 1998) suggests that increased incidence of hepatic tumors occurs at doses of BDCM that have no effect on hepatocyte labeling index, indicating that regenerative hyperplasia is not required for tumor induction, supporting the choice of linear low dose extrapolation for quantification of cancer risk associated with BDCM.

Following the EPA's 1986 Guidelines for Carcinogen Risk Assessment, BDCM is classified as Group B2: Probable Human Carcinogen ( EPA, 2002c). Under the EPA's 1999 Draft 76 Guidelines for Carcinogen Risk Assessment, BDCM is likely to be carcinogenic to humans ( EPA,

2002c). This descriptor is considered appropriate when there are no or inadequate data in humans, but the combined experimental evidence demonstrates the production or anticipated production of tumors by modes of action assumed to be relevant to humans. Based on the evidence for BDCM genotoxicity, and the lack of data supporting a mode of action for which nonlinear extrapolation to low doses are appropriate, a linear extrapolation to low doses is used for quantification of the BDCM cancer risk.

Under the 1986 cancer guidelines, the Agency calculated a cancer oral slope factor, based on renal tumors in treated male mice, of 6.2 × 10 2 per mg/ kg/ day and a unit risk of 1.8 x 10 6 ( µ g/ L) 1 ( EPA, 2002c). Based on the same endpoint, a cancer oral slope factor of 8.1 x 10 3 per mg/ kg/ day has been calculated under the Draft Guidelines for Carcinogen Risk Assessment ( EPA,

1999), corresponding to a unit risk of 2.3 x 10 7 ( µ g/ L) 1 ( EPA, 2002c). Based on this oral slope factor, drinking water concentrations of 400 µ g/ L, 40 µ g/ L, and 4 µ g/ L are estimated to be associated with estimated lifetime risks of 10 4 , 10 5, and 10 6.

Basis for RfD and MCLG EPA selected the chronic study by Aida et al. ( 1992) as the most appropriate study for derivation of the RfD. This study identified a LOAEL of 6 mg/ kg/ day based on liver fatty degeneration and granuloma in male rats, as well as a BMD of 1.9 mg/ kg/ day and a corresponding BMDL 10 of 0.8 mg/ kg/ day, based on the same endpoint. The LOAEL of 6.1 mg/ kg/ day was used as the basis for the RfD. An UF of 3000 was applied to the LOAEL: a 10 fold UF for interspecies extrapolation; another 10 fold factor for protection of sensitive human 77 subpopulations; a 10 fold factor for extrapolation from a LOAEL; and a 3 fold factor for database deficiencies, including lack of a multigeneration reproductive toxicity study and uncertainty related to possible reproductive or developmental effects suggested by epidemiological studies ( EPA, 2002c). These calculations result in an RfD of 0.002 mg/ kg/ day. The MCLG for BDCM is zero, based on its probable human carcinogenicity, and linear low dose extrapolation in light of the evidence for a genotoxic mode of action, coupled with the absence of evidence for other modes of action.

Children's Risk in Relation to the MCLG The Agency believes that the MCLG of zero is protective for both the carcinogenic and systemic effects of BDCM in children and adults. There is not sufficient evidence from studies on the systemic effects of BDCM, or on the metabolism of the compound, to determine whether children are more sensitive to the toxic effects of BDCM than are adults.

<IP_ADDRESS>. Dibromochloromethane Developmental/ Reproductive Effects In a prospective study, Windham et al. ( 2003) demonstrated that increasing levels of individual brominated THMs or total brominated THMs in the drinking water were associated with significantly shorter cycles when examined by quartile. Similar decrements were observed in follicular, but not luteal, phase length. The strongest association for an individual compound was observed for DBCM with adjusted decrements of 1.2 days ( 95% C. I.  2.0,  0.38) for mean cycle 78 length and 1.1 days ( 95% C. I.  1.9,  0.25) for mean follicular phase length at the highest quartile (  20 µ g/ L).

Ruddick et al. ( 1983) investigated developmental toxicity in pregnant Sprague Dawley rats ( 9 15/ group) administered DBCM by gavage in corn oil at dose levels of 0, 50, 100, or 200 mg/ kg/ day from gestation days 6 to 15. Although maternal toxicity was indicated by a significant decrease in body weight gain at the highest dose, no treatment related developmental toxicity was observed for DBCM. No statistical analysis was performed in this study and inspection of the data revealed no dose related effects. The power of the experiment was limited by the small number of litters per dose group. Thus, the developmental NOAEL was 200 mg/ kg/ day and a LOAEL could not be identified. Borzelleca and Carchman ( 1982) conducted a two generation reproductive study in ICR Swiss mice. Nine week old mice ( 10 males and 30 females per dose group) were continuously maintained on drinking water containing 0, 100, 1000, or 4000 mg/ L DBCM ( 0, 17, 171, or 685 mg/ kg/ day). Based on postnatal body weight in the F 2 b pups, this study identified a marginal LOAEL of 17 mg/ kg/ day for DBCM, and a NOAEL could not be determined. This LOAEL was considered to be minimal because the body weight decrease was only observed at one time point, the effect was only noted in one of the two litters in the F 2 generation, no other adverse effects were noted at the dose level, and it was unclear from the report how many litters and pups per litter were examined for postnatal body weight.

