United States Office of Science November 2005
Environmental Protection and Technology
Agency Washington, D.C.
s> EPA Office of Water
Drinking Water Criteria Document
Brominated Acetic Acids
EPA-822-R-05-007
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Acknowledgments
Chemical Manager/Lead Scientist:
Steven S. Kueberuwa, MS (OST/OW)
Contractor Authors:
Lynne Haber Ph.D. (TERA)
Bonnie Stern, M.P.H., Ph.D. (GRAM, Inc.)
Claudine Kasunic (GRAM, Inc.)
EPA Internal Reviewers:
John Lipscomb, Ph.D., DABT (NCEA/ORD)
Linda Teuschler, Ph.D. (NCEA/ORD)
OST Mail Code 4304T
EPA-822-R-03-015
Title Drinking Water Criteria Document for Brominated Acetic Acids: Final
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table of Contents
Acknowledgments 1-ii
List of Figures 1-vi
List of Tables 1-viii
Chapter I. Executive Summary 1-13
Chapter II. Physical and Chemical Properties II-1
Chapter III. Toxicokinetics ffl-1
A. Absorption ffl-1
B. Distribution ffl-6
C. Metabolism ffl-11
D. Excretion ffl-17
E. Bioaccumulation and Retention ffl-20
F. Summary ffl-23
Chapter IV. Human Exposure IV-1
A. Drinking Water Exposure IV-1
A.I National Occurrence Data for MBA, BCA, and DBA IV-1
A.I.I ICRPlants IV-2
A. 1.2 Quarterly Distribution System Average and Highest Value
for MBA, BCA, and DBA IV-3
A.2 Factors Affecting the Relative Concentrations of MBA, BCA,
and DBA IV-7
A.2.1 Disinfection Treatment IV-8
A.2.1.1 Disinfection Treatment in ICR Data Base . . IV-13
A.2.2 Bromide Concentration IV-21
A.2.2.1 Bromide Concentration in ICR Data Base . . . IV-22
A.2.3 Total Organic Carbon (TOC) Concentration in ICR
Database IV-31
A.2.4 Seasonal Shifts IV-39
A.2.4.1 Seasonal Shifts in ICR Database IV-40
B. Exposure to Sources Other Than Drinking Water IV-45
C. Overall Exposure IV-46
D. Body Burden IV-47
E. Summary IV-47
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Chapter V. Health Effects in Animals V-l
A. Short-Term Exposure V-l
B. Long-Term Exposure V-23
C. Reproductive and Developmental Effects V-25
D. Mutagenicity and Genotoxicity V-80
E. Carcinogenicity V-94
F. Summary V-95
Chapter VI. Health Effects in Humans VI-1
Chapter VII. Mechanisms of Toxicity VII-1
A. Mechanisms of Noncancer Toxicity VII-1
B. Cancer Mechanisms VII-18
C. Sensitive Subpopulations VII-19
D. Interactions VII-24
E. Summary VII-25
Chapter VIII. Quantification of Toxicological Effects VIII-1
A. Introduction to Methods VIII-1
A. 1. Quantification of Noncarcinogenie Effects VIII-1
A.I.I. Reference Dose VIE-1
A. 1.2. Drinking Water Equivalent Level Vffl-4
A.1.3. Health Advisory Values VIII-5
B. Noncarcino genie Effects VIII-10
B. 1 Monobromoacetic acid VIII-10
B.I.I One-Day Health Advisory for MBA VIH-ll
B.1.2 Ten-Day Health Advisory for MBA VIH-12
B.I.3 Longer-Term Health Advisory for MBA VIH-13
B.I .4 Reference Dose and Drinking Water Equivalent Level
for MBA Vin-14
B.2 Bromochloroacetic Acid Vffl-14
B.2.1 One-Day Health Advisory for BCA VIH-19
B.2.2. Ten-Day Health Advisory for BCA VIH-19
B.2.3. Longer-Term Health Advisory for BCA VIII-20
B.2.4 Reference Dose and Drinking Water Equivalent Level
for BCA Vin-20
B.3 Dibromoacetic Acid Vffl-21
B.3.1 One-Day Health Advisory for DBA VIII-27
B.3.2 Ten-Day Health Advisory for DBA VIII-28
B.3.3 Longer-Term Health Advisory for DBA VIII-30
B.3.4. Reference Dose and Drinking Water Equivalent Level
-------
Drinking Water Criteria Document for Brominated Acetic Acids
for DBA Vin-34
C. Carcinogenic Effects VIH-34
