A variety of decreases in organ weights were observed that were attributed to general growth retardation. In addition, increases in absolute and relative kidney and liver weights ( of approximately 10%) were observed in the P and F1 males and females. There was no doseresponse in increased kidney weight, although the increase in absolute and relative liver weight was dose related. There was no supporting histopathology in an evaluation of the liver and kidney in 10 rats/ sex/ group, and the study authors did not consider the organ weight changes to be toxicologically significant. Histopathology in the zona glomerulosa of the adrenal cortex in female rats of all DBA exposure groups of both generations was considered to be a physiologic response related to water balance and/ or stress, and not a direct exposure related effect. F1 generation male and female rats had significant increased spleen weights relative to terminal body weights.

An increase in the incidence and intensity of extramedullary hematopoiesis in the red pulp of the spleen occurred in the F1 generation female rats in the 650 ppm group and may have been treatment related. Decreased cellularity of the cortical lymphoid area of the thymus was noted in P generation females in the two highest dose groups.

The parental NOAEL for general toxicity is 50 ppm, based on increase in absolute and relative liver and kidney weights. Based on testicular histomorphology indicative of abnormal spermatogenesis in P and F1 males, the reproductive/ developmental toxicity LOAEL and NOAEL are 250 and 50 ppm, respectively. For the P generation, these drinking water concentrations Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 64 Draft, do not cite or quote correspond to a LOAEL and NOAEL of 22 and 4 mg/ kg/ day, respectively. For the F1 generation,

mean daily doses are considered to be equivalent to the mean of average consumed doses during the period from weaning to termination of the study. These doses are very similar to those for the P generation; resulting in a LOAEL and NOAEL for the F1 generation of 22.0 and 4.5 mg/ kg/ day, respectively, equivalent to drinking water concentrations of 250 and 50 ppm in drinking water, respectively.

The developmental toxicity of DBA has been reported in two related abstracts. Narotsky et al. ( 1996) studied the developmental toxicity of DBA in CD 1 mice dosed by gavage with 0,

0.11, 0.23, 0.46, 0.92, 1.8, 2.8, or 3.7 mmol/ kg/ day ( equivalent to 0, 24, 50, 100, 200, 392, 610,

and 806 mg/ kg/ day) on GD 6 15. Mice were allowed to deliver naturally and the litters were examined on postnatal days ( PND) 1 and 6. Maternal effects were limited to piloerection and motor depression at the highest dose tested. Parturition was delayed at all doses tested but the toxicologic significance of this effect is unclear. In the highest dose group ( 806 mg/ kg/ day),

prenatal mortality was increased, and only 3/ 9 litters were viable at birth. Increased postnatal mortality was seen at 610 and 806 mg/ kg/ day. Decreased pup weight was observed at 806 mg/ kg/ day on PND 1 and at 610 mg/ kg/ day on PND 6. Skeletal malformations, as indicated by short, kinked, or absent tails, were in the two highest dose groups. Based on these results, the authors concluded that DBA was a developmental toxicant. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 65 Draft, do not cite or quote In a second published abstract, DBA was administered to CD 1 mice by gavage in distilled water on GD 6 15 at doses of 0, 50, 100, or 400 mg/ kg/ day ( Narotsky et al., 1997). Maternal toxicity was not observed. Litters were removed by cesarean section on GD 17, and half of the fetuses in each litter were examined for skeletal defects and the other half for soft tissue malformations. There were no effects on prenatal survival, fetal weight, and skeletal development.

Hydronephrosis was noted at 100 and 400 mg/ kg/ day, and renal agenesis ( small kidneys) was observed at 400 mg/ kg/ day. In contrast to the Narotsky et al. ( 1996) abstract, which reported delayed parturition at 24 mg/ kg/ day and above, the second abstract showed no developmentallyadverse effects at the 50 mg/ kg/ day dose. Based on the summary data provided in these two abstracts, the LOAEL for fetal kidney malformations would be 100 mg/ kg/ day, with a corresponding NOAEL of 50 mg/ kg/ day. However, due to the limited data provided in the abstracts, the use of the reported adverse effect levels for quantitative risk assessment is not appropriate.

Klinefelter et al. ( 2000), in an abstract, reported the effects of DBA administered in drinking water on the pubertal development and adult reproductive function of male Sprague

Dawley rats ( 3 litters/ dose) exposed from GD 15 to PND 98. Pregnant and lactating dams were exposed to 0, 400, 600, or 800 ppm DBA in drinking water, equivalent to 0, 50, 75, and 100 mg/ kg/ day ( personal communication with authors). After weaning, male offspring were exposed to the same concentrations of DBA in drinking water and sacrificed on PND 98. Histologic examination of the reproductive tract was performed on one half of the sacrificed animals; the Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 66 Draft, do not cite or quote other half was used for harvesting of proximal cauda epididymis sperm for artificial insemination of LHRH synchronized females. Decreased body weight throughout the reproductivedevelopment period was observed in the high dose male offspring as compared with control animals. Decreased epididymis weight ( the percent decrease was not specified) occurred in the 75 and 100 mg/ kg/ day groups. The age at preputial separation was delayed in all treatment groups,

averaging 49, 48, and 50 days for the 50, 75, and 100 mg/ kg/ day dose groups, respectively, as compared with 42 days in controls. Histopathologic examination revealed the presence of only Sertoli cells in the seminiferous tubules of animals in all dose groups. The fertility of treated males was also affected by DBA treatment. The number of implants per corpora lutea in females artificially inseminated with sperm from treated males decreased from 70% for controls to 49%,

15%, and 15% for the 50, 75, and 100 mg/ kg/ day dose groups, respectively. Levels of the sperm protein SP22, which has been shown to be highly correlated with rodent fertility, were significantly decreased in all treatment groups. Based on adverse effects on the fertility of sperm of treated males, the lowest dose tested, 50 mg/ kg/ day, would be the LOAEL, and a NOAEL could not be determined. However, due to the limited data provided in this abstract, the use of the reported adverse effect levels for quantitative risk assessment is not appropriate. Further,

according to the study authors, a more comprehensive study using lower doses is being conducted to identify the NOAEL/ LOAEL boundary ( personal communication).

In a second recent abstract, Veeramachaneni et al. ( 2000) exposed male Dutch belted rabbits ( 10/ group) to DBA treated drinking water from GD15 throughout life. The average daily Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 67 Draft, do not cite or quote doses were reported as 0, 0.97, 5.05, and 54.2 mg/ kg/ day. The ability of the treated males to ejaculate was determined by collecting ejaculates every 3 4 days, beginning at 20 weeks of age.

One male in each of the 0.97 and 54.2 mg/ kg/ day dose groups consistently failed to ejaculate, and one male in each of the 5.05 and 54.2 mg/ kg/ day dose groups failed to ejaculate at least once. In the 54.2 mg/ kg/ day dose group, males that did ejaculate took more attempts and longer time to ejaculate compared to controls ( p< 0.05). The fertility of sperm from 24 week old males was assessed by artificial insemination of two 6 month old rabbit females per sample of sperm from each male. Conception rates were significantly decreased ( p< 0.01) in females inseminated with sperm from males in all treated groups, averaging 85%, 55%, 65%, and 55% for rabbit does inseminated with sperm from males treated with 0, 0.97, 5.05, and 54.2 mg/ kg/ day DBA,

respectively. Of the 53 pups born to females inseminated with sperm from the high dose males, 1 pup had cleft palate and cranioschisis, and 2 pups had cranioschisis. At 25 weeks, the offspring were necropsied; no differences in body weight, anogenital distance, or sex organ weights were reported relative to controls. These abstract data suggest that the lowest dose tested, 0.97 mg/ kg/ day was a LOAEL for decreased male fertility and that a NOAEL could not be determined.

However, a critical assessment of these findings cannot be conducted without a full review of the study report.

Taken together, the data provide strong evidence that DBA is a male reproductive system toxicant following oral dosing. The gavage studies of Linder and colleagues reported perturbation of spermatogenesis based on histopathology changes in seminiferous tubule staging, changes in Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 68 Draft, do not cite or quote sperm quality ( count, morphology, and motility), and in male reproductive performance ( Linder et al., 1994a; Linder et al., 1994b; Linder et al., 1995; Linder et al. 1997a). Administration of DBA in drinking water has also been reported to adversely affect both sperm quality and male reproductive performance in young males exposed continuously during gestation, lactation, and the post weaning developmental period ( Klinefelter et al., 2000, Veeramachaneni et al., 2000). In contrast, although the two generation drinking water reproductive toxicity study ( Chlorine Chemistry Council, 2001; Christian et al., 2002) reported testicular histomorphology indicative of abnormal spermatogenesis similar to that found in shorter term studies by Linder and colleagues,

no adverse treatment related effects on mating performance, fertility, gestation length, and other functional indices of successful reproductive behavior were noted at mean paternal ( P generation)

daily doses up to 52 mg/ kg/ day, and at mean F1 daily doses up to 55 mg/ kg/ day.

No effect on female reproductive success was reported in Holtzman rats administered DBA doses up to 250 mg/ kg/ day by gavage through days 1  8 of pregnancy ( Cummings and Hedge, 1998). Female CD 1 mice given gavage doses up to 801 mg/ kg/ day on GD 6 15 had decreases in viable litters, increased postnatal mortality, decreased pup weight, and increased tail abnormalities ( Narotsky et al., 1996). In a second study examining the incidence of skeletal and visceral malformations in the pups of pregnant CD 1 mice administered DBA gavage doses of up to 400 mg/ kg/ day, an increased incidence in renal malformations ( Narotsky et al., 1997) was reported beginning at 100 mg/ kg/ day. Delayed parturition was noted at all doses in the first, but not the second, study ( Narotsky et al., 1996, 1997); however, details were not reported in the Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 69 Draft, do not cite or quote abstracts and the adversity of this endpoint is unclear. In contrast, no treatment related effects on litter viability, postnatal mortality, and gross malformations were observed in the two generation drinking water reproductive/ developmental toxicity study ( Chlorine Chemistry Council, 2001;

Christian et al., 2002); as previously noted, the significant decreases in the body weight gain in pups of both the P and F1 generation were attributed to a general retardation in growth associated with decreased water consumption ( due to taste aversion) and reduced food consumption, not to a direct effect of DBA treatment. Differences in findings between the two generation study and those by Narotsky et al. ( 1996, 1997) may have been due to differences in internal doses associated with gavage versus drinking water DBA administration, species differences in susceptibility to DBA toxicity, and/ or the lower mean doses tested in the two generation study.

Gardner and Toussant ( 1999) evaluated developmental toxicity of DBA in the frog embryo teratogenesis assay  Xenopus ( FETAX) ( a 96 hour toxicity test), with and without metabolic activation. Endpoints evaluated were embryolethality ( LC 50), embyronic malformations ( EC 50), minimum concentration to inhibit growth ( MCIG), and a teratogenicity index ( TI   the ratio of the LC 50 to the EC 50). The FETAX assay is considered to be a reliable developmental toxicity screening assay; Dawson and Bantle ( 1987) have estimated that its predictive accuracy for identifying known mammalian or human developmental toxicants approaches or exceeds 85%.

At DBA concentrations of up to 12,800 mg/ L, neither 50% mortality nor 50% malformations was achieved in two of the three tests conducted without added metabolic activation; therefore,

neither the LC 50 nor the EC 50 could be estimated. In the third test, the LC 50 and EC 50 without Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 70 Draft, do not cite or quote metabolic activation were 7,354 and 11,723 mg/ L, respectively. With metabolic activation, the LC 50' s for three tests were 6,244, 69, and 3,787 mg/ L; the reasons for the low LC 50 in the second test were unclear and a pooled estimate was not calculated. The EC 50, estimated for one test only,

was 879 mg/ L. The TI was also calculated for only one test, and was 0.6 with metabolic activation and 0.1 without metabolic activation. TI values of > 1.5 suggest teratogenic potential;

therefore, under the conditions of this study, DBA did not exhibit teratogenic potential. Further,

malformations did not appear to increase in severity or prevalence with increasing DBA concentrations, with or without metabolic activation.

No data were identified for the reproductive or developmental toxicity of DBA following exposure by the inhalation or dermal route.

D. Mutagenicity and Genotoxicity Monobromoacetic acid MBA induced a positive mutagenic response in Salmonella typhimurium in the standard assay system ( NTP, 2000a). Detailed results including the tester strains evaluated and microsomal dependence of the mutagenic response were not available from the posted testing results. Giller et al. ( 1997) evaluated the mutagenicity of a series of halogenated acetic acids, including monochloro, dichloro, trichloro, monobromo, dibromo, and tribromoacetic acids in Salmonella Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 71 Draft, do not cite or quote typhimurium strain TA100 in the Ames fluctuation test. This assay is a modification of the Ames test in which bacteria are exposed to the compound under study in a liquid suspension. Rather than determining the number of mutant colonies, the fluctuation assay identifies the presence of mutants based on a change in color of the liquid medium in wells containing prototrophic mutants.

MBA was tested at concentrations of 0.03 to 30 µ g/ mL without S9 activation, and at 0.3 to 300 µ g/ mL in the presence of S9 activation. No mutagenic effect was detected in the absence of S9 activation. The study authors indicated that 10 µ g/ mL was the minimal cytotoxic dose in the absence of S9 activation. In the presence of S9 activation, mutagenic activity was observed at concentrations ranging from 20 to 75 µ g/ mL. The decrease in positive mutagenic responses at the high doses ( with S9 metabolic activation) was consistent with the onset of cytotoxicity at 100 µ g/ mL.