NTP ( 1996) conducted a short term reproductive toxicity screening study on Sprague

Dawley male and female rats. A group of male rats and two groups of female rats were treated with DBCM in drinking water at concentrations of 0, 50, 150, or 450 mg/ L ( 10 rats/ dose/ group)

during a study period of 35 days ( from gestation day 6 through parturition). Males were treated 79 on study days 6 34. Group A females were treated on study days 1 34, and were mated to treated males on study days 13 18. Group B females were mated on study day 1 to untreated males and treated from gestation day 6 through parturition. Based on measured water consumption, the authors estimated dose levels for the treated males to be 0, 4.2, 12.4, or 28.2 mg/ kg/ day, and for treated females to be 6.3, 17.4, or 46.0 mg/ kg/ day ( Group A) and 7.1, 20.0,

or 47.8 mg/ kg/ day ( Group B). The developmental toxicity of the offspring from these treated rats was compared with those from the control group. No significant reproductive/ developmental toxicity was observed at any dose level. The NOAEL for reproductive/ developmental effects identified in this study was 28.2 mg/ kg/ day for males and 47.8 mg/ kg/ day for females; no LOAEL was identified. The study sensitivity was decreased by the use of a relatively small number of animals per group and the lack of microscopic examination of the pups.

Systemic Effects DBCM causes decreased weight gain and various adverse effects in the nervous and immune systems, kidneys, and liver. The predominant effects from acute and chronic exposure to DBCM are on the liver and kidney. Tobe et al. ( 1982) administered microencapsulated DBCM in the diet to rats for 24 months and reported decreased body weight and changes in clinical chemistry parameters and gross liver appearance in males at 49 mg/ kg/ day; similar effects were seen in females at slightly higher administered doses. Tobe et al. ( 1982) identified a NOAEL of 12 mg/ kg/ day, but this NOAEL is uncertain in light of the lack of histopathological examination and the results of Aida et al. ( 1992) with BDCM, finding that adverse liver histopathology occurs at doses lower than those observed by Tobe et al. ( 1982) based on clinical chemistry and gross 80 tissue examination. NTP ( 1985) reported histologic lesions in the liver in male and female rats at 40 mg/ kg/ day and lesions in the liver and thyroid in female mice at 50 mg/ kg/ day after exposure to DBCM via gavage in oil for 104 weeks; no NOAELs were identified. NTP ( 1985) evaluated the toxicologic effects of DBCM after a subchronic ( 13 week) exposure. Hepatic lesions were identified at 60 mg/ kg/ day, with a subchronic NOAEL 30. Adjusting for exposure 5 days/ week,

the subchronic NOAEL was 21.4 mg/ kg/ day, and the LOAEL was 43 mg/ kg/ day.

Carcinogenicity Evidence is limited for DBCM carcinogenicity via ingestion. An increase in tumors occurred in the livers of male and female mice, while no increase in tumors occurred in male and female rats, when these rodents were treated with DBCM via gavage in corn oil for 2 years ( NTP,

1985). In the mouse study ( NTP, 1985), a significant increase in hepatocellular carcinomas and adenomas occurred in female high dose group ( 19/ 50 vs. 6/ 50). The combined incidence of hepatocellular adenomas or carcinomas in high dose male mice ( 27/ 50 vs. 23/ 50) was also significant, but this effect is considered marginal overall because it was only significant in the life table test but was not significant ( p= 0.065) by the incidental tumor test. The low dose male mouse group in this study was considered to be unsuitable for data analysis because an overdose killed most of the male mice. No significant increase in carcinomas and adenomas was observed in the female low dose group ( 10/ 50 vs. 6/ 50). Based on these data, the NTP concluded that there was equivocal evidence of DBCM carcinogenicity in male B6C3F1 mice, some evidence of carcinogenicity in female B6C3F1 mice, and no evidence of carcinogenicity in male or female rats.

Based on a comparison of the shape of the dose response curve for liver toxicity and 81 tumorigenicity of DBCM, Melnick et al. ( 1998) concluded that there does not appear to be a causal relationship between liver toxicity and tumor development for DBCM. De Angelo ( 2002)

reported that in a 13 week drinking water study, DBCM can induce aberrant crypt foci ( ACF),

putative early preneoplastic lesions, in the colons of male rats. However, it is not clear whether this change would result in colon cancer in exposed rats.

DBCM has shown mixed results in genotoxicity assays, with both positive ( e. g., Sekihashi et al., 2002; Simmon and Tardiff, 1978; Sofuni et al., 1996) and negative results ( e. g., Hayashi et al., 1988; NTP, 1985; Potter et al., 1996) from in vitro and in vivo studies. Synthesis of the overall weight of evidence from these studies is complicated by the use of a variety of testing protocols, different strains of test organisms, different activating systems, different dose levels,

different exposure methods ( gas versus liquid) and, in some cases, problems due to evaporation of the test chemical. EPA ( 1994b) has previously determined that the weight of evidence for DBCM mutagenicity and genotoxicity is inconclusive. Recent studies in Salmonella strains engineered to contain rat GST theta suggest that the mutagenicity of the brominated THMs may be mediated by glutathione conjugation ( DeMarini et al., 1997).

Following the EPA's 1986 Guidelines for Carcinogen Risk Assessment, DBCM is classified as Group C: Possible Human Carcinogen, based on inadequate human data and limited evidence of carcinogenicity in animals ( EPA, 2002c). Based on the 1999 Draft Guidelines for Cancer Risk Assessment, there is suggestive evidence of human carcinogenicity of DBCM, but the data are not sufficient to assess human carcinogenic potential. A compound is described as having suggestive evidence of carcinogenicity when the evidence from human or animal data is suggestive of carcinogenicity, but is judged not sufficient for a conclusion as to human 82 carcinogenic potential ( e. g., a marginal increase in tumors that may be exposure related, or evidence is observed only in a single study). The evidence for DBCM carcinogenicity is weaker than the evidence for the other brominated THMs. However, the evidence for DBCM carcinogenicity is strengthened by the evidence for carcinogenicity of BDCM and bromoform,

both of which are considered likely to be carcinogenic to humans. BDCM, DBCM, and bromoform are all brominated THMs with increasing numbers of bromine atoms, suggesting that DBCM is likely to share toxic effects common to both BDCM and bromoform.