C. 1. Monobromoacetic acid VIII-35
C.2. Bromochloroacetic acid VIII-36
C.3. Dibromoacetic acid VIII-37
D. Summary VIII-38
Chapter IX. References IX-1
-------
Drinking Water Criteria Document for Brominated Acetic Acids
List of Figures
Figure II-1. The Chemical Structures of MBA, BCA, and DBA II-3
Figure III-l. Proposed Metabolism of DBA" Ill-14
-------
Drinking Water Criteria Document for Brominated Acetic Acids
List of Tables
Table n-1. Physical and Chemical Properties of Brominated Acetic Acids II-4
Table HI-1. Toxicokinetic Data for BCA and DBA in F344 Rats IH-3
Figure III-l. Proposed Metabolism of DBAa Ill-14
Table IV-1. Bromoacetic Acids
Quarterly Distribution System Average and Highest Value IV-5
Table IV-2. MBA by Disinfection Method
(Quarterly Distribution System Average) IV-14
Table IV-3. BCA Acid by Disinfection Method
(Quarterly Distribution System Average) IV-16
Table IV-4. DBA by Disinfection Method
(Quarterly Distribution System Average) IV-18
Table IV-5. MBA by Influent Bromide Concentration
(Quarterly Distribution System Average) IV-24
Table IV-6. BCA by Influent Bromide Concentration
(Quarterly Distribution System Average) IV-26
Table IV-7. DBA by Influent Bromide Concentration
(Quarterly Distribution System Average) IV-2 8
Table IV-8. MBA by Influent Total Organic Carbon (TOC) Concentration
(Quarterly Distribution System Average) IV-3 4
Table IV-9. BCA by Influent Total Organic Carbon (TOC) Concentration
(Quarterly Distribution System Average) IV-3 6
Table IV-10. DBA by Influent Total Organic Carbon (TOC) Concentration
(Quarterly Distribution System Average) IV-3 8
Table IV-11. MBA by Sample Quarter
(Quarterly Distribution System Average) IV-42
Table IV-12. BCA by Sample Quarter
(Quarterly Distribution System Average) IV-43
Table IV-13. DBA by Sample Quarter
(Quarterly Distribution System Average) IV-44
Table V-l. Body and Liver Weight Changes Induced by BCA and DBA V-9
Table V-2. Immunotoxicity of DBA in Female B6C3F1 mice V-15
Table V-3. General toxicity of DBA in Female B6C3F1 mice V-16
Table V-4. Reproductive and Developmental Toxicity of BCA following Peri-conception
Exposure (Combined Data for Female Groups A and C) V-32
Table V-5. Sperm Quality Parameters in Rats Given 14 Daily Doses of DBA V-46
Table V-6. Reproductive Outcomes in Rats Following Oral Dosing with DBA V-49
Table V-7. Outcome of Artificial Insemination of Sperm from Rats
Dosed with DBA V-50
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table V-8. Reproductive Organ Weights and Sperm Counts
in Rats Given Daily Doses of DBA V-51
Table V-9. Sperm Quality Parameters in Rats Given Daily Doses of DBA V-53
Table V-10. Average Consumed Daily Doses (mg/kg/day) for Male and Female Sprague-Dawley
Rats in the Two-Generation Reproductive/Developmental Toxicity Study V-68
Table V-l 1. Incidences of Exposure-Related Histopathologic Findings in the Testes of Rats
Consuming DBA in Drinking Water V-71
Table V-12. Genotoxicity Studies of MBA V-83
Table V-13. Genotoxicity Studies of DBA V-93
Table V-l 4. Summary of Genotoxicity Data for Brominated Acetic Acids V-94
Table Vffl-1. Summary of Oral Studies of MBA Toxicity VIH-ll
Table Vffl-2. Summary of Oral Studies of BCA Toxicity VIH-14
Table Vffl-3. Summary of Oral Studies of DBA Toxicity VIII-21
-------
Drinking Water Criteria Document for Brominated Acetic Acids
FOREWORD
The Safe Drinking Water Act, as amended in 1996, requires the Administrator of the U.S.