Similar results were reported in a published abstract by Kohan et al. ( 1998), who tested the mutagenicity of the same series of halogenated acetic acids as Giller et al. ( 1997). S.

typhimurium tester strains TA98 and TA100 were incubated with MBA, with or without S9 activation, in a microsuspension assay. MBA ( 0.1 µ mole) induced a positive mutagenic response in both strains + S9 at subtoxic concentrations ( personal communication). Other than DBA ( as described below), none of the other halogenated acetic acids induced a positive mutagenic response when tested up to cytotoxic concentrations. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 72 Draft, do not cite or quote Several measures of DNA damage response have been reported for MBA. Giller et al.

( 1997) evaluated DNA repair responses to MBA using the SOS chromotest, which measures the induction of DNA repair. In this assay, Escherichia coli strain PQ37 was exposed to concentrations ranging from 1 to 1000 µ g/ mL MBA without metabolic activation, and from 3 to 3000 µ g/ mL with metabolic activation by S9 mix. Toxic concentrations were 300 µ g/ mL and higher, regardless of S9 activation. MBA failed to induce the DNA repair response at any of the concentrations tested, regardless of metabolic activation.

Giller et al. ( 1997) also evaluated chromosome damage using a newt micronucleus test.

Pleurodeles waltl larvae were exposed to varying concentrations of MBA in the absence of S9 for 12 days. The highest concentration tested in the assay was half the minimum concentration that led to detectable physiological disturbances in a preliminary test. Fifteen larvae per dose group were exposed to 10, 20, or 40 µ g/ mL MBA ( renewed daily) and the number of micronucleated erythrocytes in a sample of 1000 erythrocytes was determined. MBA did not increase the number of micronuclei at any of the tested concentrations.

Stratton et al. ( 1981) reported that MBA concentrations of 100 µ M ( 13.9 mg/ L) induced DNA strand breaks in L 1210 mouse leukemia cells as measured in an alkaline elution assay.

MBA was added to the cell culture medium and the cells were incubated in the treated medium for 1 hour in the absence of S9. The cells were rinsed and harvested immediately or incubated in MBA free medium for 1 or 6 hours before measuring DNA strand breaks. The number of DNAstrand breaks was increased compared to controls immediately after the 1 hour treatment, and Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 73 Draft, do not cite or quote increased even further following post treatment incubations. The study authors suggested that the observed increase in DNA strand breaks is consistent with depurination of alkylated DNA over time to form alkali labile apurinic DNA sites, suggesting that MBA can induce direct DNA damage.

The results of genotoxicity studies for MBA are summarized in Table V 12. Based on these limited studies, it remains unclear if MBA is genotoxic. A positive mutagenic response was observed in the Ames assays. In the single study with sufficient detail for full evaluation, the positive finding only with S9 activation suggests metabolic activation of MBA to a genotoxic form, but there have been no metabolism studies to identify potentially mutagenic metabolites.

The observed effect in that study is unlikely to be due to altered pH, since the mutagenicity was observed in the absence of cytotoxicity. Although MBA was mutagenic in the Ames assay, it did not induce a DNA repair response in the SOS chromotest. In addition, measures of DNA damage,

including micronuclei and DNA strand breaks, have yielded inconsistent results. Taken together,

these data are not sufficient to conclude that MBA is genotoxic. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 74 Draft, do not cite or quote Table V 12. Genotoxicity Studies of MBA Endpoint Assay system Results ( wo/ w activation) Comments Reference Gene mutationbacteria Salmonella typhimurium + Detailed results were not available NTP, 2000a Salmonella typhimurium TA100 /+ Tested to cytotoxic doses in Ames fluctuation protocol Giller et al., 1997 Salmonella typhimurium TA98, TA100 /+ Positive in TA98 and TA100 in suspension assay. Data provided in a published abstract Kohan et al., 1998 Clastogenicity Micronuclei in Pleurodeles waltl larvae ( Newt) erythrocytes / NT None Giller et al., 1997 DNA damage Escherichia coli strain PQ37 SOS chromotest / Tested to cytotoxic doses Giller et al., 1997 Single strand breaks in L 1210 mouse leukemia cells +/ NT MBA was shown to be cytotoxic to L 1210 cells from 50

M. Stratton et al., 1981 NT = Not tested Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 75 Draft, do not cite or quote Bromochloroacetic acid BCA induced a positive mutagenic response in Salmonella typhimurium in the standard assay system ( NTP, 2000b). Detailed results, including the tester strains evaluated and microsomal dependence of the mutagenic response, were not available from the posted testing results, but should be released upon completion of the full cancer bioassay. Two related studies evaluated the ability of BCA to induce oxidative DNA damage. Austin et al. ( 1996) investigated the hypothesis that compounds that induce lipid peroxidation might show increased potential as genotoxic agents. The capacity of a series of haloacetic acids, including BCA and DBA ( DBA results described below), to induce lipid peroxidation was measured by an increase in production of thiobarbituric acid reactive substances ( TBARS) in the liver. As an indicator of genotoxicity,

oxidative DNA damage in the liver was measured by an increase in 8 hydroxydeoxyguanosine ( 8

OHdG) levels. Male B6C3F1 mice ( number per group varied from 3 to 6) were exposed to single oral doses of 0, 30, 100, or 300 mg/ kg BCA by gavage in distilled water. In a time course experiment, mice were given 300 mg/ kg BCA and livers were harvested at 1, 3, 5, 7, 9, and 12 hours after dosing for measurement of liver TBARS. TBARS levels peaked at 3 hours after dosing and reached levels approximately 5 fold greater than background. TBARS levels returned to pre exposure levels between 7 and 9 hours after dosing. BCA at 300 mg/ kg increased 8 OHdG levels to a maximum of 2 to 3 fold above controls over a period of 12 hours. Both TBARS and 8 OHdG levels increased with increasing dose when measured at 3 hours, and were maximal at Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 76 Draft, do not cite or quote the highest dose tested. For both TBARS and 8 OHdG, the increases were significant ( p< 0.05)

beginning at 30 mg/ kg.

Parrish et al. ( 1996) evaluated whether the ability of brominated acetic acids to induce oxidative stress responses was due to peroxisome proliferation. The effects of BCA on oxidative DNA damage and peroxisome proliferation were measured in the livers of male B6C3F1 mice.

The animals ( 6/ treatment group) were given drinking water containing 0, 100, 500, or 2000 mg/ L BCA for 3 weeks. The approximate corresponding doses, calculated using the default waterintake value of 0.25 L/ kg/ day ( U. S. EPA, 1988), are 25, 125, and 500 mg/ kg/ day. No doserelated change in body weight was observed, but absolute and relative liver weight increased at the high dose. Two responses indicative of peroxisome proliferation ( increased cyanide insensitive Acyl CoA oxidase activity and increased 12 hydroxylation of lauric acid) were also studied,

because peroxisome proliferation has been linked with the hepatocarcinogenic effect of trichloroacetate. An additional dose group exposed to 3000 mg/ L BCA ( 750 mg/ kg/ day) was evaluated for the Acyl CoA activity measurements. BCA had no effect on either measure of peroxisome proliferation after exposures up to 3000 mg/ L. BCA did induce oxidative DNA damage, with 8 OHdG levels in nuclear DNA of the liver significantly increased ( p< 0.05)

beginning at the lowest dose, 25 mg/ kg/ day. The level of 8 OHdG increased to a maximum of approximately 2 fold at the highest dose ( 500 mg/ kg/ day). It is not clear at this time whether the parent BCA or one or more of BCA metabolites is responsible for the observed increase in oxidative stress. The lack of correlation of 8 OHdG levels with Acyl CoA activity or 12 Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 77 Draft, do not cite or quote hydroxylation of lauric acid suggests that peroxisome proliferation is not causally associated with BCA induced oxidative stress.

The results of Austin et al. ( 1996) and Parrish et al. ( 1996) do not provide evidence of a direct genotoxic effect of BCA, although these results coupled with the positive results in the Ames assay suggest that BCA induced oxidative stress might result in downstream genotoxicity through oxidative DNA damage.

Dibromoacetic acid DBA induced a positive mutagenic response in Salmonella typhimurium in the standard assay system ( NTP, 2000c). Detailed results, including the tester strains evaluated and microsomal dependence of the mutagenic response, were not available from the posted testing results, but should be released upon completion of the full cancer bioassay. Giller et al. ( 1997)

evaluated mutagenicity of a series of halogenated acetic acids, including DBA, in the Ames fluctuation test as described previously for MBA. DBA was tested at concentrations ranging from 3 to 3000 µ g/ mL without S9 activation, and from 10 to10,000 µ g/ mL in the presence of S9 fraction. Toxic concentrations were 1000 µ g/ mL without, and 10,000 µ g/ mL with, metabolic activation. Genotoxicity of DBA was detected at concentrations of 10 to 750 µ g/ mL without activation, and at 30 to 3000 µ g/ mL with activation. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 78 Draft, do not cite or quote Similar results were reported for the microsuspension Ames assay of DBA reported in a published abstract by Kohan et al. ( 1998). DBA ( 2.0 µ mole) induced a positive mutagenic response in strains TA98 and TA100 + S9 at subtoxic concentrations ( personal communication).

The absence of a positive response without S9 activation contrasts with the report of Giller et al.

( 1997).

Saito et al. ( 1995) analyzed indoor swimming pool water from four pools for the presence of trace halogenated contaminants, and also for mutagenicity using the Ames Salmonella typhimurium assay with strains TA98 and TA100, with and without metabolic activation. As part of the study, the mutagenicity of DBA ( reported 90% purity) was also investigated in strains TA 98 and TA100. DBA was mutagenic with and without metabolic activation in strain TA100, at a minimum positive concentration of 640 µ g/ plate in each assay. No mutagenic activity was identified in strain TA98.

The ability of DBA to induce DNA repair responses has been evaluated in two separate reports. Giller et al. ( 1997) tested the ability of DBA to induce DNA damage using the SOS chromotest as described previously for MBA. E. coli strain PQ37 was exposed to 10 to 10,000 µ g/ mL DBA without metabolic activation, and to 3 to10,000 µ g/ mL with S9 metabolic activation.

Toxic concentrations were reported as 1000 µ g/ mL without and 10,000 µ g/ mL with S9 activation. DBA induced a positive response regardless of metabolic activation. The concentrations that induced DNA repair were 200 to 750 µ g/ mL without activation, and 100 to 3000 µ g/ mL with activation. Mayer et al. ( 1996) presented a scheme for the concentration and Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 79 Draft, do not cite or quote analysis of water samples for trace analytes, and coupled it with the umu Microtest, which measures induction of DNA repair. DBA was positive in the umu Microtest, with and without metabolic activation. Details of this report were in German, and were not available in English for a more thorough review.

As described above for BCA, Austin et al. ( 1996) tested the capacity of a series of haloacetic acids, including BCA and DBA, to induce lipid peroxidation and oxidative DNA damage. Male B6C3F1 mice were exposed to single oral doses of 0, 30, 100, or 300 mg/ kg DBA by gavage in distilled water. In a time course study, mice were given 300 mg/ kg DBA, and livers were harvested at 1, 3, 5, 7, 9, and 12 hours after dosing for measurement of liver TBARS.

TBARS levels peaked rapidly, 1 hour after dosing for DBA, to levels approximately 5 fold greater than background, and returned to pre exposure levels between 7 and 9 hours after dosing.

DBA at 300 mg/ kg rapidly increased 8 OHdG levels 2 to 3 fold. In contrast to TBARS, the increase in 8 OHdG levels was sustained over a 12 hour period. Both TBARS and 8 OHdG levels, when measured at 1 hour, increased with increasing dose and were maximal at 300 mg/ kg,

the highest dose tested. For TBARS, the increases were significant ( p< 0.05) beginning at 300 mg/ kg, and increases in 8 OHdG levels were significantly greater than controls beginning at 30 mg/ kg.

Parrish et al. ( 1996) tested whether the ability of brominated acetic acids to induce oxidative stress responses was due to peroxisome proliferation. The effects of DBA on oxidative DNA damage and peroxisome proliferation were measured in the livers of male B6C3F1 mice. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 80 Draft, do not cite or quote The animals ( 6/ group) were given drinking water containing 100, 500, and 2000 mg/ L DBA. The approximate doses calculated from a default water intake value of 0.25 L/ kg/ day are 25, 125, and 500 mg/ kg/ day ( U. S. EPA, 1988). No dose related change in body weight was observed, but absolute and relative liver weight increased at the mid and high dose for DBA. Two responses indicative of peroxisome proliferation ( increased cyanide insensitive Acyl CoA oxidase activity and increased 12 hydroxylation of lauric acid) were also studied, because peroxisome proliferation has been linked with the hepatocarcinogenic effect of trichloroacetate. An additional dose group exposed to 3000 mg/ L DBA ( 750 mg/ kg/ day) was evaluated for the Acyl CoA activity measurements. DBA induced Acyl CoA activity to a maximum of 3 fold after exposures up to 3000 mg/ L, but did not induce the 12 hydroxylation of lauric acid. The study authors did not explain the inconsistency in the different responses obtained with these two measures of peroxisome proliferation. DBA induced oxidative DNA damage, with 8 OHdG levels in hepatic nuclear DNA significantly increased ( p< 0.05) at the highest dose ( 500 mg/ kg/ day) to a maximum of approximately twice the control response. The overall lack of correlation of 8 OHdG levels with Acyl CoA activity or 12 hydroxylation of lauric acid suggests that peroxisome proliferation is not causally associated with the oxidative stress induced by brominated acetic acids.