Under the 1986 Guidelines for Carcinogen Risk Assessment, the Agency calculated a cancer oral slope factor of 8.4 × 10 2 per mg/ kg/ day and a unit risk of 2.4 x 10 6 ( µ g/ L) 1 based on liver tumors in female mice dosed via oral gavage in corn oil ( EPA, 2002c). Dose response assessment is not recommended under the 1999 guidelines for chemicals for which the weight of evidence is described as " suggestive evidence of human carcinogenicity, but not sufficient to assess human carcinogenic potential."

Basis for RfD and MCLG EPA selected the NTP subchronic study in rats ( NTP, 1985) as the most appropriate basis for derivation of the RfD and DWEL. This study identified a NOAEL of 30 mg/ kg/ day ( duration adjusted to 21.4 mg/ kg/ day). An UF of 1000 was applied to the subchronic NOAEL: a 10 fold factor for interspecies extrapolation; a 10 fold factor for protection of sensitive human subpopulations; and a 10 fold factor for subchronic to chronic extrapolation ( EPA, 2002c). The resulting RfD was 0.02 mg/ kg/ day, corresponding to a DWEL of 0.7 mg/ L for a 70 kg adult drinking 2 L of water/ day. 83 A relative source contribution ( RSC) of 80% is used for DBCM, based on an analysis of the anticipated potential for exposure to DBCM in disinfected tap water via ingestion, inhalation,

and dermal contact, and because exposure via other media ( outdoor air, food, and soil) are anticipated to be low ( EPA, 2002c). There are some uncertainties in the 80% RSC that are related to the availability of adequate concentration data for DBCM in media other than water.

The MCLG of 0.06 mg/ L for DBCM is based on a noncarcinogenic endpoint ( the RfD)

with an additional safety factor of 10 to account for possible carcinogenicity. Thus, the MCLG is calculated as: MCLG = ( 0.02 mg/ kg/ day × 70 kg × 0.8)/( 2 L/ day × 10) = 0.056 mg/ L, rounded to 0.06 mg/ L.

Children's Risk in Relation to the MCLG Developmental and reproductive toxicity data are available for DBCM and were considered in the derivation of the RfD for DBCM. The NOAEL in the two generation reproductive study of Borzelleca and Carchman ( 1982) ( 17 mg/ kg/ day) is comparable to the duration adjusted NOAEL ( 21.4 mg/ kg/ day) that was the basis for the RfD. There is not sufficient evidence from studies on the systemic effects of DBCM, or on the metabolism of the compound, to determine whether children are more sensitive to the toxic effects of DBCM than are adults. Nevertheless, the Agency believes that the MCLG of 0.06 mg/ L is protective of children's health, because developmental or reproductive effects have not been found to occur below the level of the critical effect ( liver toxicity) used to derive the current RfD, and because development of the MCLG includes the standard UF of 10 for protection of sensitive populations.

There is, however, some uncertainty in this conclusion. 84 <IP_ADDRESS>. Bromoform Developmental/ Reproductive Effects In a prospective study, Windham et al. ( 2003) demonstrated that increasing levels of individual brominated THMs or total brominated THMs in the drinking water were associated with significantly shorter cycles when examined by quartile. Similar decrements were observed in follicular, but not luteal, phase length. For bromoform, the adjusted decrements were 0.79 days ( 95% C. I.  1.4,  0.14) for mean cycle length and 0.78 days ( 95% C. I.  1.4,  0.14) for mean follicular phase length at the highest quartile (  12  g/ L). The strongest association for an individual compound was observed for DBCM.

Ruddick et al. ( 1983) investigated developmental toxicity in pregnant Sprague Dawley rats ( 9  15/ group) administered bromoform by gavage in corn oil at dose levels of 0, 50, 100, or 200 mg/ kg/ day from gestation days 6 15. No maternal toxicity was observed at any dose level;

some fetal skeletal anomalies were observed. Incidences of both fetuses and litters with interparietal deviations were increased at the mid and high dose groups compared with the controls. Furthermore, incidences of both fetuses and litters with sternebra aberrations increased in a dose related fashion. No statistical analysis was performed by the authors, but an independent statistical analysis ( EPA, 2002c, using the Fisher Exact test) demonstrated that the increase in sternebral anomalies was significantly different from controls at 200 mg/ kg/ day. A trend test showed a statistically significant dose related trend ( p< 0.002) for this endpoint;

stepwise analysis indicated that this trend was no longer significant when the two highest doses ( 100 mg/ kg/ day and 200 mg/ kg/ day) were removed from the analysis. These findings suggest that 85 the NOAEL and LOAEL for developmental toxicity from this study were 50 mg/ kg/ day and 100 mg/ kg/ day, respectively. The NOAEL for maternal toxicity was 200 mg/ kg/ day, and a maternal LOAEL could not be determined.

The prenatal and postnatal effects of bromoform on fertility and reproduction were investigated by NTP ( 1989a) in Swiss CD 1 mice using a continuous reproductive breeding protocol. Mice were administered bromoform by gavage in corn oil at dose levels of 0, 50, 100,

or 200 mg/ kg/ day for 105 days, including a seven day pre cohabitation phase and a 98 day cohabitation phase. F 1 litters in the control and 200 mg/ kg/ day groups were raised to sexual maturity ( approximately 74 days) while receiving the same treatment as their parents. At sexual maturity, males and females from different litters within the same treatment group were cohabited for seven days and then housed individually until delivery. In the 200 mg/ kg/ day group, the body weights of the dams at delivery were consistently less than the controls. No effect on any fertility or reproductive parameter ( numbers of litters per pair, litter size, proportion of live pups, sex ratio of live pups, and pup body weight) was observed in the P generation. However, postnatal survival of the F 1 pups in the 200 mg/ kg/ day group was significantly less than in the control group. As for the P generation, there was no effect on any mating, fertility or reproductive parameter in the F 1 generation. At sacrifice, male and female F 1 mice administered 200 mg/ kg/ day exhibited increased relative liver weights and decreased relative kidney weights compared to controls. Histopathological examination revealed minimal to moderate hepatocellular degeneration in the livers of the 200 mg/ kg/ day male and female mice. Therefore,

based on liver histopathology, decreased postnatal survival, and other signs of toxicity ( e. g.,

increased relative liver and decreased relative kidney weights) the developmental NOAEL and 86 LOAEL from this study were 100 mg/ kg/ day and 200 mg/ kg/ day, respectively. The NOAEL and LOAEL for maternal toxicity were also 100 mg/ kg/ day and 200 mg/ kg/ day, respectively, based on consistently decreased body weights of pregnant dams at delivery.