Environmental Protection Agency (EPA) to publish maximum contaminant level goals (MCLGs)
and promulgate National Primary Drinking Water Regulations for each contaminant that, in the
judgment of the Administrator, may have an adverse effect on public health and that is known or
anticipated to occur in public water systems. The MCLG is non-enforceable and is set at a level
at which no known or anticipated adverse health effects in humans occur and which allows for
an adequate margin of safety. Factors considered in setting the MCLG include health effects
data and sources of exposure other than drinking water.
This document provides the health effects basis to be considered in establishing the
MCLG for dibromoacetic acid. To achieve this objective, data on pharmacokinetics, human
exposure, acute and chronic toxicity to animals and humans, epidemiology, and mechanisms of
toxicity were evaluated. Specific emphasis is placed on data providing dose-response
information. Thus, although the literature search and evaluation performed in support of this
document were comprehensive, only the reports considered most pertinent in the derivation of
the MCLG are cited in this document. The comprehensive literature search in support of this
document includes information published up to February 2005; however, more recent
information may have been added during the review process.
When adequate health effects data exist, Health Advisory values for less than lifetime
exposure(l-day, 10-day, and longer term, approximately 10% of an individual's lifetime) are
included in this document. These values are not used in setting the MCLG, but serve as informal
guidance to municipalities and other organizations when emergency spills or contamination
situations occur. The Reference Dose (RfD) provides information on long-term toxic effects
other than carcinogenicity. The RfD is based on the assumption that thresholds exist for certain
toxic effects such as cellular necrosis, but may not exist for other toxic effects such as some
carcinogenic responses. It is expressed in terms of milligrams per kilogram per day (mg/kg/day).
In general, the RfD is an estimate (with uncertainty spanning perhaps an order of magnitude) of a
daily exposure to the human population (including sensitive subgroups) that is likely to be
without an appreciable risk of deleterious effects during a lifetime. The RfD is used in
establishing the Lifetime Health Advisory for noncancer effects.
The carcinogenicity assessment provides information on two aspects of the carcinogenic
risk assessment for the agent in question: (1) the EPA classification and (2) quantitative estimates
of risk from oral exposure. The classification reflects a weight-of-evidence judgment of the
likelihood that the agent is a human carcinogen and the conditions under which the carcinogenic
effects maybe expressed. Quantitative risk estimates are presented in three ways. The slope
factor is the result of the application of a low-dose extrapolation procedure and is presented as
the risk per mg/kg/day. The unit risk is the quantitative estimate in terms of risk per micrograms
per liter (|-ig/L) drinking water. The third form in which risk is presented is a drinking water
concentration providing cancer risks of 1 in 10,000, 1 in 100,000, or 1 inl,000,000.
EPA/OW/OST/HECD
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Development of the hazard identification and dose-response assessments for
dibromoacetic acid has followed the general guidelines for risk assessments as set forth by the
National Research Council (1983) and The Presidential/Congressional Commission on Risk
Assessment and Risk Management (1997). Other guidelines that were used in the development of
this assessment include the following: Guidelines for Carcinogen Risk Assessment (U.S. EPA,
2005), Guidelines for Developmental Toxicity Risk Assessment (U.S. EPA, 1991), Guidelines
for Reproductive Toxicity Risk Assessment (U.S. EPA, 1996b), Guidelines for Neurotoxicity
Risk Assessment (U.S. EPA, 1998), Recommendations for and Documentation of Biological
Values for Use in Risk Assessment (U.S. EPA, 1988), and Health Effects Testing Guidelines
(U.S. EPA,1997).
EPA/OW/OST/HECD
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Chapter I. Executive Summary
Three brominated acetic acids, monobromoacetic acid (MBA), dibromoacetic acid
(DBA), and bromochloroacetic acid (BCA), have been selected for evaluation in this drinking
water criteria document on the basis of (1) their occurrence in drinking water as chlorine
disinfection byproducts, and (2) the availability of toxicological data on their potential human
health effects. MBA, BCA, and DBA are water-soluble hygroscopic crystals in pure form, and
are very soluble in water. Brominated acetic acids are formed during ozonation or chlorination of
water that contains bromide ions and organic matter, primarily humic and fulvic acids.