Effects of DBA have also been evaluated at the chromosome level in one study. Giller et al. ( 1997) conducted the newt micronucleus test for DBA, as described previously for MBA.

None of the DBA concentrations that were tested ( 20, 40, or 80 µ g/ mL, in the absence of S9)

significantly increased the number of erythrocytes with micronuclei. The co clastogenic effects of Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 81 Draft, do not cite or quote various water pollutants on chromosomal aberrations induced by mitomycin C in various mammalian cell lines were reported by Sasaki and Kinae ( 1995). The primary focus of the report was on the co clastogenic effects of toxic metals such as lead and mercury; however, some organic chemicals were also tested. Post treatment with DBA ( microsomal activation status not available in the English summary) at concentrations up to approximately 15 µ g/ mL resulted in a strong dose related increase in chromosomal aberrations induced by mitomycin C. Details of this report ( including a complete description of the test system) were in Japanese and were not available in English for further review.

Of the brominated acetic acids, the database for DBA is most complete, as summarized in Table V 13. DBA has provided nearly uniformly positive results in the assays tested. The positive effects have been reported regardless of S9 activation, suggesting that mutagenicity is independent of metabolism by cytochrome P450s, similar to DCA whose metabolism does not involve microsomal activation but is mediated by NADPH and GSH ( Lipscomb et al., 1995;

Cornett et al., 1997; Stacpoole, 1998). DNA damage secondary to generation of oxidative stress has been reported by Austin et al. ( 1996), and is likely to be independent of peroxisome proliferation ( Parrish et al., 1996). The induction of DNA damage responses, including SOS repair system ( Giller et al., 1997) and the umu microtest ( Mayer et al., 1996), supports the potential mutagenicity of DBA. On the other hand, no induction of micronucleated erythrocytes was reported ( Giller et al., 1997), suggesting that DBA was not clastogenic in the newt test system. No evaluation of micronuclei has been reported in the more standard mouse micronucleus Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 82 Draft, do not cite or quote assay. The clastogenicity of DBA has not been reported in other assays using a standard protocol,

but DBA has been reported to be co clastogenic ( Sasaki and Kinae, 1995). As a whole, these data support the conclusion that DBA is mutagenic and genotoxic, although the nature of the DNA damage induced by DBA remains unclear.

Table V 14 provides a summary of the genotoxicity data for MBA, BCA, and DBA. The data are inadequate for determining whether MBA or BCA are genotoxic, but suggest that DBA is genotoxic. The mechanism by which these different brominated acetic acids might lead to DNA damage is not clear from these data. The mutagenicity of MBA, but not DBA, might be metabolism dependent. The data are very sparse for BCA, but for the single endpoint evaluated,

BCA and DBA shared the ability to induce oxidative DNA damage. Thus, this mechanism remains a viable explanation for the onset of DNA damage, and perhaps mutagenicity of DBA and BCA. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 83 Draft, do not cite or quote Table V 13. Genotoxicity Studies of DBA Endpoint Assay system Results ( wo/ w activation) Comments Reference In vitro assays Gene mutationbacteria Salmonella typhimurium + Detailed results were not available NTP, 2000c Salmonella typhimurium TA100 +/+ Tested to cytotoxic doses in Ames fluctuation protocol Giller et al., 1997 Salmonella typhimurium TA98, TA100 /+ Positive in TA98 and TA100 in suspension assay. Data provided in a published abstract Kohan et al., 1998 Salmonella typhimurium TA98 and TA100 +/+ Positive in TA 100. Reference in Japanese; only study summary in English was reviewed. Saito et al., 1995 Clastogenicity Micronuclei in Pleurodeles waltl larvae ( newt) erythrocytes / NT None Giller et al., 1997 Not specified. + Co clastogenic with mitomycin C. Reference in Japanese; only study summary in English was reviewed. Sasaki and Kinae, 1995 DNA damage Escherichia coli strain PQ37 SOS chromotest +/+ Tested to cytotoxic doses Giller et al., 1997 umu Microtest +/+ Reference in German; only study summary in English was reviewed. Mayer et al., 1996 In vivo assays DNA damage Oxidative DNA damage mouse liver in vivo + 2 to 3 fold induction in 8OHdG levels Austin et al., 1996 Oxidative DNA damage mouse liver in vivo + 2 to 3 fold induction in 8OHdG levels Parrish et al., 1996 Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 84 Draft, do not cite or quote Table V 14. Summary of Genotoxicity Data for Brominated Acetic Acids MBA BCA DBA Assay  S9 + S9  S9 + S9  S9 + S9 Mutagenicity  + + ± a +

SOS DNA repair induction   NTb NT + +

DNA damage ± + +

Chromosome damage  NT ± a. Mixed or equivocal results are denoted with a ± .

b. NT = not tested.

E. Carcinogenicity Concern for the potential carcinogenic hazard of the brominated acetic acids is based on the tumorigenicity of chlorinated acetic acids observed in rodent cancer bioassays ( Boorman et al., 1999). Carcinogenicity testing data for the brominated acetic acids are limited to results reported in published abstracts, although both BCA and DBA have been slated for complete 2

year cancer bioassays ( NTP, 2000b; NTP, 2000c).

So and Bull ( 1995) reported in a published abstract that DBA increased the numbers of aberrant crypt foci in the colon of F344 rats. Male rats were administered an initiating dose of azoxymethane and were exposed to 1000 mg/ L DBA in drinking water for up to 20 weeks. The number of aberrant crypt foci and the complexity of the foci were increased in the animals given DBA as compared to animals given the initiating compound only. These findings may be of particular significance because the colon has been implicated as a potential cancer site in humans Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 85 Draft, do not cite or quote exposed to disinfectant by products ( Boorman et al., 1999). Stauber et al. ( 1995), in an abstract,

reported that preliminary data suggest that BCA and DBA induce hepatic tumors in B6C3F1 mice. No experimental details were provided in the brief summary.

F. Summary Monobromoacetic acid The toxicity data for MBA are very limited. The oral LD 50 for MBA was reported as 177 mg/ kg in male rats ( Linder et al., 1994a). MBA is a dermal irritant when topically applied to the skin of rabbits ( Eriksson et al., 1994). The systemic toxicity of MBA has not been well studied by any route of exposure. Reproductive toxicity studies are limited to a single dose or 14 day oral gavage study assessing MBA spermatotoxicity ( Linder et al., 1994a), and have not demonstrated either general toxicity or spermatoxicity. A published abstract ( Randall et al., 1991) reported decreased maternal weight gain, decreased live fetus size, and an increased incidence of softtissue malformations in female rats orally exposed to MBA on GD 6 15; the LOAEL and NOAEL for these effects were 100 and 50 mg/ kg/ day, respectively. However, the full study has not been published and, thus, these data are of limited utility in both hazard characterization and risk assessment. The carcinogenicity of MBA has not been evaluated by any route of exposure. The Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 86 Draft, do not cite or quote genotoxicity data base is limited to four in vitro and one in vivo newt larvae studies the results of which are mixed.

Bromochloroacetic acid The database for BCA toxicity is limited. BCA is predicted to be a severe dermal irritant,

based on QSAR modeling ( Eriksson et al., 1994). Several oral toxicity studies of BCA have identified the kidney and liver as target organs of systemic toxicity, although reported effects have been minimal and/ or equivocal ( NTP, 1998; Austin et al., 1996). Although BCA did not induce male reproductive organ toxicity or affect sperm quality or male fertility in rats in an NTP ( 1998)

reproductive and developmental screening assay, three more recent studies reported the occurrence of reduced sperm quality and decreased male fertility in mice and rats. In a published abstract by Luft et al ( 2000), male mice treated with 72 mg/ kg/ day BCA for 14 days had impaired sperm quality and reduced fertility; no effects were observed at 24 mg/ kg/ day. In two studies reported in an as yet unpublished manuscript ( Klinefelter et al, 2002a), male rats treated with BCA doses ranging from 8 to 216 mg/ kg/ day showed a variety of adverse effects, including significant impairment in sperm motility, abnormal sperm morphology, and altered spermiation. In one of these studies, fertility was assessed by in utero insemination of untreated females with sperm from treated males; significantly reduced fertility was observed at all doses tested ( 8, 24, Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 87 Draft, do not cite or quote and 72 mg/ kg/ day), although there was no dose response. The LOAEL for the Klinefelter et al.

( 2002a) study was 8 mg/ kg/ day and a NOAEL could not be determined.

In the reproductive and developmental screening assay conducted by NTP ( 1998), BCA treatment at 50 mg/ kg/ day for 30 35 days adversely affected the ability of female rats to conceive and carry a full litter to term. Adverse reproductive effects were most prominent early in gestation, as demonstrated by significantly increased pre implantation losses and decreased total implants per litter, and nonsignificant but elevated post implantation losses and increased number of resorptions. The statistical power of the screening assay was seriously limited by the small sample sizes and the low number of pregnancies in each dose group. Nonetheless, based on statistically significant and toxicologically relevant reproductive and developmental end points, the LOAEL and NOAEL for this study were 50 mg/ kg/ day and 19 mg/ kg/ day, respectively. No effects on male reproductive endpoints ( testicular histopathology, epididymal sperm measures,

spermatid head counts, sperm morphology, or sperm motility) were observed in the NTP ( 1998)

screening study. It is unclear why these results differed from those of the Klinefelter et al. ( 2002a)

study. The genotoxicity database for BCA is very limited. Although positive results have been reported in a bacterial mutagenicity assay ( NTP, 2000b), in vivo studies do not provide evidence of a direct genotoxic effect ( Austin et al., 1996; Parrish et al., 1996). BCA ( either parent or metabolite) induces oxidative stress in the livers of orally treated rodents as measured by increased 8 OHdG; these findings suggest that oxidative DNA damage might result from BCA exposure. However, the data are insufficient to comprehensively evaluate the potential Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 88 Draft, do not cite or quote genotoxicity of BCA. Similarly, the carcinogenic potential of BCA is not known. In a published abstract, Stauber et al. ( 1995) reported that BCA induces liver tumors in mice, but there are no published reports of a full bioassay. A 2 year NTP toxicity and carcinogenesis study with BCA is scheduled to be conducted in the near future ( NTP, 2000b).

Dibromoacetic acid The toxicity database for DBA is limited and has been developed largely to explore its effects on the male reproductive tract. The oral LD 50 was reported to be 1737 mg/ kg in male rats ( Linder et al., 1994a). The liver has been reported to be a systemic target organ of DBA induced toxicity, although only minimally adverse effects have been observed in short term studies ( Parrish et al., 1996; NTP, 1999). Phillips et al. ( 2002, published abstract) evaluated the neurobehavioral toxicity and neuropathology of DBA administered in drinking water to male and female Sprague Dawley rats for 6 months. Neurotoxic effects included mild gait abnormalities,

hyptonia, decreased forelimb and hindlimb grip strength, decreased sensorimotor responsiveness ( as measured by responses to a tail pinch and click), and decreased motor activity.

Neuropathologic examination showed significant myelin fragmentation, axonal swelling, and axonal degeneration in the white matter of the spinal cord, and eosinophilic or faintly basophilic,

occasionally vacuolated swelling, indicative of degenerating axons, in the spinal cord gray matter.

The LOAEL for neurobehavioral effects was 20 mg/ kg/ day, the lowest dose tested. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 89 Draft, do not cite or quote Linder and his colleagues have studied the effects of DBA on spermatogenesis and the resulting consequences for male fertility, using a number of different experimental protocols,

including a single high dose study ( Linder et al., 1994a), a 14 day study ( Linder et al., 1994b),

and a longer term study ( Linder et al., 1995; Linder et al. 1997a). In all of these studies, DBA was clearly spermatotoxic in rats. Based on histopathologic changes in spermiation, the equivocal LOAEL for the 14 day study was the lowest dose tested, 10 mg/ kg/ day ( Linder et al., 1994b). In longer term studies in which male Sprague Dawley rats were exposed to DBA for up to 79 days,

the equivocal LOAEL for histopathologic changes in spermiation was 10 mg/ kg/ day and the corresponding NOAEL was 2 mg/ kg/ day. The severity of the DBA induced male reproductivetract toxicity was both dose and duration dependent. Extensive reproductive tract histopathology was only partially reversed in rats administered 250 mg/ kg/ day by oral gavage for 42 days followed by a 6 month recovery period, indicating that structural damage to the reproductive organs was permanent under the conditions of this dosing regime. In an abstract,

Veeramachaneni et al. ( 2000) reported that rabbits exposed to DBA in utero from GD 15 to parturition, during lactation, and during the post weaning period through 24 weeks of age,

exhibited reduced sperm fertility; the lowest dose tested, 0.97 mg/ kg/ day, was the LOAEL. In contrast to these findings, a reproductive toxicity study by Vetter et al. ( 1998) did not observe significant spermatotoxic effects in male Crl: CD( SD) BR rats treated with single oral gavage doses of 600 or 1200 mg/ kg DBA. However, mild testes histopathology was observed in both dose groups; the LOAEL for acute reproductive effects was 600 mg/ kg and a NOAEL could not be Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 90 Draft, do not cite or quote determined. The reasons for the differences in DBA induced spermatotoxicity between the Vetter et al. ( 1998) study and those of Linder and his colleagues are unclear. In the recent twogeneration reproductive toxicity study ( Chlorine Chemistry Council, 2001; Christian et al., 2002),

impaired spermatogenesis was observed in male rats of the P and F1 generations at DBA drinking water concentrations of 250 ppm and above ( equivalent to a LOAEL of 22 mg/ kg/ day for the P generation, and at least 22 mg/ kg/ day for the F1 generation); abnormal pathology of the testes and epididymes was noted in some males of the F1 generation at 650 ppm ( equivalent to a LOAEL of not less than 75 mg/ kg/ day). However, in contrast with the shorter term study that showed adverse mating performance effects at 250 mg/ kg/ day and higher ( Linder et al., 1995), no adverse treatment related effects on mating performance, gestation length, fertility, pup mortality and viability, and other functional indices of successful reproductive behavior were observed at DBA drinking water concentrations up to 650 ppm ( 52 to 132 mg/ kg day). Alternatively, these studies in combination may define a NOAEL/ LOAEL boundary for functional effects of DBA on reproduction. The weight of evidence indicates that DBA is a potent male reproductive system toxicant and exerts its primary effects by interfering with the normal processes of spermatogenesis; however, the data are mixed as to whether these effects interfere with normal reproductive function.