Systemic Effects Bromoform causes decreased weight gain and various adverse effects in the nervous system, kidney, liver, and thyroid, but the predominant effects from acute and chronic exposure to bromoform are on the liver. Tobe et al. ( 1982) administered bromoform microencapsulated in the diet of rats for 24 months and reported gross liver lesions and changes in clinical chemistry parameters in male rats at 90 mg/ kg/ day, but this study was not appropriate as the basis for the RfD, because there was no histopathological analysis. NTP ( 1989b) conducted a chronic oral study in rats and mice ( 50/ sex/ group) administered bromoform at 0, 100, or 200 mg/ kg/ day by gavage in corn oil. This study observed histologic lesions in the liver ( fatty changes and chronic inflammation) at 100 mg/ kg/ day in male and female rats, and fatty changes in the liver of female mice at the same dose. Female mice exposed to 200 mg/ kg/ day also exhibited follicular cell hyperplasia of the thyroid gland. This study identified a LOAEL of 100 mg/ kg/ day based on liver effects in both male and female rats and in female mice; no NOAEL was identified. For male mice,

a NOAEL of 100 mg/ kg/ day was identified.

NTP ( 1989b) also conducted a subchronic oral study in rats and mice. Male and female F344/ N rats ( 10/ sex/ dose) received 0, 12, 25, 50, 100, or 200 mg/ kg/ day bromoform by gavage in oil, 5 days/ week for 13 weeks. Male and female B6C3F1 mice received 0, 25, 50, 100, 200, or 400 mg/ kg/ day bromoform, for the same length of time as the rats. In male rats, a dose 87 dependent increase in the frequency of hepatocellular vacuolation was observed, which reached statistical significance at 50 mg/ kg/ day. This effect was not observed in female rats. In male mice, a dose dependent increase in the number of hepatocellular vacuoles was seen at 200 mg/ kg/ day. Thus, this study identified a NOAEL and LOAEL of 25 mg/ kg/ day and 50 mg/ kg/ day, respectively, based on hepatocellular vacuolation in male rats, and a NOAEL and LOAEL of 100 mg/ kg/ day and 200 mg/ kg/ day, respectively, based on hepatocellular vacuoles in male mice.

Carcinogenicity An increase in tumors occurred in the large intestine in male and female rats, while no statistically significant increase in the incidence of tumors was seen in male or female mice, when these rodents were treated with bromoform via gavage in corn oil ( NTP, 1989b). The NTP concluded that there was some evidence for carcinogenic activity in male rats, clear evidence in female rats, and no evidence in male or female mice. De Angelo ( 2002) reported that in a 13

week drinking water study, bromoform can induce ACF, putative early preneoplastic lesions, in the colons of male rats. However, it is not clear whether this change would result in colon cancer in exposed rats. In a study in which bromoform was administered by intraperitoneal injection in mice, the number of lung tumors per mouse in the mid dose group was significantly elevated over controls ( Theiss et al., 1977).

Bromoform has shown mixed results in mutagenicity assays, with both positive ( Simmon and Tardiff, 1978; Zeiger, 1990) and negative ( Hayashi et al., 1988; Ishidate et al., 1982; NTP,

1989a) results from in vitro and in vivo studies. Synthesis of the overall weight of evidence from 88 these studies is complicated by the use of a variety of testing protocols, different strains of test organisms, different activating systems, different dose levels, different exposure methods ( gas versus liquid) and, in some cases, problems due to evaporation of the test chemical. Although data are mixed, the weight of evidence indicates that bromoform is genotoxic. In addition, recent studies in Salmonella strains engineered to contain rat GST theta suggests that the mutagenicity of the brominated THMs may be mediated by glutathione conjugation ( De Marini et al., 1997;

Landi et al., 1999).

The evidence is sufficient to consider bromoform a Group B2: Probable Human Carcinogen via ingestion under the EPA's 1986 Guidelines for Cancer Risk Assessment ( EPA,

2002c). Based on the observed rare large intestinal tumors observed in both sexes of rats tested and positive genotoxicity results, bromoform can be described under the 1999 Draft Guidelines for Carcinogen Risk Assessment as likely to be carcinogenic to humans ( EPA, 2002c). Based on the evidence for bromoform genotoxicity, and the lack of data supporting a mode of action for which nonlinear extrapolation to low doses is appropriate, a linear extrapolation to low doses is used to quantify the bromoform cancer risk.

The Agency calculated a cancer oral slope factor of 7.9 x 10 3 per mg/ kg/ day and a unit risk of 2.3 x 10 7 per ( µ g/ L) ( EPA, 2002c), based on the incidence of intestinal tumors in rats, as specified in the 1986 Guidelines for Carcinogen Risk Assessment ( EPA, 1986). Calculating a cancer slope factor based on the Draft Guidelines for Carcinogen Risk Assessment ( EPA, 1999),

using the same endpoint results in a value of 4.5 x10 3 ( mg/ kg/ day) 1, corresponding to a unit risk of 1.3 x 10 7 ( µ g/ L) 1 ( EPA, 2002c). Based on this calculation, drinking water concentrations of 89 approximately 800 µ g/ L, 80 µ g/ L and 8 µ g/ L are estimated to be associated with lifetime cancer risks of 10 4, 10 5 and 10 6 , respectively.