Formation of chlorinated acetic acids is higher in the presence of humic acid fractions of water
than in the presence of fulvic acid, suggesting that a similar relationship may hold for brominated
acetic acids. Bromide ions occur naturally in surface water and ground water, with seasonal
fluctuations, and may increase due to saltwater intrusions under conditions of drought or as a
result of pollution. In the presence of sufficient concentrations of bromide ion, the formation of
brominated compounds may be favored over formation of chlorinated compounds. Brominated
acetic acid concentrations in drinking water are typically in the order of BCA>DBA>MBA.
No toxicokinetic studies of MBA have been identified in the literature. Although
quantitative information on BCA and DBA toxicokinetics is limited to the findings in a single
comparative toxicokinetic study with rats, the data demonstrate that both compounds are rapidly
absorbed from the gastrointestinal tract, almost completely metabolized, and minimally excreted
in the urine and feces. Following a single intravenous dose, neither BCA nor DBA appeared to
EPA/OW/OST/HECD 1-1
-------
Drinking Water Criteria Document for Brominated Acetic Acids
bind significantly to plasma proteins or accumulate in blood cells, and the unbound fraction in
plasma and the plasma:blood concentrations was close to unity. Further, the apparent volume of
distribution was similar to the total body water volume for rats, leading the authors to conclude
that both compounds were uniformly distributed outside the vascular system and did not
sequester in peripheral tissues. However, in the absence of specific tissue measurements, the
distribution of BCA and DBA cannot be ascertained. The mechanisms by which brominated
acetic acids are metabolized remains unclear. Potential pathways of brominated acetic acid
metabolism to glyoxylic acid have been proposed based on the observed metabolism of 1,1,2,2-
tetrabromoethane, and based on analogy to chlorinated acetic acids Metabolic data from a
number of studies demonstrate that chlorinated acetic acids undergo oxidative dehalogenation by
glutathione transferase zeta (GST-Zeta) activity and preliminary data indicate that a similar
metabolic pathway is likely to occur for the brominated acetic acids. It is not clear whether the
lexicologically effective moiety is the parent compound or an active metabolite. Both BCA and
DBA are rapidly cleared from the blood, following single oral or intravenous dosing, although
these data are inconsistent with the results of repeated-exposure drinking water studies. Based on
current information, brominated acetic acids appear to be rapidly excreted and to have little
propensity for bioaccumulation. DBA administered at high concentrations to pregnant Sprague-
Dawley females was reliably measured in placental tissue and in fetal plasma at concentrations
that were generally similar to those measured in maternal plasma. However, quantifiable levels
of DBA in the milk of the lactating rats were not detected, leading to the conclusion that DBA
freely crosses the placenta and distributes to the fetus during gestation, but does not appear to
EPA/OW/OST/HECD 1-2
-------
Drinking Water Criteria Document for Brominated Acetic Acids
bioaccumulate. In contrast, preliminary data from a published abstract reported the presence of
DBA in the milk of lactating female rats at concentrations higher than those in blood serum,
suggesting to the authors that DBA might accumulate in milk.
EPA's Information Collection Rule (ICR) database contains extensive information on
concentrations of MBA, BCA, and DBA in drinking-water systems, and on how those
concentrations vary with input-water characteristics and treatment methods. The database
contains information from six quarterly samples from 7/97 to 12/98, from approximately 300
large systems covering approximately 500 plants. The mean concentrations of BCA were 1.47
and 3.61 |-ig/L from groundwater and surface water respectively. The mean concentrations of
DBA were 0.82 and 1.09 |-ig/L in groundwater and surface water, respectively. Statistical
analysis of these data indicated that the mean concentrations of MBA, BCA, and DBA in surface
water were significantly higher than the mean concentrations of these chemicals in groundwater,
with BCA > DBA > MBA in both surface water and groundwater.