In a study on DBA effects on female reproductive capacity ( Cummings and Hedge, 1998),

reproductive outcomes were not adversely affected in rats administered oral gavage doses of up to 250 mg/ kg/ day DBA on GD 1 8, although DBA induced a significant increase in serum 17   Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 91 Draft, do not cite or quote estradiol at the highest dose tested. The reproductive toxicity NOAEL for this study was 250 mg/ kg/ day and a LOAEL could not be determined. In three published abstracts summarizing the findings of developmental toxicity studies, DBA was reported to: adversely affect pre and postnatal mortality, decrease pup weight, and induce skeletal ( tail) and soft tissue ( kidney)

malformations in mice exposed in utero to DBA ( Narotsky et al., 1996; Narotsky et al., 1997);

and delay the pubertal development and reduce the sperm fertility of male rats exposed in utero,

during lactation, and during the post weaning period ( to PND 98) to DBA in drinking water ( Klinefelter et al., 2000). Delayed parturition was also observed at 24 mg/ kg/ day in one of the mouse studies ( Narotsky et al., 1996) but the biological significance of this finding is unclear. The LOAEL and NOAEL for soft tissue kidney defects ( hydronephrosis) in the mouse study ( Narotsky et al., 1997) were 100 and 50 mg/ kg/ day, respectively. In the pubertal development and sperm fertility study ( Klinefelter et al., 2000), reduced sperm fertility was observed in all dosed male offspring; the LOAEL was 50 mg/ kg/ day and a NOAEL could not be determined. The results described in these abstracts, however, cannot be fully evaluated until a complete report of findings is published. In the two generation drinking water study ( Chlorine Chemistry Council,

2001; Christian et al., 2002), no effects were observed on female reproductive function.

Treatment related developmental effects included a statistically significant delay in preputial separation and vaginal patency, and/ or a reduction in anogenital distance on LD 22, observed in F1 or F2 pups exposed during gestation and lactation to 650 ppm DBA. These findings were Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD V 92 Draft, do not cite or quote attributed to the significant growth retardation observed in these animals, which was secondary to decreased water consumption ( due to taste aversion) by both pups and their mothers.

The immunotoxicity of DBA administered in drinking water has been evaluated in four studies in mice ( NTP, 1999). A number of different end points were assessed, including thymus and spleen weights, number and type of spleen cells, macrophage activation, natural killer ( NK)

cell activity, and specific and general IgM antibody forming responses. The most sensitive and reliable measure was a decrease in spleen IgM antibody forming cell responses, representing a clear decrease in immune system function, accompanied by an increase in the number of spleen macrophages. The LOAEL and NOAEL for these endpoints were approximately 70 and 38 mg/ kg/ day, respectively. No data were identified for the toxicity of DBA following exposure by the dermal or inhalation routes.

The weight of evidence for DBA mutagenicity/ genotoxicity indicates that DBA is mutagenic and genotoxic, although the nature of the DNA damage induced by DBA remains unclear. The potential for DBA carcinogenicity is not known. In published abstracts, So and Bull ( 1995) reported that DBA induces aberrant crypt foci in the colon of rats, and Stauber et al.

( 1995) reported that DBA induces liver tumors in mice. However, no complete reports of DBA cancer bioassays have been published. A 2 year NTP toxicity and carcinogenicity study with DBA is scheduled to be conducted in the near future ( NTP, 2000a). Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VI 1 Draft, do not cite or quote Chapter VI. Health Effects in Humans No studies were identified that directly evaluated human health effects of exposure to MBA, BCA, or DBA via any route. Rather, most of the human health data for brominated acetic acids are as components of complex mixtures of water disinfection byproducts. These complex mixtures of disinfection byproducts have been associated with increased potential for bladder,

rectal, and colon cancer ( reviewed by Boorman et al., 1999) and adverse effects on reproduction ( reviewed by Nieuwenhuijsen et al., 1999).

Most studies of human health effects following exposure to water disinfectant byproducts have used total trihalomethanes as the exposure metric, and the risks attributable to brominated acetic acids typically have not been reported. In one study by Klotz and Pyrch ( 1999), a population based case control study was conducted on the relationship between drinking water exposure to trihalomethanes, haloacetonitriles, and haloacetic acids and neural tube defects. The study included 112 cases of neural tube defects in 1993 and 1994 in New Jersey. A total of 248 controls were selected randomly from all New Jersey births. No significant relationship between total trihalomethanes and neural tube defects was observed for analysis of all cases, cases restricted to subjects with known residency at conception, or cases restricted to isolated cases of neural tube defects. However, a statistically significant difference between cases and controls was observed when cases were restricted to subjects with known residency at conception and to cases with isolated neural tube defects. Based on this more stringent case definition, a prevalence odds Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VI 2 Draft, do not cite or quote ratio ( POR) of 2.1 was reported ( 95% confidence interval, 1.1  4.0) for the highest tertile of 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 haloaceticacid tertiles. The specific haloacetic acids that were measured as part of the total haloacetic acidexposure estimate were not specified. Based on the results of the study, the authors concluded that the haloacetic acids did not exhibit a clear association with neural tube defects. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 1 Draft, do not cite or quote VII. Mechanisms of Toxicity A. Mechanisms of Noncancer Toxicity Little is known about the molecular mechanisms of toxicity of the brominated acetic acids.

It has not been determined conclusively whether the parent compound or a metabolite is the toxic moiety; and there are clear differences in the potency and spectrum of effects induced by MBA,

BCA, and DBA. For example, MBA is more acutely toxic than DBA, but, unlike DBA, is not spermatotoxic ( Linder et al., 1994a). The available data on the mechanisms of toxicity of the brominated acetic acids for selected endpoints are described here.

One proposed cellular basis for the toxicity of MBA is through direct alkylation of sulfhydryl and amino groups via its ability to inhibit a number of mammalian enzymes in in vitro studies. However, the data are only suggestive, due to the use of high MBA concentrations and purified proteins, and in vitro test systems. Incubation of purified human thioredoxin reductase with MBA at pH 6.5 inhibited enzyme activity by > 99% ( Gorlatov et al., 1998). Although the concentration of MBA used in the reaction was not presented, the MBA concentration can be estimated to be 0.7 mM, based on reaction volumes and moles of compound used. Similar incubations at pH 6.5 and pH 8 led to carboxymethylation of specific selonocysteine residues of thioredoxin reductase ( Gorlatov et al., 1998), suggesting that MBA can inhibit enzyme activity through alkylation of sulfhydryl groups. MBA reactions with amino groups in proteins have also Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 2 Draft, do not cite or quote been suggested as a mechanism for enzyme inhibition under certain physiological conditions. For example, Ito et al. ( 1994) reported that 100 mM MBA nearly completely inhibited purified human urinary DNase I activity and resulted in modifications of critical histidine residues. Shapiro et al.

( 1988) reported that 30 mM MBA resulted in carboxymethylation of several histidine sites in purified human angiogenin and inactivated the protein. Whitney ( 1970) reported the inhibition of human carbonic anhydrase B following incubation of human erythrocytes with 5 mM MBA. While these data show that MBA can alkylate cellular proteins and disrupt their normal function in vitro,

the enzyme inhibition studies were carried out primarily for the purpose of identifying critical amino acid residues for normal protein function, not for determining the mechanism of action of MBA toxicity. The threshold concentration for MBA induced enzyme inhibition was not reported,

and the relevance of these findings to in vivo toxicity is not clear. In vitro studies lack a number of biological characteristics that can modulate toxicologic responses in the intact organism, including hepatic metabolism, toxicokinetics, and the presence of additional protein systems. Further, the concentrations of MBA used in these studies are similar to or higher than the high doses used in animal studies, and may not be directly comparable to low dose environmental exposures. In addition, critical high affinity enzyme targets that lead to the observed toxic effects of MBA have not been identified, although the alkylation of DNA in cells in tissue culture was reported to induce DNA damage ( Stratton et al., 1981). Taken together, these data demonstrate the potential for MBA to adversely impact cellular macromolecules, but whether this ability is responsible for MBA toxicity has not been clearly shown. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 3 Draft, do not cite or quote Several in vivo animal studies have demonstrated that MBA, BCA, and DBA are developmental toxicants ( Randall et al., 1991 ( abstract); NTP, 1998; Narotsky et al., 1996, 1997 ( abstracts); Klinefelter, 2000c ( abstract)), although the spectrum of adverse developmental effects, the associated toxic potencies, and the critical periods for gestational exposure differ significantly among these three compounds. These developmental toxicity studies are described in detail in Chapter V.

The results of several mouse whole embryo testing studies provide mechanistic support for the potential for developmental toxicity of the brominated acetic acids in vivo and suggest possible mechanisms of embryotoxicity. However, in vitro studies such as whole embryo culture ( WEC) have limited utility for predicting either the spectrum of adverse developmental effects or the associated toxic potencies in intact organisms. In addition to maternal influences in the whole animal during gestation and lactation, potentially adverse developmental responses observed in vitro can be modified by hepatic metabolism, toxicokinetics, the activity of additional protein systems, and other physiologic and biochemical processes. Further, the chemical concentrations required to induce developmental effects in in vitro experimental systems such as WEC are usually much higher than low dose environmental exposures. Thus, these in vitro data are hypothesis generating only, and must be supplemented by mechanistic data from studies conducted in vivo. To date, the data from in vivo and in vitro developmental toxicity studies are limited and do not provide significant information on possible or likely mechanisms of Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 4 Draft, do not cite or quote developmental toxicity for brominated acetic acids, particularly mechanisms which might explain observed differences in in vivo toxicity among brominated acetic acid compounds.

In vitro studies with mouse whole embryo culture ( WEC) have demonstrated that MBA,

DBA, and BCA have the potential to induce developmental toxicity, including skeletal ( e. g.,

neural tube defects, pharyngeal arch defects) and soft tissue ( e. g., cardiac and eye defects)

malformations ( Hunter et al., 1996, 1999 ( abstract)). Ward et al. ( 1997, 1998) studied the effects of BCA and DBA on protein kinase C ( PKC) activity in mouse WEC as a possible mechanism of developmental toxicity ( PKC is a signal transduction enzyme that controls the activity of a variety of proteins involved in cell growth and differentiation via phosphorylation).

Both BCA and DBA, in the concentration range of 0.3  3 mM, inhibited purified rat brain PKC in a dose dependent manner. These compounds also inhibited PKC activity in homogenates of GD 9 embryos. A follow up study was conducted to evaluate the relationship between BCA and DBA's ability to inhibit PKC and their observed embryotoxicity ( Ward et al., 2000). Groups of 6 12 whole CD 1 mouse embryos ( early somite stage conceptuses) were cultured for up to 24 hours with 300 µ M DBA, 300 µ M BCA, Bis I ( a specific PKC inhibitor with previously defined embryotoxic effects ( Ward et al., 1998), staurosporine ( a potent, but non specific PKC inhibitor known to interact with the cell cycle) or Bis V ( negative control). These concentrations of BCA and DBA were chosen to induce embryotoxicity in nearly all embryos as evidenced by morphological abnormalities, primarily neural tube defects ( data shown for DBA only), but not embryolethality. Neither BCA nor DBA disrupted the cell cycle. However, flow cytometry Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 5 Draft, do not cite or quote revealed the accumulation of sub G1 events ( indicative of apoptosis) with BCA and staurosporine, but not DBA, Bis I or Bis V. For BCA, sub G1 events were particularly pronounced in the head region but not in the heart. Although sub G1 events in the head region were also increased by DBA treatment ( 2 to 3 fold increase), this increase was not statistically significant. Thus, BCA and staurosporine, but not Bis I or DBA, induced apoptosis. These mixed results for the specific PKC inhibitor Bis I and nonspecific PKC inhibitor staurosporine make it unclear whether the ability of BCA to inhibit PKC is related to the induced apoptotic response.

However, because the two inhibitors have differing PKC isoform specificities, a direct role of PKC inhibition cannot be ruled out. The study authors suggested that other possible mechanisms of dysmorphogenesis may include kinase mediated disruption of signal transduction pathways in the neurulation stage embryo.