Basis for RfD and MCLG EPA derived the RfD for bromoform based on hepatocellular vacuolization in the liver of male rats in the NTP ( 1989b) subchronic study. This study identified a NOAEL of 25 mg/ kg/ day and a LOAEL of 50 mg/ kg/ day. After adjustment for dosing 5 days/ week, the duration adjusted NOAEL and LOAEL were 17.9 and 35.7 mg/ kg/ day, respectively. The duration adjusted BMDL 10 was 2.6 mg/ kg/ day. The subchronic study was preferred to the chronic study as the basis for the RfD. This is because there was much less uncertainty in the estimate of the BMDL 10 for the subchronic study, due to the doses in the subchronic study being closer together, and in light of the high response at the lowest dose of the chronic study. Furthermore, a NOAEL was not identified in the chronic study, and the critical effect was consistent across both the subchronic and chronic studies. Thus, the RfD was based on a NOAEL of 17.9 mg/ kg/ day and a composite UF of 1000: 10 fold factors each for interspecies variation and for protection of sensitive subpopulations; and a 10 fold factor for subchronic to chronic extrapolation. The database for bromoform includes systemic toxicity studies in two species, a two generation study in mice, and a developmental toxicity study in rats. In light of the other data on the THMs, for which the critical effect is systemic toxicity, the single data gap of a developmental toxicity study in a second species is insufficient to require a database UF. The resulting RfD is 0.02 mg/ kg/ day. 90 The MCLG for bromoform is zero, based on its probable human carcinogenicity, and linear low dose extrapolation in light of the evidence for a genotoxic mode of action, coupled with the absence of evidence for other modes of action.

Children's Risk in Relation to the MCLG The Agency believes that the MCLG of zero is protective for both the carcinogenic and systemic effects of bromoform in children and adults. There is no evidence from studies on the systemic effects of bromoform, or on the metabolism of the compound, that children are more sensitive to the toxic effects of bromoform than are adults.

2.3. HALOACETIC ACIDS There are only limited data available for determining children's risk resulting from exposure to haloacetic acids ( HAAs). The haloacetic acids have been tested to varying degrees in developmental and reproductive toxicity studies in animals, as described below for each HAA covered in this document. Human epidemiology data on the developmental and reproductive toxicity of the haloacetic acids are lacking. Most of the human health data for halogenated acetic acids are as components of complex mixtures of water disinfection byproducts. Although most studies of human health effects following exposure to water disinfection byproducts have used total trihalomethanes as the exposure metric, Klotz and Pyrch ( 1999), conducted a case control study on the relationship between neural tube defects and drinking water exposure to trihalomethanes, haloacetonitriles or haloacetic acids. ( See Section 2.1 for more study details.) A statistically significant prevalence odds ratio ( POR) was reported for the highest tertile ( third) of 91 trihalomethane exposure; however, only a slight non statistically significant excess risk ( POR 1.2:

95% confidence interval 0.5 2.6) was found for cases when analyzed based on total haloacetic acids tertiles. The specific haloacetic acids that were measured as part of the total haloacetic acid exposure estimate were not identified. Based on the results of the study, the authors concluded that there was no clear association of HAAs with neural tube defects.

Only a subset of haloacetic acids has been tested in standard cancer bioassays. The chlorinated acetic acids have been tested, but published reports of full cancer bioassays are not yet available for their brominated counterparts. Cancer studies in animals are described separately for each compound below.

The data are insufficient to determine whether there are age dependent differences in the metabolism of the haloacetic acids that might lead to differences in health risk. The enzymes responsible for the metabolism of monochloroacetic acid ( MCA) and monobromoacetic acid ( MBA) have not been identified. The metabolism of the dihaloacetic acids has been more extensively studied. Dichloroacetic acid ( DCA), bromochloroacetic acid ( BCA) and dibromoacetic acid ( DBA) all are metabolized to glyoxylic acid through a reaction catalyzed by glutathione S transferase Zeta ( GST Zeta) ( Tong et al., 1998a; Tong et al., 1998b). Glyoxylic acid is in turn metabolized through a variety of competing pathways to form glycine, glycolate,

oxalate, or CO 2 ( Stacpoole et al., 1998). Since DCA is a potential metabolite of trichloroacetic acid ( TCA) ( Bull 2000; Lash et al., 2000), this enzyme pathway may also be relevant for evaluating the susceptibility of children to TCA. Even in the cases where relevant metabolizing enzymes have been identified, there is no information on age dependent changes in the expression or activity of these enzymes. It is noteworthy, however, that Stacpoole et al. ( 1998) reported that 92 the toxicokinetics of DCA is similar in children and adults given pharmacological doses of this compound, suggesting that there are no significant age dependent differences in DCA metabolism.

The health implications of any differences between children and adults in metabolic capacity are also difficult to determine for the haloacetic acids; neither the mode of action nor toxic forms are known.