The concentrations of MBA in surface water treated with chlorine were similar to those
treated with chlorine followed by chloramine. BCA and DBA concentrations were lower when
free chlorine was used both in the treatment plant and the distribution system. Although
ozonation appeared to significantly reduce the formation of BCA, there were no significant
differences in MBA or DBA concentrations with the use of ozone in treating surface water as
compared to the common (non-ozonation) chemical-disinfection processes. In addition there
were no significant differences between the two treatments using ozonation in treating surface
water for MBA, BCA and DBA.
EPA/OW/OST/HECD 1-3
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Consistent with the findings of other investigators, and the chemistry of the formation of
bromoacetic acids, a regression analysis of the ICR data indicated that, with the exception of
MBA in surface water, there was a significant correlation between influent bromide
concentration and the mean concentrations of BCA and DBA in surface water and groundwater.
In addition, for a given influent bromide concentration range, the mean concentrations of BCA
were generally significantly higher that the mean concentrations of DBA and MBA in both
surface water and groundwater.
A regression analysis of the ICR data indicated that there was a significant correlation
between influent total organic carbon (TOC) concentration and the mean concentrations of MBA,
BCA, and DBA in surface water. This is consistent with the formation of brominated acetic
acids from the reaction of humic acid and hypobromous acid, a compound formed by the reaction
of bromide ion with ozone and/or chlorine in the disinfection process. In addition, for a given
influent TOC concentration range in surface water, the mean concentrations of BCA were
significantly higher than the mean concentrations of DBA, which were significantly higher than
the MBA mean concentrations.
Based on only two seasons of monitoring, statistical analysis indicated that the mean
concentrations of MBA in surface water were significantly higher in the summer than in the
spring Also, based on only two seasons of monitoring, the mean concentrations of BCA in
surface water were higher in summer than in winter. Aside from these exceptions, there were no
consistently significant differences in the mean concentrations of MBA, BCA or DBA between
one season and another in either surface water or groundwater. Seasonal variations in
EPA/OW/OST/HECD 1-4
-------
Drinking Water Criteria Document for Brominated Acetic Acids
brominated acetic acids may be dependent on seasonal fluctuations in bromide-ion concentration,
which were not monitored.
The data on exposure to sources other than drinking water are limited, but MBA has been
used in industry and in hospitals. Between 1981 to 1983, approximately 5000 workers were
potentially exposed to MBA.. No data were located on exposure to MBA, BCA, or DBA in
food, air, or via dermal exposure. No data could be located on body burden levels of MBA, BCA,
or DBA..
The available toxicity database for the brominated acetic acids is limited and many
toxicity endpoints have not been fully explored. However, there is a large body of ongoing work,
particularly for BCA and DBA. Preliminary results for many studies have been reported in
published abstracts and are included in this document to provide a sense of the spectrum of
effects induced by the brominated acetic acids.
Monobromoacetic acid
The toxicity data for MBA are very limited. The oral LD50 for MBA was reported as 177
mg/kg in male rats. Oral gavage single-dose (0 or 100 mg/kg) and 14-day studies (0 or 25
mg/kg/day) have been conducted to assess the spermatotoxicity of MBA. Neither general
toxicity nor spermatotoxicity were observed with either dosing regimen. In a published abstract
on the developmental toxicity of MBA, decreased maternal-weight gain, decreased live-fetus
size, and increased incidence of soft-tissue malformations were reported at gavage doses of >50
EPA/OW/OST/HECD 1-5
-------
Drinking Water Criteria Document for Brominated Acetic Acids
mg/kg/day administered during gestation days (GD) 6-15. No multi-generation reproductive
toxicity, subchronic or chronic studies have been conducted with MBA.
MBA is a dermal irritant, with a lowest-observed-effect concentration (LOEC) of 0.2 M
following a one-hour occluded dermal exposure in rabbits. No data were identified for the
toxicity of MBA following exposure by the inhalation route.
No data were identified on the carcinogenicity of MBA. The genotoxicity data for MBA
have provided mixed results. MBA was mutagenic in Salmonella typhimurium and induced
DNA single-strand breaks in vitro, but did not induce SOS DNA repair (a DNA-repair system
induced in response to DNA damage) in bacteria or micronuclei in a newt-larvae system.