Hunter et al. ( 1999), in a published abstract, evaluated the ability of known haloacetic acid metabolites to induce dysmorphogenesis in the mouse WEC system. The potency of glycolate,

glyoxylate, and oxalate were tested. Glycolate induced a low incidence of neural tube defects ( NTDs) at 1000 µ M, while no effects were induced at this concentration for glyoxylate or oxalate. For all three compounds, the severity of effects increased with increasing concentration.

The concentrations of MBA, DBA, and BCA at which dysmorphogenesis was observed in the same test system were much lower than those for identified metabolites. This result suggests that the developmental toxicity of the brominated acetic acids is not due to the metabolites glycolate, Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 6 Draft, do not cite or quote glyoxylate, or oxalate. However, other as yet unidentified intermediate metabolites may be implicated in brominated acetic acid toxicity.

Andrews et al. ( 1999a), in a published abstract, extended the use of whole embryo culture studies by evaluating the potential for developmental effects of BCA and DBA in rat embryo cultures, as compared with mouse embryo cultures reported previously by other investigators ( Hunter et al., 1996; Ward et al., 1996, 1997). Results for DBA were comparable with those from the mouse WEC studies and the toxic potencies of DBA and BCA were similar. In a follow up study, Andrews et. al. ( 1999b, abstract) reported on the potential for embryotoxicity of mixtures of DCA, DBA, and BCA in rat WEC. The experimental results for the mixtures were adequately predicted ( data were not shown) by dose additivity, as proposed by a quantitative structure activity relationship ( QSAR) model developed by Richard and Hunter ( 1996; described below). The effects on dysmorphogenesis were similar to those observed in single compound WEC ( Hunter et al. 1996).

The potential for developmental toxicity among haloacetic compounds, including the mono, di, and tri substituted fluoro, chloro, and bromoacetic acids was also studied using WEC by Hunter et al. ( 1996), who were mainly interested in determining if structure activity relationships could be established relating the type and degree of halogen substitution to the severity and spectrum of possible developmental effects. For both MBA and DBA, malformations were increased at sublethal doses and the spectrum of effects was similar, but MBA was significantly more potent than DBA. Neither MBA nor DBA reduced the pH of the culture Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 7 Draft, do not cite or quote medium, precluding this as the mechanism responsible for the observed developmental toxicity.

Overall, the effects of most haloacetic acids were qualitatively similar and the ranking of toxic potency was monobromo > monochloro > dibromo > trichloro, and tribromo > acetate > dichloro > trifluoro > difluoro. Using these data, Richard and Hunter ( 1996) developed a QSAR model in order to test predictions regarding the toxic potency of haloacetic acids, and offer insight into the mechanism( s) of the developmental toxicity of this class of compounds. The potencies predicted by this model were compared with the potencies determined experimentally. Experimentally, the potencies of the monohaloacetic acids increased with halogen size ( iodo > bromo > chloro >

fluoro), and the model was able to correctly predict this trend, although slightly overestimating chloroacetic acid potency, and slightly underestimating those of fluoroacetic acid and bromoacetic acid. Experimentally, and as predicted by the model, the same trend held for the three dihaloacetic acids ( dibromo > dichloro > difluoro), although the model predicted more similar potencies of the difluoro and the dichloro compounds than were seen experimentally. However, the model was unable to accurately predict the toxic potency of trihaloacetic acids, overestimating tribromo and underestimating trichloro and trifluoro acetic acid potencies. Richard and Hunter ( 1996) used this model to predict the developmental toxicity potencies of several untested haloacetic acids,

including BCA. The predicted potency of BCA was similar to that of DBA, and the relative potencies of the brominated acetic acids were MBA> BCA

DBA. The results of the study demonstrated an increase in teratogenic potency with increasing pKa values. Since pKa increased with decreased degree of halogenation, this relationship accounted for increasing potency with Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 8 Draft, do not cite or quote decreasing degree of halogenation. The authors hypothesized that the fit of the data supported a common mechanism of action for haloacetic acids, with differing potencies engendered by the type and degree of halogen substitution. However, insufficient data are available to confirm this hypothesis.

The most well studied noncancer endpoint of concern for brominated acetic acids is male reproductive toxicity. Some evidence also suggests that liver, kidney, and immune system toxicity can occur. With the exception of effects of DBA on spermatogenesis, the data are limited and only tentative conclusions regarding mechanisms of toxicity can be made. Unifying ideas on mechanisms of toxicity across the class of brominated acetic acids will be discussed below for liver, kidney, and reproductive effects, respectively.

Effects of oral dosing on the liver have been reported for BCA ( Parrish et al., 1996; NTP,

1998) and DBA ( Parrish et al., 1996; NTP, 1999; Chlorine Chemistry Council, 2001; Christian et al., 2002). However, in most cases, minimal evaluations were conducted, and observed effects were limited to increased liver weight and marginal histopathological changes including cytoplasmic vacuolization. Both increased liver weight and cytoplasmic vacuolization are consistent with liver glycogen accumulation ( NTP, 1998), a phenomenon that occurs with dichloroacetic acid ( DCA) ( Kato Weinstein et al., 1998)

The induction of lipid peroxidation and oxidative DNA damage in the livers of mice treated with BCA or DBA, in the absence of peroxisome proliferation ( Austin et al., 1996; Parrish et al., 1996), is consistent with the potential for liver toxicity of these compounds. The Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 9 Draft, do not cite or quote metabolism of BCA and DBA has not been sufficiently characterized to clearly identify the intermediates involved in lipid peroxidation. Both compounds are apparently metabolized in a manner similar to DCA, a weak peroxisome proliferator ( De Angelo et al, 1989). Similar reactive intermediates derived from BCA and DBA might be responsible for their ability to induce lipid peroxidation.

Recent studies have demonstrated the metabolism of both BCA and DBA, as well as that of DCA, is mediated by GST Zeta ( Tong et al., 1998a. Cornett et al. ( 1999) proposed that DCAinduced toxicity results from the inhibition of GST Zeta, which is also known as maleylacetoacetate isomerase, an enzyme involved in tyrosine catabolism. Cornett et al. ( 1999)

found that DCA exposure increased the urinary excretion of maleylacetone, a reactive metabolite of tyrosine catabolism. Based on these data, the authors suggested that increases in reactive tyrosine metabolites might contribute to adverse effects induced by DCA. BCA and DBA also inhibit GST Zeta activity ( Anderson et al., 1999), suggesting that perturbation of tyrosine catabolism might also be involved in the toxicity induced by brominated acetic acids. The formation of reactive intermediates or oxidative stress responses ( such as lipid peroxidation)

either by BCA or DBA directly or through tyrosine metabolites may be particularly important in the potential for tumorigenicity of these compounds. Stauber et al. ( 1995) has reported in an abstract that BCA and DBA induced liver tumors in mice; however, a complete presentation of this study has not been published and, thus, these findings cannot be evaluated. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 10 Draft, do not cite or quote Another potential target for brominated acetic acids is the kidney. BCA treatment increased renal tubular dilatation/ degeneration in female rats, but these changes were not statistically significant ( NTP, 1998). In males in the same study, no treatment related changes in kidney weight or labeling index were detected and histopathology was not reported. NTP ( 1999)

reported increased kidney weight following oral dosing with DBA in female mice. One potential mechanism hypothesized for the observed renal effects might be metabolism of brominated acetic acids to oxalic acid, a demonstrated kidney toxicant that causes tubule damage by forming oxalate crystals ( Kennedy et al., 1993; Webster et al., 2000), but this hypothesis has not been directly tested. In addition, the nature and extent of kidney toxicity induced by BCA is not well characterized in the currently available data. Further, in pharmacokinetic studies with DCA, the chlorinated analog of BCA, only 2 4% of the administered dose was recovered as urinary oxalate following high dose gavage exposures ( James et al., 1998); thus, it is questionable whether sufficient oxalate would be formed during the metabolism of brominated acetic acids to induce kidney toxicity by this mechanism.

The major area of emphasis for toxicity studies for the brominated acetic acids, has been on potential reproductive effects. Convincing evidence of adverse spermatogenic effects and decreased male fertility is available for DBA ( Linder et al., 1994a; Linder et al., 1994b; Linder et al., 1995; Linder et. al., 1997a; Klinefelter et al., 2000; Veeramachaneni et al., 2000; ; Chlorine Chemistry Council, 2001; Christian et al., 2002) and BCA ( Klinefelter et al., 2002a). In the single Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 11 Draft, do not cite or quote study identified, MBA had no effect on spermatogenesis, under treatment conditions similar to those that yielded positive indications of spermatotoxicity for DBA ( Linder at al., 1994a).

Several potential mechanisms of male reproductive toxicity have been explored. Linder et al. ( 1994a) identified potential targets for the effects of DBA, based on the spectrum and timing of effects on different stages of sperm development 2, 14, and 28 days after administration of a single, high dose of DBA. The study authors suggested that DBA induced a sequence of two developmental changes in epididymal sperm: abnormal head development in Step 10 or earlier spermatids, and abnormal flagellar development as the spermatids passed through the cauda. This early paper also noted that the retention of Step 19 spermatids ( the most common effect of DBA on sperm development occurring at the lowest doses) is an effect observed following treatments that alter hormone status. The observation that DBA reduced circulating testosterone levels is consistent with the observed effects on Step 19 spermatids ( Linder et al., 1994a). Another potential target for DBA might be Sertoli cells, because the presence of testicular debris suggested to the study authors that disruption of the endocytic activity of these cells might be occurring. The hypothesis that Sertoli cells are a target for DBA induced spermatotoxicity was supported by the results of a later multiple dosing study by this same group of investigators ( Linder et al., 1997a). In this study, the authors reported that the spectrum of late spermatid dysmorphogenesis and the formation of atypical residual bodies was consistent with the disruption of normal Sertoli cell function. The appearance of large intraepithelial vacuoles in the Sertoli cell cytoplasm was cited as additional evidence for damage to these cells. The authors further noted Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 12 Draft, do not cite or quote that the disruption of normal Sertoli cell function could be explained by DBA induced damage to the cell cytoskeleton, many proteins of which play a direct role in the developmental functions of Sertoli cells ( Linder et al., 1997a). In light of the observed dose severity response, with low DBA doses causing retention of Step 19 spermatids and higher DBA doses causing overt changes in sperm morphology, it is possible that there are multiple targets for DBA induced toxicity in the male reproductive tract and that toxicity might be induced by more than one mechanism.

Although the cellular mechanisms of brominated acetic acid spermatotoxicity have not been clearly identified, a hypothesis consistent with the existing data is that the disruption of normal Sertoli cell function is due to alkylation of critical cellular proteins. There are multiple pathways by which DBA might alter the function of key Sertoli cell proteins. For example, DBA itself, reactive DBA metabolites, metabolites formed by the disruption of tyrosine catabolism,

and/ or reactive oxygen species induced secondary to DBA treatment might be involved in amino or sulfhydryl group modifications of critical cellular proteins. Alternatively, these same moieties might be directly cytotoxic, as suggested by the observed Sertoli cell morphology changes following DBA treatment ( Linder et al., 1997a).

More recent work has tentatively identified one of the key protein targets of brominated acetic acid spermatotoxicity. As part of a study of BCA spermatotoxicity and fertility assessment,

Klinefelter et al ( 2002a) analyzed 120 sperm proteins extracted from male Sprague Dawley rats treated with DBA for 14 days. A significant reduction in two of these proteins, SP22 and SP9,

was observed following treatment, and the shape of the dose response curve for SP22 mirrored Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 13 Draft, do not cite or quote that of reduced male fertility. The study authors concluded that BCA, like DBA, is capable of perturbing spermatogenesis and fertility, and that SP22 appears to be useful as a sperm biomarker of fertility. Additional studies have suggested that SP22 represents a protein found within the cytoplasm of round spermatids that migrates to the plasma membrane overlying the equatorial segment of the sperm head later on in the spermatogenic process ( Jeffrey, 1999), and appears to play an important role in the fertilization process ( Klinefelter et al., 2002b). Additionally, the nuclear form of SP22 may be a positive regulator of the androgen receptor ( Takahashi et al.,

2001), and it has been postulated that haloacetic acids acting on SP22 and other sperm proteins may indirectly compromise androgen dependent maintenance of spermatogenesis ( Klinefelter et al., 2002b).