2.3.1. Monochloroacetic Acid Developmental/ Reproductive Effects Johnson et al. ( 1998) studied the developmental toxicity of MCA by exposing pregnant Sprague Dawley rats on gestation days 1 22 to 0 or 1570 mg/ L MCA in drinking water ( reported by the authors to be equivalent to doses of 0 or 193 mg/ kg/ day). Dams were sacrificed on gestation day 22, and implantation sites, resorption sites, fetal placements, fetal weights, placental weights, crown rump lengths, gross fetal abnormalities, and abnormal abdominal organs were recorded. In addition, the hearts were evaluated microscopically for abnormalities. There was no microscopic evaluation of other internal anomalies, and no evaluation of skeletal anomalies. All dams survived without evidence of toxicity. Although the authors reported that " weight gain during pregnancy was not significantly different for any group of dams," the average maternal weight gain in the exposed rats was markedly reduced ( 18 g/ dam, compared to 122.1 g/ dam in the controls). The magnitude of the decreased weight gain was confirmed by the study authors ( Johnson, personal communication), but it remains unclear why this difference was not considered statistically significant. No adverse effects on reproductive or developmental endpoints were reported. There was no effect on any other evaluated endpoint of maternal or fetal toxicity. 93 Based on the marked decrease in weight gain, the single dose tested of 193 mg/ kg/ day is a maternal LOAEL and a developmental NOAEL. There was no evidence to suggest any developmental toxicity resulting from exposure to MCA. This lack of effect occurred despite the apparently large differences in maternal weight gain during exposure. Complete fetal examinations were not conducted, and the study is limited by the small size of the exposed group.

Smith et al. ( 1990 in a published abstract) reported on the results of a developmental toxicity study in which pregnant Long Evans rats were treated with MCA at doses of 0, 17, 35,

70, or 140 mg/ kg/ day by gavage on gestation days 6 15. MCA treatment significantly reduced maternal weight gain at 140 mg/ kg/ day but did not affect maternal deaths or organ weight at any doses. No treatment related effects on resorptions/ litter or pup weight were observed. The mean frequency per litter of soft tissue malformations ranged from 1.2% in the control group to 6.4% in the high dose group ( 140 mg/ kg/ day); the change was not considered to be dose related. The highest dose of MCA caused a significantly increased incidence of malformations of the cardiovascular system, mainly levocardia ( primarily a defect between the ascending aorta and right ventricle). There were no skeletal malformations observed in this study. Based on the malformations of the cardiovascular system, the LOAEL for developmental toxicity was 140 mg/ kg/ day, and the next lower dose of 70 mg/ kg/ day can be considered a NOAEL. Maternal toxicity was observed at the LOAEL for developmental toxicity, as evidenced by a significant reduction in maternal weight gain. Complete details of the study methods and results were not published.

MCA has been evaluated in in vitro systems for developmental toxicity. Hunter et al.

( 1996) treated early somite staged conceptuses ( 3 6 somites) from CD 1 mice to MCA 94 concentrations ranging from 0 to 500 µ M. Statistically significant, concentration dependent increases in malformations were seen at sub lethal doses. MCA has also been tested in developmental screening assays in non mammalian systems. The Hydra system is an assay that determines the degree to which a test chemical can perturb embryonic development at maternally subtoxic doses. It has been purposely designed to overestimate developmental hazard potential and its primary utility is as a screening tool for identifying compounds for in vivo developmental toxicity testing ( Fu et al., 1990). This assay system compares the toxicity to adult Hydra and to artificial " embryos" made from disassociated and randomly reaggregated terminally differentiated and pluripotent stem cells. Using this system, Fu et al. ( 1990) studied MCA's developmental toxicity. Based on the ratio of minimal effective toxic concentrations for adults and artificial embryos, they reported that MCA was more than 8 times more toxic to in vitro development than to the adult component of the assay. Ji et al. ( 1998) also studied the teratogenic potential of MCA using Hydra. The teratogenic potential was estimated by the ratio of toxicity to regenerative inhibition ( T 50/ I 50), which yielded a ratio of 6.16. The study authors concluded that MCA had a high teratogenic potential. The in vitro studies using mammalian whole embryo culture and Hydra provide support for the developmental toxicity of MCA observed in vivo ( Smith et al., 1990).

No studies were located on the reproductive toxicity of MCA.

Systemic Effects Effects in several short term MCA toxicity studies included neurotoxicity at high doses near the LD 50. In lower dose studies, increased nasal discharge and lacrimation were observed in 95 rats beginning at 7.5 mg/ kg/ day, but no adverse effects were observed in mice treated at similar doses ( NTP, 1992b). Subchronic and chronic studies suggest that the primary targets for MCAinduced toxicity include the heart, spleen and nasal epithelium. In a 13 week oral gavage study in rats, decreased heart weight was observed at 30 mg/ kg/ day and cardiac lesions progressed in severity with increasing dose. Liver and kidney toxicity were only observed at higher doses ( NTP, 1992b). In 2 year oral gavage studies, decreased survival was observed at 15 mg/ kg/ day in rats and decreased survival and nasal and forestomach hyperplasia were observed in mice at 100 mg/ kg/ day ( NTP, 1992b). A more recent study appears to confirm these findings ( DeAngelo et al., 1997); this study identified a critical effect of increased spleen weight.

Carcinogenicity MCA did not induce a carcinogenic response in two chronic rodent bioassays ( NTP,

1992b; DeAngelo et al., 1997). There was no evidence for carcinogenicity in the NTP ( 1992b)

study; however, the route of compound administration was via gavage, only two doses were tested, and significant mortality was observed in high dose male rats, high dose male mice, and low and high dose female rats. The high mortality rates may have compromised the power and sensitivity of the study to detect MCA associated tumor effects.

In the DeAngelo et al. ( 1997) drinking water study, only male F344 rats were tested.

Female F344 rats may be more sensitive to MCA associated effects, as evidenced by some of the toxicity findings in the subchronic oral gavage study ( NTP, 1992b; Bryant et al., 1992). MCA has not been tested for carcinogenicity in a drinking water assay in a second species. 96 MCA has yielded mixed results in genotoxicity assays. MCA was negative in the S.

typhimurium reverse mutation assay ( Mortelmans et al., 1986) but positive in the mouse lymphoma gene mutation assay in the absence of metabolic activation ( McGregor et al., 1987). It was positive for the induction of sister chromatid exchanges in Chinese hamster ovary cells in the absence of S9 ( liver enzyme preparation containing cytochrome P450) activation ( but not in the presence of S9) ( Galloway et al., 1987), but was negative for sister chromatid exchanges in a Chinese hamster lung fibroblast system ( Sawada et al., 1987). MCA did not induce strand breaks in rats or mice ( Chang et al., 1991). In a more recent study including gene mutation assays in Escherichia coli and Salmonella TA100 and the newt ( Pleurodeles waltl larvae) micronucleus test, MCA gave uniformly negative results ( Giller et. al, 1997).