Bromochloroacetic acid
Oral toxicity studies of BCA have identified the kidney, liver, and developmental organs as
potential targets of toxicity. Increased liver weight was observed at the highest drinking water
dose tested (500 mg/kg/day) in B6C3F1 male mouse given BCA in drinking water in a 21-day
study evaluating peroxisomal proliferation and oxidative damage. Marginal increases in liver
weight were induced at the highest dose tested (39 mg/kg/day in drinking water) in rats evaluated
for target-organ toxicity as part of a 26- or 30-day reproductive and developmental screening
assay. In this assay, treatment-related liver histopathological changes (cytoplasmic vacuolization)
were observed beginning at 5 mg/kg/day, and became more prominent at 39 mg/kg/day in rats
given BCA for 30 days. The biological significance of these changes was unclear, as control
males in the parallel 26-day study exhibited the same lesion and there was no dose response.
EPA/OW/OST/HECD 1-6
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Overall, the authors concluded that the high dose of 39 mg/kg/day was sufficient to cause
systemic toxicity, and this value is considered to be a marginal LOAEL.
No standard developmental toxicity studies have been conducted with BCA. The
reproductive toxicity of BCA has been assessed in females and male rats exposed to BCA in
drinking water for 30 to 34 days during the peri-conception period, which included a 12-day pre-
mating period of exposure, exposure during cohabitation on days 13-18, and/or gestational
exposures of varying duration (e.g., GD 1-12, GD 1-21). In females exposed for 30 days, the
NOAEL for reproductive and developmental effects (decreased live fetuses/litter and decreased
total implants/litter) was 19 mg/kg/day, and the NOAEL for maternal toxicity was also 19
mg/kg/day, based on kidney toxicity in the pregnant dam. In females exposed only from GD 6 to
parturition, no dose-dependent increases in either maternal or fetal toxicity were observed. No
effects of BCA on male fertility and sperm quality were noted. In another study (reported in a
published abstract), male Sprague-Dawley rats administered BCA by gavage for 14 days
exhibited impaired sperm motility, abnormalities in sperm morphology, altered spermiation, and
reduced fertility (evaluated by in utero insemination of untreated females). The LOAEL was 8
mg/kg/day, and a NOAEL could not be determined. Adverse effects on sperm quality and male
fertility were also reported (in a published abstract) in male mice exposed to BCA for 14 days. A
decrease in the mean number of litters per male and a decrease in the percent of live litters per
mated female were observed, with a NOAEL of 24 mg/kg/day. No multigeneration reproductive
toxicity study has been conducted for BCA. However, BCA is currently undergoing 90-day
subchronic and 2-year chronic bioassays.
EPA/OW/OST/HECD 1-7
-------
Drinking Water Criteria Document for Brominated Acetic Acids
QSAR modeling predicted a LOEC for skin corrosion of 0.7 M for BCA, indicating a
potential for dermal irritation. No data were identified for the toxicity of BCA following
exposure by the inhalation route.
In a published abstract, BCA was reported to induces liver tumors in mice, but there are
no published reports of a full bioassay with BCA. BCA was mutagenic in Salmonella
typhimurium and induced oxidative DNA damage, as measured by an increase in the DNA
adduct 8-hydroxydeoxyguanisine (8-OH-dG), in the livers of mice given BCA in drinking water.
The data are insufficient to evaluate either the genotoxicity or the potential carcinogenicity of
BCA.
Dibromoacetic acid
In a range-finding study, the reproductive and developmental toxicity of DBA,
administered in deionized drinking water to male and female Sprague-Dawley rats, was
evaluated. Animals were exposed beginning 14 days prior to cohabitation and continuing
through gestation and lactation (a total of 63 to 70 days of treatment). Apparent taste aversion
was associated with an exposure-dependent reduction in water consumption, which was
paralleled by a reduction in food intake at all concentrations, resulting in decreased body weights
in parental animals and postweanling pups at the two highest doses tested. The only observed
adverse reproductive effect was a possible reduction in mating performance in the highest dose
group (1000 ppm), as evidenced by a slight, but nonsignificant, increase in the number of days of
EPA/OW/OST/HECD 1-8
-------
Drinking Water Criteria Document for Brominated Acetic Acids
cohabitation and a decrease in the number of mated pairs (6/10 in the 1000 ppm group versus 9-
10/10 in all other groups). No effects were observed on the incidence of pre- and post-
implantation losses, live litter sizes, and gross external morphology or sex ratios in the pups.