The results of several studies have suggested that DBA may act as a reproductive toxicant by interfering with steroidogenesis. In the Linder et al. ( 1994a) study, alterations in sperm parameters at high doses were accompanied by a sharp attenuation of serum testosterone levels,

suggesting a steroidogenic effect. During steroidogenesis, cholesterol is converted by steroidogenic acute regulatory protein ( StAR) and P450 side chain cleavage enzyme ( P450 scc) to pregnenolone, a precursor of progesterone. 3  hydroxysteroid dehydrogenase ( 3  HSD)

catalyzes the conversion of pregnenolone to progesterone, which is then converted through a series of catalyzed steps to testosterone. Balchak et al. ( 2000) showed that in vitro exposure of preovulatory follicles, obtained from immature Sprague Dawley female rats primed with a subcutaneous injection of pregnant mare serum gonadotrophin, to 50 µ g/ mL DBA over a 24 hour Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 14 Draft, do not cite or quote period blocked release of progesterone stimulated by human chorionic gonadotrophin ( hCG)

without having a comparable effect on estradiol. Using a similar protocol, Goldman and Murr ( 2002) conducted a series of experiments to establish a dose response for the effects of DBA on progesterone secretion and to identify the site( s) of action along the initial segment of the steroidogenic pathway. Progesterone release was significantly depressed following 24 hour incubation with 50 µ g/ mL DBA, but not with 2 or 10 µ g/ mL, under both baseline and hCGstimulated conditions. The suppression in progesterone release at 50 µ g/ mL was shown ( by analysis of the incubated follicles) to be due to a DBA induced reduction in follicular progesterone content. No effects of DBA treatment on estradiol secretion were observed. In a second experiment, follicular cultures were supplemented with pregnenolone to evaluate any effects of DBA ( 50 µ g/ mL) on 3  HSD catalyzed conversion to progesterone. Under these conditions, progesterone release was increased up to 13 fold compared to that of untreated controls. Thus, DBA did not attenuate progesterone release in the presence of pregnenolone. The increase in progesterone secretions observed during the 24 hour incubation period indicated to the authors that DBA induced depression of progesterone release observed in the earlier experiment was not likely to be due to effects on follicular viability. Determination of follicular progesterone levels showed that the progesterone content was significantly elevated about 3 3.5

fold under both the hCG stimulated and non stimulated conditions, even though pregnenolone supplementation suppressed the DBA attenuation of progesterone release. These findings imply an increase in progesterone synthesis, both in the presence and absence of hCG supplementation, Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 15 Draft, do not cite or quote that was not reflected in the progesterone release data. Pregnenolone treatment did not have an effect on estradiol secretion, something not reflected in the release data. In a third experiment,

follicular cultures were supplemented with 22 R hydroxycholesterol ( 22R HC). According to the study authors, 22R HC can serve as a membrane permeable precursor for pregnenolone synthesis,

circumventing transport within the mitochondrial membrane by the StAR protein; its presence in DBA treated follicular cultures allows for the assessment of DBA effects on the activity of P450 cholesterol side chain cleavage enzyme ( Pscc). Supplementation with 22R HC eliminated the DBA induced attenuation effect on baseline progesterone release, although the attenuation in the hCG stimulated secretion was still present. The study authors concluded that exposure to DBA may have an effect on the StAR mediated transport of cholesterol within the mitochondrial membrane and an effect on receptor or postreceptor events triggered by hCG, but only at high doses.

B. Cancer Mechanisms As described in Section V. D, there are a number of studies on the genotoxicity of MBA,

BCA, and DBA. The data are inadequate for determining whether MBA or BCA are genotoxic,

but suggest that DBA is genotoxic. Some comparisons to the chlorinated acetic acids may also be useful. However, in the absence of data from completed cancer bioassays, consideration of the mechanism of carcinogenicity for MBA, BCA, or DBA is premature and overly speculative. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 16 Draft, do not cite or quote C. Sensitive Subpopulations No data are available to determine whether sensitive subpopulations exist with regard to differences in age or genetic susceptibility The developmental toxicity of MBA, BCA, and DBA has been evaluated, to a limited degree, as described below. No multi generation reproductive study has been conducted for MBA or BCA, although a recent two generation study of DBA did not find evidence that the developing fetus is more sensitive than adults to the effects of DBA ( Chlorine Chemistry Council, 2001; Christian et al., 2002). No information on age dependent changes in the expression of genes encoding enzymes thought to metabolize bromoacetic acids was identified. Further, it is not yet clear whether the parent compound or one ( or more)

intermediate metabolites is the toxic moiety of concern. In the absence of these data, a full determination of fetal or early age susceptibility and differences in sensitivity associated with genetic variability among individuals cannot be made.

For MBA, data relevant to potential fetal sensitivity is limited to a single developmental study reported in a published abstract ( Randall et al., 1991). The induction of fetal effects only at doses that also affected maternal weight does not suggest that the fetus is more sensitive.

However, because the study was only presented in an abstract, and the available reports do not cover the full range of developmental endpoints, no firm conclusions can be drawn from the limited database. As for MBA, the data for the potential developmental toxicity of BCA are limited to a single reproductive and developmental toxicity screening assay ( NTP, 1998). The Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 17 Draft, do not cite or quote NOAEL for decreased live fetuses/ litter and decreased total implants/ litter was 19 mg/ kg/ day,

while this dose was also considered a NOAEL for general toxicity in adult males and females.

Thus, for both MBA and BCA, the data are limited, but the available data do not support the hypothesis that fetuses or children are more sensitive than adults.

The data for DBA are also insufficient for a full evaluation of fetal and childhood susceptibility. In two published abstracts ( Narotsky et al., 1996, 1997), DBA was reported to induce developmental toxicity in CD 1 mice at doses lower than those which produced maternal toxicity. However, these data are preliminary and have only been reported in abstract form. Thus,

sufficient detail on the results is not available and age related differences in sensitivity cannot be clearly determined. In addition, the NOAEL for both maternal and fetal toxicity reported in these abstracts was well above the NOAEL for effects on spermatogenesis.

Although DBA is clearly spermatotoxic in mature animals ( Linder et al., 1994a; Linder et al., 1994b; Linder et al., 1995; Linder et. al., 1997a; Chlorine Chemistry Council, 2001; Christian et al., 2002), the data are conflicting as to whether the developing male reproductive tract is particularly sensitive. In a published abstract, Klinefelter et al. ( 2000) reported developmental delays ( delayed preputial separation) in male rats exposed to DBA in utero from GD15 through PND 98 at doses of 50 mg/ kg/ day and higher. However, due to significant individual variability in this developmental measure, the statistical and biological significance of this finding is unclear,

and insufficient detail is available from the published abstract to evaluate these results. Klinefelter and his colleagues are currently conducting a follow up in utero exposure study using lower doses Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 18 Draft, do not cite or quote ( personal communication). However, in the two generation reproductive/ developmental toxicity study ( Chlorine Chemistry Council, 2001; Christian et al., 2002), delayed preputial separation in males and delayed vaginal patency in females in the F1 generation was attributed to a general retardation of growth associated with decreased water intake and food consumption and not to a direct treatment effect. Veeramachaneni et al. ( 2000) reported in an abstract that exposure of rabbits in utero from gestation day 15 to 24 weeks of age reduced the fertility of sperm from treated males. The lowest dose tested, 0.97 mg/ kg/ day was considered to be the LOAEL. This LOAEL for fertility changes was 10 fold lower than the LOAEL of 10 mg/ kg/ day reported in Linder et al. ( 1997a) for altered histopathology. However, this difference could reflect interspecies differences ( rabbits versus rats), or differences in the duration of dosing ( 24 weeks versus 79 days), as well as increased sensitivity of the developing male reproductive tract.

Further, sufficient detail on this study is not available from the abstract to adequately assess these findings. A two generation study of DBA administered in drinking water to rats ( Chlorine Chemistry Council, 2001; Christian et al., 2002) found no evidence that rats exposed to DBA in utero and for the first 71 days of life are more susceptible than adults to the effects of DBA on spermatogenesis. Therefore, the data are not sufficient to determine the relative sensitivity to DBA of the developing versus mature male reproductive tract. Similarly, the data on the male reproductive effects of BCA are insufficient to determine the relative sensitivity to DBA of the developing versus mature male reproductive tract. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 19 Draft, do not cite or quote In contrast to the lack of information on age dependent differences in the activity of enzymes involved in bromoacetic acid metabolism, several genetic differences in these enzymes have been identified that may engender differences in susceptibility to bromoacetic acids.

GSTZeta has been shown to convert BCA and DBA to glyoxylate. Blackburn et al. ( 2000) reported on human polymorphism of GST Zeta, in which three polymorphic forms of the gene, GST* A,

GST* B, and GST* C, were identified. Based on in vitro experiments with purified proteins encoded by these three forms of GST Zeta, GST* A had a 3.6 fold higher activity toward DCA than the other two human forms. The functional consequences of the polymorphism of GST Zeta cannot be verified in the absence of studies of DBA or BCA metabolism in humans polymorphic for GST Zeta. Further, it is unclear what consequences the observed polymorphism would have on DBA or BCA induced toxicity in humans because it is not known whether the parent compound or one ( or more) of its metabolites is the toxic moiety. A similar analysis for MBA metabolism is not possible because the enzymes involved in MBA metabolism have not yet been identified.

As noted above, DBA and BCA induce liver effects consistent with glycogen accumulation. DCA, the chlorinated analog of DBA, has been shown to increase hepatic glycogen accumulation ( Kato Weinstein et al., 1998) and these authors have suggested that prolonged glycogen accumulation can become irreversible, resulting in liver injury. The enzymatic basis for increased hepatic glycogen accumulation remains unclear. However, it is possible that individuals with glycogen storage disease ( an inherited deficiency or alteration in any one of the enzymes Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 20 Draft, do not cite or quote involved in glycogen degradation) represent another group that may be more susceptible to DBA or BCA toxicity.

Genetic deficiencies in glyoxylate metabolism enzymes, including alanine: glyoxylate aminotransaminase ( AGT) and D glycerate dehydrogenase have been shown to be responsible for primary hyperoxaluria type I and type II, respectively. These disorders result in systemic oxalate overload and induce subsequent kidney toxicity ( Webster et al., 2000). Since bromoacetic acid metabolism could contribute to the total oxalate load, these individuals might have increased susceptibility for kidney toxicity.

No quantitative evaluation has been conducted on the health impact of environmental exposures for individuals harboring polymorphisms in genes related to glycogen storage,

antioxidant response, or oxalate synthesis. In each of these cases, a significant background load of the stressor may be present; thus, the excess risk associated with low doses of brominated acetic acids is not clear, and the data are insufficient to determine whether any of these groups constitute sensitive populations. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 21 Draft, do not cite or quote D. Interactions No studies were identified that evaluated interactions between brominated acetic acids and chemicals other than water disinfection byproducts. The only endpoint for which mixtures of haloacetic acids have been evaluated experimentally is developmental toxicity in the wholeembryo culture system. As described above, the results of Andrews et al. ( 1999b) in this in vitro test system support the dose additivity suggested by Hunter et al. ( 1996), but these findings have limited utility for prediction of interactions in intact animals. Although QSAR by Richard and Hunter ( 1996) predicts that haloacetic acids would be similar in their mechanisms of action for developmental toxicity and, thus, have the potential for additivity, in vivo rodent developmental toxicity studies have demonstrated marked differences among MBA, BCA, and DBA in the spectrum of induced toxic effects, the chemical potencies associated with these effects, and the critical periods for gestational exposures. Additionally, differences in the genotoxic/ mutagenic potential among MBA, BCA, and DBA suggest that these compounds might exert at least some of their toxic effects via distinct mechanistic pathways.

E. Summary One proposed cellular basis for the toxicity of MBA is its ability to inhibit enzyme activity through direct alkylation of sulfhydryl and amino groups. This hypothesis is supported by in vitro Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 22 Draft, do not cite or quote studies using purified human enzymes ( Gorlatov et al., 1998; Ito et al., 1994; Shapiro et al., 1988;

Whitney, 1970) and some evidence for DNA alkylation ( Stratton et al, 1981), but a direct relationship between such reactions with cellular macromolecules in vivo and the observed toxic effects of MBA has not been established, and there are several limitations in extrapolating from the in vitro data.

DBA and BCA have been associated with liver, kidney, and reproductive and developmental toxicity in a variety of toxicity studies. In short term studies, both BCA ( Parrish et al., 1996; NTP, 1998) and DBA ( Parrish et al., 1996; NTP, 1999) induce changes in liver weight and/ or mild histopathologic alterations, indicating the potential for liver injury with increasing dose and/ or exposure duration. Potential mechanisms for the induction of adverse liver effects include glycogen accumulation ( Kato Weinstein et al., 1998), perturbations of carbohydrate homeostasis ( Bull et al, 2000), or toxicity due to the formation of reactive metabolites from haloacetic acid or tyrosine metabolism pathways ( Austin et al., 1996; Parrish et al., 1996;

Stacpoole et al., 1998; Cornett et al., 1999). The kidney may also be a target for brominated acetic acids ( NTP, 1998; NTP, 1999). This might reflect direct toxicity related to the formation of reactive metabolites as described above for liver toxicity, or may reflect toxicity secondary to oxalate formation ( Kennedy et al., 1993; Webster et al., 2000).

The major area of emphasis for toxicity studies for the brominated acetic acids,

particularly for DBA, has been on potential reproductive effects. MBA did not induce spermatotoxicity in the one available study ( Linder et al., 1994a). DBA induced effects on sperm Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 23 Draft, do not cite or quote development at doses below those causing overt toxicity, and spermatotoxicity at doses causing overt systemic toxicity ( Linder et al., 1994a, 1997a). Although no evidence of BCA induced spermatotoxicity was found in the NTP ( 1998) reproductive and developmental toxicity screening asssay, Luft et al. ( 2000) reported in an abstract that BCA decreased male fertility, and Klinefelter et al. ( 2002a) demonstrated that BCA impaired sperm quality and also reduced male fertility.

These data suggest that both DBA and BCA are male reproductive toxicants. One suggested target for the spermatotoxicity is the Sertoli cells ( Linder et al., 1997a). Although the cellular mechanisms of brominated acetic acid spermatotoxicity have not been identified, the modification of key proteins necessary for Sertoli cell function or direct cytotoxicity by DBA or reactive metabolites might be involved. The results of several studies have suggested that DBA at high doses may act as a reproductive toxicant by interfering with the early stages of steroidogenesis ( Balchak et al., 2000; Goldman and Murr, 2002), possibly by altering the StAR mediated transport of cholesterol within the mitochondrial membrane and thereby affecting the synthesis of pregnenolone. Brominated acetic acids may also interfere with the process of spermatogenesis.