Following the EPA's 1986 Guidelines for Carcinogen Risk Assessment ( EPA, 1986),

MCA is best classified as Group D: Not Classifiable as to Human Carcinogenicity. According to the 1999 Draft Guidelines for Carcinogen Risk Assessment ( EPA, 1999), the data on MCA are inadequate for an assessment of human carcinogenic potential.

Basis for RfD and MCLG The Agency derived the MCA RfD from the LOAEL of 3.5 mg/ kg/ day for increased spleen weight in the chronic rat drinking water study ( DeAngelo et al., 1997). A composite UF of 1000 was used: a 10 fold factor for interspecies extrapolation, a 10 fold factor for human variability, a 3 fold factor for extrapolation from a minimal LOAEL, and a 3 fold factor for database deficiencies, including lack of adequate developmental toxicity studies in two species and 97 the lack of a multi generation reproductive study. The resulting RfD is 0.004 mg/ kg/ day,

corresponding to a DWEL of 0.14 mg/ L for a 70 kg adult drinking 2 L of water/ day.

EPA calculated the MCLG by applying a relative source contribution ( RSC) of 20% to the DWEL: MCLG = 0.14 mg/ L x 0.2 = 0.028 mg/ L ( rounded to 0.03 mg/ L). The resulting proposed MCLG for MCA is 0.03 mg/ L. The default RSC of 20% was chosen in accordance with the exposure decision tree approach in EPA's Human Health Methodology ( EPA, 2000f),

taking into account the likelihood of exposure to MCA from sources other than tap water, such as ambient air and food. The available data are sufficient to demonstrate that food and air are relevant exposure sources in addition to drinking water, but the data are inadequate to calculate an RSC.

Children's Risk in Relation to the MCLG There is no evidence from studies on the systemic effects of MCA that children are more sensitive to the toxic effects of MCA than are adults. However, these data are limited. No data on potential metabolic differences between children and adults for MCA were located.

The only developmental effects that have been reported for MCA are malformations of the cardiovascular system reported in a published abstract ( Smith et al., 1990). The NOAEL was 70 mg/ kg/ day, and the LOAEL for developmental toxicity was 140 mg/ kg/ day, where maternal toxicity also occurred ( Smith et al., 1990). In contrast, Johnson et al. ( 1998) reported that the single dose tested of 193 mg/ kg/ day induced maternal, but not developmental effects.

Taken together, these data suggest that the developing fetus is no more sensitive to MCA than are adults. The LOAEL for systemic toxicity of 3.5 mg/ kg/ day ( DeAngelo et al., 1997) is 20x lower 98 than the NOAEL for developmental effects of 70 mg/ kg/ day. The MCLG of 0.03 mg/ L for systemic effects is adequately protective of fetuses and children.

2.3.2. Dichloroacetic Acid Developmental/ Reproductive Effects In two studies reported by Smith et al. ( 1992), pregnant Long Evans rats ( approximately 20/ dose) received DCA by gavage on gestation days 6 15 at doses of 0, 900, 1400, 1900, or 2400 mg/ kg/ day ( first study) or 0, 14, 140, or 400 mg/ kg/ day ( second study). Dams were sacrificed on gestation day 20, and both maternal toxicity and fetal toxicity were assessed. Dose related increases in mortality occurred in dams dosed at 1400 mg/ kg/ day and above, body weight gain was significantly reduced at 140 mg/ kg/ day and above, and increased liver weight was observed in all the treated groups. Significant implantation loss occurred at 900 mg/ kg/ day and above, and the number of live fetuses per litter was reduced at 900 mg/ kg/ day and above. Fetal weight and crown rump length were significantly lower at levels of 400 mg/ kg/ day and above. Dose related increases were also reported for external, soft tissue, cardiovascular, urogenital, and orbital malformations in the developing fetuses at doses of 140 and above. The authors identified a developmental NOAEL of 14 mg/ kg/ day and a LOAEL of 140 mg/ kg/ day. Based on increased liver weight, the maternal NOAEL and LOAEL were also 14 and 140 mg/ kg/ day, respectively.

Epstein et al. ( 1992) reported findings from a series of experiments in pregnant Long

Evans rats exposed to DCA by gavage. There were three separate, sequential phases of this study; in each phase, the dams were exposed for a specific 1 to 3 day period during gestation and were sacrificed on gestation day 20. Both maternal and fetal toxicity were assessed, including 99 histological examination of the fetuses. In all three phases of the study, no treatment related maternal toxicity was observed ( based on body weight and organ weight data). In the first phase of the study, dams were exposed to 1900 mg/ kg/ day during gestation days 6 8, 9 11, or 12 15, in order to observe the effects of DCA during specific periods of organogenesis. A decrease in average fetal weight was reported in the dose group exposed on days 6 to 8, but no malformations were reported. In the groups dosed on days 9 11 and 12 15, the mean percentage of cardiac malformations per litter was significantly ( p

0.001) increased. In the second phase of the study, pregnant dams were administered a single dose of 2400 mg/ kg on gestation days 10,

11, 12, or 13. Fetal weights in each exposure group were similar to control values. Significant ( p

0.05) increases in cardiac malformations were reported in groups exposed on day 10 or day 12. In the third phase of the study, a single dose of 3500 mg/ kg was administered to dams on gestation days 9, 10, 11, 12, or 13. This higher dose level, 3500 mg/ kg, resulted in a slightly higher incidence of cardiac defects ( 3.6% vs. 2.9% in controls), and the increase was significant ( p

0.05) on day 12. The results from these studies suggest that acute high dose treatments of DCA at specific developmental stages can induce developmental toxicity in the absence of maternal toxicity.