Two sperm proteins, SP22 and SP9, were significantly decreased following 14 day treatment of male rats with BCA, and the shape of the dose response curve for SP22 mirrored that of reduced male fertility observed in these animals ( Klinefelter et al., 2002a). SP22 is a protein that appears to play an important role in the fertilization process ( Klinefelter et al., 2002b), possibly by regulating the androgen receptor ( Takahashi et al., 2001), and it has been suggested that Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 24 Draft, do not cite or quote haloacetic acids acting on SP22 and other sperm proteins may indirectly compromise androgendependent maintenance of spermatogenesis ( Klinefelter et al., 2002b).

All three brominated acetic acids have been reported to induce developmental effects ( Randall et al., 1991; NTP, 1998; Narotsky et al., 1996; Narotsky et al., 1997), although the spectrum of developmental endpoints observed does not necessarily suggest a common mechanism of action for the in vivo studies. Results of whole embryo testing were consistent with a common mechanism of action, because a QSAR model adequately described the potency of a mono and di halogenated series of haloacetic acids ( Hunter et al., 1996; Richard and Hunter,

1996); further testing of haloacetic acid mixtures in whole embryo culture was consistent with the QSAR model predictions ( Andrews et al., 1999b). Brominated acetic acids also induced dysmorphogenesis at doses lower than their known metabolites in the whole embryo testing system ( Hunter et al., 1999), suggesting that the parent compound or unidentified metabolites upstream of glyoxylate are responsible. Ward et al. ( 2000) proposed a potential role of apoptosis induction in developmental toxicity of brominated acetic acids, based on results in a wholeembryo culture system. It should be noted, however, that although the findings from wholeembryo culture systems indicate the potential for developmental toxicity of brominated acetic acids, these in vitro results are limited in their utility to predict both the spectrum of effects and the toxic potencies of these compounds in in vivo animal systems due to the modulating influences of a variety of other physiologic and biochemical processes in intact organisms. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 25 Draft, do not cite or quote There are no data available for identifying susceptible populations. A two generation reproductive study exists for DBA ( Chlorine Chemistry Council, 2001; Christian et al., 2002), but no multi generation studies have been conducted for MBA or BCA. In addition, no data on agedependent changes in the expression of genes involved in brominated acetic acids were found.

Based on the results of in vivo developmental toxicity studies, DBA, but not MBA or BCA,

induced fetal toxicity at doses lower than those associated with maternal effects, suggesting that,

at least for DBA, the fetus might be susceptible. These results were only published in abstracts and, thus, complete study reports are not available to evaluate these findings. However, these preliminary studies found fetal and maternal effects only at doses well above those causing effects on sperm and male reproduction, indicating that protection against the latter effect will also provide adequate protection to children and fetuses.

There are also limited data on potential susceptible populations based on genetic differences. Blackburn et al. ( 2000) characterized human polymorphisms in GST Zeta, which metabolizes DBA and BCA to glyoxylate. However, in the absence of data on whether the parent compound or a metabolite is the active moiety, the functional consequences of this polymorphism with regard to brominated acetic acid toxicity are not clear. Individuals having underlying defects in glycogen storage may be susceptible to liver effects of brominated acetic acids, and individuals lacking certain enzymes of glyoxylate metabolism may be at risk for BCA or DBA induced kidney toxicity. If the formation of reactive oxygen or lipid intermediates is responsible for the toxicity of brominated acetic acids, then deficits in the activity of anti oxidant enzymes could also represent a Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VII 26 Draft, do not cite or quote source of increased susceptibility. All of these possibilities remain speculative, and none has been tested directly in in vivo studies. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 1 Draft, do not cite or quote Chapter VIII. Quantification of Toxicological Effects The quantification of toxicological effects of a chemical consists of separate assessments of noncarcinogenic and carcinogenic health effects. Chemicals that do not produce carcinogenic effects are believed to have a threshold dose below which no adverse, noncarcinogenic health effects occur. Carcinogens are assumed to act without a threshold unless there are data elucidating a nonmutagenic mode of action and demonstrating a threshold for the precursor events that commit a cell to an abnormal tumorigenic response.

A. Introduction to Methods A. 1. Quantification of Noncarcinogenic Effects A. 1.1. Reference Dose In quantification of noncarcinogenic effects, a Reference Dose ( RfD) ( formerly called the Acceptable Daily Intake ( ADI)) is calculated ( U. S. EPA, 2001). The RfD is " an estimate ( with uncertainty spanning approximately an order of magnitude) of a daily exposure to the human population ( including sensitive subgroups) that is likely to be without appreciable risk of deleterious effects over a lifetime" ( U. S. EPA, 1993). The RfD is derived from a no observed Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 2 Draft, do not cite or quote adverse effect level ( NOAEL), lowest observed adverse effect level ( LOAEL), or a NOAEL surrogate such as a benchmark dose identified from a subchronic or chronic study, and divided by a composite uncertainty factor( s). The RfD is calculated as follows:

RfD = NOAEL or LOAEL UF × MF where:

NOAEL = No observed adverse effect level from a high quality toxicological study of an appropriate duration LOAEL = Lowest observed adverse effect level from a high quality toxicological study of an appropriate duration. In situations where there is no NOAEL for a contaminant but there is a LOAEL, the LOAEL can be used for the RfD calculation with the inclusion of an additional uncertainty factor.

UF = Uncertainty factor chosen according to EPA/ NAS guidelines MF = Modifying factor Selection of the uncertainty factor to be employed in calculation of the RfD is based on professional judgment while considering the entire database of toxicological effects for the chemical. To ensure that uncertainty factors are selected and applied in a consistent manner, the Office of Water ( OW) employs a modification to the guidelines proposed by the National Academy of Sciences ( NAS, 1977, 1980). According to the EPA approach ( U. S. EPA, 1993),

uncertainty is broken down into its components, and each component of uncertainty is given a Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 3 Draft, do not cite or quote quantitative rating. The total uncertainty factor is the product of the component uncertainties. The individual components of the uncertainty are as follows:

UF H A factor of 1, 3, or 10 used when extrapolating from valid data in studies using long term exposure to average healthy humans. This factor is intended to account for the variation in sensitivity ( intraspecies variation) among the members of the human population.

UF A An additional factor of 1, 3, or 10 used when extrapolating from valid results of long term studies on experimental animals when results of studies of human exposure are not available or are inadequate. This factor is intended to account for the uncertainty involved in extrapolating from animal data to humans ( interspecies variation).

UF S An additional factor of 1, 3, or 10 used when extrapolating from less thanchronic results on experimental animals when there are no useful long term human data. This factor is intended to account for the uncertainty involved in extrapolating from less than chronic NOAELs to chronic NOAELs.

UF L An additional factor of 1, 3, or 10 used when deriving an RfD from a LOAEL, instead of a NOAEL. This factor is intended to account for the uncertainty involved in extrapolating from LOAELs to NOAELs.

UF D An additional factor of 1, 3, or 10 used when deriving an RfD from an " incomplete" database. This factor is meant to account for the inability of any single type of study to consider all toxic endpoints. The intermediate factor of 3 ( approximately ½ log 10 unit, i. e., the square root of 10) is often used when there is a single data gap exclusive of chronic data. It is often designated as UF D.

On occasion, EPA also uses a modifying factor in the determination of the RfD. A modifying factor is an additional uncertainty factor that is greater than zero and less than or equal to 10. The magnitude of the MF depends upon the professional assessment of scientific Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 4 Draft, do not cite or quote uncertainties of the study and database not explicitly treated above ( e. g., the number of species tested). The default value for the MF is 1.

In establishing the UF or MF, it is recognized that professional scientific judgment must be used. The total product of the uncertainty factors and modifying factor should not exceed 3000. If the assignment of uncertainty results in a UF/ MF product that exceeds 3000, then the database does not support development of an RfD. The quantification of toxicological effects of a chemical consists of separate assessments of noncarcinogenic and carcinogenic health effects.

Unless otherwise specified, chemicals which do not produce carcinogenic effects are believed to have a threshold dose below which no adverse, noncarcinogenic health effects occur, while carcinogens are assumed to act without a threshold.

A. 1.2. Drinking Water Equivalent Level The drinking water equivalent ( DWEL) is calculated from the RfD. The DWEL represents a drinking water specific lifetime exposure at which adverse, noncarcinogenic health effects are not anticipated to occur. The DWEL assumes 100% exposure from drinking water. The DWEL provides the noncarcinogenic health effects basis for establishing a drinking water standard. For ingestion data, the DWEL is derived as follows: Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 5 Draft, do not cite or quote DWEL = ( RfD) × BW WI where:

BW = 70 kg adult body weight WI = Drinking water intake ( 2 L/ day)

A. 1.3. Health Advisory Values In addition to the RfD and the DWEL, EPA calculates Health Advisory ( HA) values for noncancer effects. HAs are determined for lifetime exposures as well as for exposures of shorter duration ( 1 day, 10 day, and longer term). The shorter duration HA values are used as informal guidance to municipalities and other organizations when emergency spills or contamination situations occur. The lifetime HA becomes the MCLG for a chemical that is not a carcinogen.

The shorter term HAs are calculated using an equation similar to the RfD and DWEL;

however, the NOAELs or LOAELs are derived from acute or subchronic studies and identify a sensitive, noncarcinogenic endpoint of toxicity. The HAs are derived as follows:

HA = NOAEL or LOAEL × BW UF × WI where:

NOAEL or LOAEL = No or lowest observed adverse effect level in mg/ kg bw/ day BW = Assumed body weight of a child ( 10 kg) or an adult ( 70 kg) Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 6 Draft, do not cite or quote UF = Uncertainty factor, in accordance with EPA or NAS/ OW guidelines WI = Assumed daily water intake of a child ( 1 L/ day) or an adult ( 2 L/ day)

Using the above equation, the following drinking water HAs are developed for noncarcinogenic effects:

° 1 day HA for a 10 kg child ingesting 1 L water per day.

° 10 day HA for a 10 kg child ingesting 1 L water per day.

° Longer term HA for a 10 kg child ingesting 1 L water per day.

° Longer term HA for a 70 kg adult ingesting 2 L water per day.

Each of these shorter term HA values assumes that the total exposure to the contaminant comes from drinking water.

The lifetime HA is calculated from the DWEL, and takes into account exposure from sources other than drinking water. It is calculated using the following equation:

Lifetime HA = DWEL × RSC where:

DWEL= Drinking water equivalent level RSC = Relative source contribution. The fraction of the total exposure allocated to drinking water. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 7 Draft, do not cite or quote A. 2 Quantification of Carcinogenic Effects Under the new U. S. EPA ( 1999) draft cancer risk assessment guidelines, the U. S. EPA assesses the carcinogenic potential of a chemical compound in a narrative characterization, and uses one of the following five standard descriptors to express the conclusion regarding the weight of evidence for carcinogenic hazard potential:

° Carcinogenic to Humans ° Likely to be Carcinogenic to Humans ° Suggestive Evidence of Carcinogenic Potential ° Inadequate Information to Assess Carcinogenic Potential ° Not Likely to be Carcinogenic to Humans Each standard descriptor is presented only in the context of a chemical specific, weight

ofevidence narrative. Additionally, more than one conclusion may be reached for an agent ( e. g., an agent is " likely to carcinogenic" by inhalation exposure and " not likely to be carcinogenic" by oral exposure. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 8 Draft, do not cite or quote If toxicological evidence leads to the classification of the contaminant as a genotoxic,

probable or possible human carcinogen, mathematical models are used to calculate the estimated excess cancer risk associated with ingestion of the contaminant in drinking water. The data used in these estimates usually come from lifetime exposure studies in animals. In order to predict the risk for humans from animal data, animal doses must be converted to equivalent human doses.

This conversion includes correction for noncontinuous exposure, less than lifetime studies and differences in size. It is assumed that the average adult human body weight is 70 kg and that the average water consumption of an adult human is two liters of water per day.

For contaminants with a carcinogenic potential, chemical levels are correlated with a carcinogenic risk estimate by employing a cancer potency ( unit risk) value together with the assumption for lifetime exposure via ingestion of water. Under the 1986 Carcinogen Risk Assessment Guidelines, the cancer unit risk was usually derived from a linearized multistage model with a 95% upper confidence limit providing a low dose estimate; that is, the true risk to humans, while not identifiable, is not likely to exceed the upper limit estimate and, in fact, may be lower. Excess cancer risk estimates may also be calculated using other models such as the one hit,

Weibull, logit and probit models. There is little basis in the current understanding of the biological mechanisms involved in cancer to suggest that any one of these models is able to predict risk more accurately than any of the others. Because each model is based upon differing assumptions, the estimates that are derived for each model can differ by several orders of magnitude. Drinking Water Criteria Document for Brominated Acetic Acids 2 A" point of departure" ( POD) marks the beginning of extrapolation to lower doses. The POD is an estimated dose ( expressed inhuman equivalent terms) near the lower end of the observed range, without significant extrapolation to lower doses ( U. S. EPA, 2003).

EPA/ OW/ OST/ HECD VIII 9 Draft, do not cite or quote Under the new U. S. EPA ( 1999) draft cancer risk assessment guidelines, dose response assessment is performed in two steps: assessment of observed experimental data to derive a point of departure ( POD) 2, followed by extrapolation to lower exposure to the extent that is necessary for environmental exposures of interest. Extrapolation is based on extension of a biologicallybased model if supported by substantial data. Otherwise, default approaches can be applied that are consistent with current understanding of mode( s) of action of the agent. These approaches may assume either linearity or nonlinearity of the dose response relationship, or both. The linear approach is used when there is an absence of sufficient information on modes of action or the mode of action information indicates that the dose response curve a low dose is or is expected to be linear. A range of models may be used for the linear approach. A default approach for nonlinearity can be to use a reference dose or a reference concentration ( U. S. EPA, 1999).

The scientific data base used to calculate and support the setting of cancer risk rates has an inherent uncertainty due to the systematic and random errors in scientific measurement. In most cases, only studies using experimental animals have been performed. Thus, there is uncertainty when the data are extrapolated to humans. When developing cancer risk rates, several other areas of uncertainty exist, such as the incomplete knowledge concerning the health effects of contaminants in drinking water, the impact of the experimental animal's age, sex, and species, the nature of the target organ system( s) examined and the actual rate of exposure of the internal Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 10 Draft, do not cite or quote targets in experimental animals or humans. Dose response data usually are available only for high levels of exposure, not for the lower levels of exposure at which a standard may be set. When there is exposure to more than one contaminant, additional uncertainty results from a lack of information about possible synergistic or antagonistic effects. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 11 Draft, do not cite or quote B. Noncarcinogenic Effects B. 1 Monobromoacetic acid Table VIII 1 summarizes the available studies on the oral toxicity of MBA.

Table VIII 1. Summary of Oral Studies of MBA Toxicity Reference Species Route Exposure Duration; Doses Endpoints NOAEL LOAEL Comments ( mg/ kg/ day)

Linder et al., 1994a Sprague Dawley Rat ( male) Oral Gavage in water Acute Single dose; 100 to 200 mg/ kg Lethality, clinical observation  LD50 177 mg/ kg Doses not specified.

Linder et al., 1994a Sprague Dawley Rat ( male) Oral Gavage in water Acute Single dose; 0, 100 mg/ kg Sperm analysis, reproductive tract histopathology 100  None Randall et al., 1991 Long Evans Rat ( female) Oral Gavage in water Gestation day 6 15; 0, 25, 50, 100 mg/ kg/ day Decreased maternal weight gain, decreased live fetus size, increased incidence of softtissue malformations 50 100 Published abstract does not provide adequate details for definitive review.

Linder et al., 1994a Sprague Dawley Rat ( male) Oral Gavage in water 14 day; 0, 25 mg/ kg/ day Sperm analysis, reproductive tract histopathology 25  Only listed endpoints were evaluated Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 12 Draft, do not cite or quote B. 1.1 One Day Health Advisory for MBA The oral toxicity data for MBA are very limited. Linder et al. ( 1994a) reported an LD 50 of 177 mg/ kg in Sprague Dawley rats. Clinical signs included excess drinking, hypomobility, labored breathing, and diarrhea. However, LD 50 studies are not suitable for the development of one day health advisories. The only other single dose study was reported by these same authors, where male Sprague Dawley rats were given 0 or 100 mg/ kg/ day MBA and were evaluated for evidence of spermatotoxicity. No other endpoints were evaluated. The single dose tested was the approximate LD 01 and the NOAEL for male reproductive effects in this study. Due to the absence of an observed effect, the use of a single dose precluding assessment of the dose response, and the limited endpoints evaluated, this study is not sufficient for derivation of a One day health advisory.

B. 1.2 Ten Day Health Advisory for MBA Two studies of appropriate duration were identified for derivation of a Ten day health advisory. However, limited details or inadequate study designs preclude their use. In a published abstract, Randall et al. ( 1991) evaluated the developmental toxicity of MBA in female Long

Evans rats dosed with 0, 25, 50, or 100 mg/ kg/ day MBA on gestation days 6 15. The authors reported decreased maternal weight gain, decreased live fetus size, and increased incidence of Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 13 Draft, do not cite or quote soft tissue malformations at the highest dose. Based on these data, the NOAEL would be 50 mg/ kg/ day and the LOAEL for maternal and developmental effects would be 100 mg/ kg/ day.

However, this study is available only as a published abstract that has not undergone scientific peer review. Thus, the data are to be viewed as preliminary, and the results of this study are not sufficient for derivation of a health advisory. A single published study of adequate duration was identified for MBA. In this study, no effects on the male reproductive tract were seen in male rats treated with 0 or 25 mg/ kg/ day MBA by gavage in water for 14 days ( Linder et al., 1994a); no LOAEL was identified. The absence of an identified effect, the use of a single dose level precluding assessment of the dose response, and the limited array of endpoints examined preclude the use of this study as the basis for derivation of the Ten day health advisory. In light of the qualitative differences in toxicity between MBA and DBA, it is not appropriate to assume that male reproductive toxicity is the most sensitive endpoint for MBA, and it is, therefore, not appropriate to consider this free standing NOAEL for a limited number of endpoints as sufficiently protective for general toxicity, nor to use it for derivation of a health advisory.

B. 1.3 Longer Term Health Advisory for MBA There are no studies of suitable duration for derivation of a Longer Term Health Advisory for MBA. Developmental toxicity data, such as that reported in an abstract by Randall et al.

( 1991), are appropriate for the derivation of Longer Term health advisories only if systemic Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 14 Draft, do not cite or quote toxicity studies of adequate duration have been conducted and show that developmental toxicity is the most sensitive endpoint. In the absence of such systemic toxicity studies for MBA, no Longer

Term Health Advisory can be derived.

B. 1.4 Reference Dose and Drinking Water Equivalent Level for MBA There are no studies ( subchronic or chronic toxicity studies that evaluate a range of systemic endpoints) suitable for derivation of an RfD for MBA. As for the Longer Term Health Advisory, developmental toxicity data are not appropriate for the derivation of an RfD in the absence of systemic toxicity studies of adequate duration. In addition, the only available developmental toxicity study ( Randall et al., 1991) was published as an abstract.

B. 2 Bromochloroacetic acid Table VIII 2 summarizes the available studies on the oral toxicity of BCA. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 15 Draft, do not cite or quote Table VIII 2. Summary of Oral Studies of BCA Toxicity Reference Species Route Exposure Duration; Doses Endpoints NOAEL LOAEL Comments ( mg/ kg/ day)

NTP, 1998 Sprague Dawley Rat ( male and female) Drinking Water 14 day; 0, 3, 10, 28, 41 mg/ kg/ day Clinical observation, body weight, body weight gain 41  None Luft et al., 2000 C57BL/ 6 Mouse ( male) Oral Gavage in water 14 day; 0, 8, 24, 72, 216 mg/ kg/ day Decrease in mean number of litters per male, decreased percent of litters per bred female 24 72 Published abstract does not provide adequate details for definitive review.

Results of histopathology analysis were not reported.

Klinefelter et al., ( 2002a), manuscript Sprague Dawley Rat ( male) Oral, Gavage in water 14 day, 0, 24, 72, 216, mg/ kg/ day ( range finding study) Decreased serum hormone levels, sperm abnormalities, altered spermiation  24 The LOAEL was the lowest dose tested.

Klinefelter et al., ( 2002a), manuscript Sprague Dawley Rat ( male) Oral, Gavage in water 14 day, 0, 8, 24, 72 mg/ kg/ day ( dose response study) Decreased progressive sperm motility, significantly reduced fertility, decrease in sperm protein SP22.  8 Decreased serum hormone levels not observed.

Decrease in sperm protein SP22 paralleled the reduction in fertility.

The LOAEL was the lowest dose tested. Drinking Water Criteria Document for Brominated Acetic Acids Table VIII 2. Summary of Oral Studies of BCA Toxicity Reference Species Route Exposure Duration; Doses Endpoints NOAEL LOAEL Comments ( mg/ kg/ day)

EPA/ OW/ OST/ HECD VIII 16 Draft, do not cite or quote Parrish et al., 1996 B6C3F1 Mouse ( male) Drinking water 21 days; 0, 25, 125, 500 mg/ kg/ day Increased liver weight, increased oxidative DNA damage 125 500 500 mg/ kg/ day is considered a marginal LOAEL.

Doses were calculated from default waterintake estimates.

NTP, 1998 Sprague Dawley Rat ( male) Drinkin g water 30 or 26 days; 0, 5, 15, 39 mg/ kg/ day Liver weight and histopathology , sperm quality 15 39 Marginal LOAEL; males had marginal liver weight and histopathology changes Drinking Water Criteria Document for Brominated Acetic Acids Table VIII 2. Summary of Oral Studies of BCA Toxicity Reference Species Route Exposure Duration; Doses Endpoints NOAEL LOAEL Comments ( mg/ kg/ day)

EPA/ OW/ OST/ HECD VIII 17 Draft, do not cite or quote NTP, 1998 Sprague Dawley Rat ( female) Drinking water Group A: 34 days periconception ( 12 days premating up to 5 days cohabitation, and up to 21 days gestation); 0, 6, 19, 50 mg/ kg/ day Group C: 30 days periconception ( 12 days premating up to 5 days cohabitation, and up to 16 days gestation); 0, 6, 19, 50 mg/ kg/ day Group B: 17 days  gestation day 6 to postnatal day 1; 0, 10, 25, 61 mg/ kg/ day Decreased live fetuses per litter and decreased total implants per litter Decreased live fetuses per litter, decreased total implants per litter, kidney histopathology changes Maternal and fetal toxicity 19 19 61 50 50  Effects on fetuses were based on a pooled analysis of Group A and Group C data Kidney histopathological changes are considered as a equivocal NOAEL at 19 mg/ kg/ day.

Increase in postimplantation loss was not statistically significant and lacked a doseresponse Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 18 Draft, do not cite or quote B. 2.1 One Day Health Advisory for BCA No studies of suitable duration were identified for derivation of a One day health advisory for BCA.

B. 2.2. Ten Day Health Advisory for BCA Five studies of suitable duration were identified for derivation of a Ten day health advisory for BCA. However, one of these studies was available only as a published abstract ( Luft et al., 2000) and two of the available studies either identified only marginal effects or were not designed to evaluate systemic toxicity ( Parrish et al., 1996; dose range finding study by NTP,

1998). The reproductive/ developmental effects in the remaining two studies of BCA ( NTP, 1998;

Klinefelter et al., 2002a) were not considered to be suitable for the derivation of a 10 Day health advisory for a 10 kg child, because the study was conducted on sexually mature animals.

Although equivocal liver effects ( marginal increases in liver weights and marginal histopathology)

were observed in the NTP ( 1998) study, there was no dose response and no effect on hepatic labeling index indicative of cellular proliferation and regeneration. Further, similar marginal histopathology ( i. e., mild cytoplasmic vacuolization) was also observed in control animals. A nonstatistically significant increase in kidney tubule dilatation/ degeneration was also observed in dosed animals, but it was unclear whether these findings were treatment related. Other organs and Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 19 Draft, do not cite or quote endpoints were not evaluated in this study and sample sizes were small, thus limiting the statistical power of this experiment. These studies are discussed in detail in Chapter V. The limitations of these studies, including low statistical power and toxicity evaluation of only a small number of endpoints, preclude their use in the derivation of a Ten day health advisory.

B. 2.3. Longer Term Health Advisory for BCA The toxicity database for BCA is very limited. There are no studies of sufficient duration for derivation of Longer term health advisory for BCA. As noted for MBA and in the previous section for BCA, the developmental effects noted in the NTP ( 1998) study are inadequate as the basis for the Longer term health advisory, in the absence of a subchronic study that adequately evaluated systemic toxicity. No multi generation reproductive toxicity study has been conducted.

Subchronic and chronic toxicity testing of BCA is planned or in progress ( NTP, 2000b).

B. 2.4 Reference Dose and Drinking Water Equivalent Level for BCA As discussed in the previous section on the Longer term health advisory, the toxicity database for BCA is currently limited and there are no suitable studies of appropriate design and duration to derive an oral RfD at this time. Drinking Water Criteria Document for Brominated Acetic Acids EPA/ OW/ OST/ HECD VIII 20 Draft, do not cite or quote B. 3 DBA Table VIII 3 summarizes the available studies on the oral toxicity of DBA.

Table VIII 3. Summary of Oral Studies of DBA Toxicity Reference Species Route Exposure Duration; Doses Endpoints NOAEL LOAEL Comments mg/ kg/ day General Toxicity Studies ( by duration of treatment)

Linder et al., 1994a Sprague Dawley rat ( male) Oral Gavage in water Acute single dose; 1000 to 2000 mg/ kg Lethality, clinical observation  LD50 1737 mg/ kg Doses not specified.

Parrish et al., 1996 B6C3F1 mouse ( male) Drinking water 21 days; 0, 25, 125, 500 mg/ kg/ day Increased liver weight, oxidative DNA damage 25 125 Doses were calculated from default water intake estimates NTP, 1999 B6C3F1 mouse ( female) Drinking water 28 days;

Study ( 1); 0, 19, 39, 73, 150, 285 mg/ kg/ day Study ( 2); 0, 20, 38, 70, 143, 280 mg/ kg/ day Study ( 3); 0, 16, 35, 69, 134, 229 mg/ kg/ day Study ( 4); 0, 14, 33, 68, 132, 236 mg/ kg/ day Decreased antibody forming cell response 38 70 Absolute and relative liver weight were increased beginning at 14 mg/ kg/ day. This was not chosen as the critical effect in the absence of histopathology or clinical chemistry data to confirm that the effect was adverse.