The developmental toxicity of DCA has also been evaluated in in vitro systems. Saillenfait et al. ( 1995) exposed explanted embryos from Sprague Dawley rats for 46 hours to 0, 1.0, 2.5,

3.5, 5.0, 7.5, or 10 mM DCA. A significant, dose dependent decrease in crown rump length was seen at 3.5 mM and above. In addition, several effects ( brain and eye defects, reduced embryonic axis) were seen at 2.5 mM. A similar study with CD 1 mouse whole embryo culture exposed to DCA for 24 hours found significant increases in neural tube defects at treatment concentrations of 100 5.9 mM and above ( Hunter et al., 1996). Another whole embryo culture study was designed to test mechanisms of haloacetic acid induced dysmorphogenesis ( Ward et al., 2000). The authors reported that DCA treatment increased the accumulation of sub G1 events ( a measure of cells with less than the normal 2n complement of DNA found during the G1 stage of the cell cycle).

Because the characteristic breakage of DNA during apoptosis leads to this accumulation of sub

G1 events, this measure is often used as an indicator of apoptosis. The results suggested to the authors that the developmental toxicity of DCA, particularly the induction of embryonic neuraltube defects, may be mechanistically associated with its ability to increase apoptosis. Bantle et al.

( 1999) reported equivocal results for DCA in the Frog Embryo Teratogenesis Assay  Xenopus ( FETAX). Although these in vitro studies cannot be used for dose response assessment, they support the in vivo observations that DCA can cause developmental effects.

No single or two generation reproductive toxicity studies have been conducted with DCA. However, there are data indicating that DCA induces male reproductive tract toxicity.

Cicmanec et al. ( 1991) observed testicular degenerative changes in beagle dogs exposed to 12.5 mg/ kg/ day DCA in gelatin capsules for 90 days. Testicular changes included syncytial giant cell formation and degeneration of the testicular germinal epithelium. This dose was the study LOAEL, based on the testicular changes, liver vacuolization and brain histopathology. Toth et al.

( 1992) found decreased absolute weight of the preputial gland and epididymis at the lowest dose tested ( 31.3 mg/ kg/ day) in male Long Evans rats ( 18 to 19/ group) treated with gavage doses of DCA for 10 weeks, but the absolute testes weight was not affected. Relative liver weights were also increased in this dose group. At the two higher doses ( 62.5 and 125 mg/ kg/ day), significant decreases in the percentage of motile sperm, epididymal sperm counts and sperm motion 101 parameters were observed. Gross pathologic examination did not reveal lesions at any dose in the testis or epididymal epithelium. Histological evidence of impaired spermiation was noted in the 62.5 and 125 mg/ kg/ day dose groups and was attributed to retention of late step spermatids in the seminiferous tubules. Male fertility was assessed by mating treated males with untreated females on day 70 ( final day of treatment); no effect on fertility was seen at any dose. However, the study authors indicated that the experimental protocol may have been insensitive for fertility evaluation,

because an overnight mating period was used and multiple matings were possible. Based on reduced weight in sexual accessory organs ( preputial gland and epididymis), the LOAEL for this study was 31.3 mg/ kg/ day, and a NOAEL could not be determined.

Linder et al. ( 1997a) examined the reproductive tract toxicity and spermatotoxicity of DCA. Male Sprague Dawley rats ( 8/ group) were given daily gavage doses of 0, 18, 54, 160,

480, or 1440 mg/ kg DCA for 14 days. The only evidence of general toxicity was a statistically significant decrease in final body weight at 480 and 1440 mg/ kg/ day. A variety of male reproductive tract toxicity parameters were affected. Rats in the 54 mg/ kg/ day group exhibited clear histopathological effects on spermiation indicative of spermatotoxicity, which increased in severity with increasing dose. At doses of 160 mg/ kg and higher, the percentage of fused cauda sperm was statistically increased and percent motile sperm were statistically decreased as compared with controls. The characteristic pattern of response included altered spermiation ( including retention of mature sperm), atypical formation and resorption of residual cytoplasm,

abnormal sperm morphology, and decreased motility, all of which increased in magnitude and severity with dose. Epididymis weight was decreased at dose levels of 480 and 1440 mg/ kg/ day.

The LOAEL for this study was 54 mg/ kg/ day, and 18 mg/ kg/ day was an equivocal NOAEL. This 102 determination is based on unequivocal histopathological changes indicative of spermatotoxicity in the male reproductive tract at the LOAEL. The susceptibility of the developing male reproductive tract has not been tested for DCA.

Systemic Effects Stacpoole et al. ( 1998) reviewed the clinical data for DCA. The sodium salt of DCA has been employed experimentally for more than 25 years in clinical medicine and human research.

Over 40 patient years of cumulative DCA exposure have been recorded in more than 50 infants and children with congenital lactic acidosis; several patients were treated with daily oral doses of approximately 25 mg/ kg for up to 5 years. Typical doses range from 25 to 100 mg/ kg, and subsequent clinically adverse findings have been mainly restricted to mild neurological effects,

such as a reduction in anxiety or sedative effects in both adults and children that lasted for several hours. A doses of 50 to 100 mg/ kg, there were a few cases of reversible peripheral neuropathy.

No effects on reproductive organs or reproductive parameters have been reported, although the clinical studies have not specifically targeted evaluation of reproductive toxicity. Limited pharmacokinetic data in children ranging in age from 18 months to 10 years, who were treated with DCA for control of lactic acidosis due to malaria, showed kinetic parameters after a single 50 mg/ kg infusion similar to those obtained in healthy or acidotic adults ( Stacpoole et al., 1998).