Based on a lack of reproductive effects, the parental and reproductive/developmental NOAEL
for this study was 1000 ppm.
In a recently completed a two-generation reproductive follow up study, male and female
Sprague-Dawley rats were administered DBA in drinking water, continuously from initiation of
exposure of the P generation through weaning of the F2 offspring. Decreased body weight gains
were observed in high-dose P males and females, and at all exposure levels for Fl male and
females, attributed to a general retardation in growth caused by decreased water and food
consumption. Observed delays in sexual maturation in the Fl high-dose group were also
considered to be due to growth retardation. No adverse treatment-related effects were observed
on any reproductive index or developmental parameters, except for statistically significant, dose-
related increase in the number of males exhibiting altered spermatogenesis (i.e., retained Step 19
spermatids in Stage IX and X tubules, and increased or abnormal residual bodies in affected
seminiferous tubules) in the P and Fl groups, and testicular malformation in four males in the
high dose Fl group. The NOAEL was 50 ppm for both P and Fl generations.
A series of rat oral gavage studies on the effects of DBA on spermatogenesis and male
fertility, have been conducted using a number of different experimental protocols, including a
single high-dose study, a 14-day study, and several longer-term studies. The results indicated
that DBA is clearly spermatotoxic, as demonstrated by histopathology indicative of altered
EPA/OW/OST/HECD 1-9
-------
Drinking Water Criteria Document for Brominated Acetic Acids
spermiation. In the 14-day study, the LOAEL was 10 mg/kg/day, based on mild histopathologic
changes (retention of Step 19 spermatids) and aNOAEL could not be determined. In the longer-
term studies, male rats were treated for up to 79 days, with the highest dose group (250
mg/kg/day) being treated for only 42 days due to the onset of overt toxicity. However, fertility
was assessed at various time points up to 213 days by mating treated males with untreated
females. The severity of male reproductive-tract toxicity was both dose- and duration-dependent.
In the group of males given 250 mg/kg/day, fertility was impaired throughout the 6-month
recovery period following cessation of treatment, indicating that damage was structural and likely
permanent. Mild histopathologic changes (retention of Step 19 spermatids) were observed
beginning at 10 mg/kg/day while adverse effects on sperm quality were reported beginning at 50
mg/kg/day. The NO AEL for spermatotoxicity was 2 mg/kg/day. In a published abstract, rats
were exposed in utero from GD 15 through postnatal day (PND) 98 to DBA in drinking water
and reproductive development and adult reproductive function were assessed. Male
reproductive-tract development (as indicated by delayed preputial separation), as well as
spermatogenic and fertility endpoints, were adversely affected at the lowest dose tested (50
mg/kg/day). In another published abstract, exposure of male rabbits to DBA in drinking water,
for a period beginning in utero on GD 15 and continuing to 24 weeks of age, reduced the
conception rates of females artificially inseminated with sperm from these treated males. The
LOAEL was 0.97 mg/kg/day and a NOAEL could not be determined. Full reports of these
studies have not yet been published .
EPA/OW/OST/HECD 1-10
-------
Drinking Water Criteria Document for Brominated Acetic Acids
A recently-published reproductive-toxicity study did not demonstrate significant
spermatotoxic effects in male rats treated with single oral-gavage doses of DBA, as evidenced by
the absence of treatment-related effects on sperm motility, morphology, and membrane
permeability at doses > 600 mg/kg, although mild testes histopathology was reported. The
LOAEL histopathology was 600 mg/kg and a NOAEL could not be determined. The lack of
reported spermatotoxicity in this study is in contrast to an earlier single-dose gavage study, in
which significantly adverse effects of DBA on sperm count, motility, and morphology at doses
> 1250 mg/kg were observed in a different strain of rats. Variation among studies in DBA
spermatotoxic effects may have been associated with differences in rat strain, experimental
design, end-point measurements, and other study variables.
DBA administered to Holtzman rats via oral gavage on GD 1-8 did not impair female
fertility although a 170 % increase in serum 17p-estradiol was observed at 250, but not 500,
mg/kg/day. In two published abstracts, DBA was reported to induce reproductive and
developmental toxicity in pregnant CD-I mice administered DBA on GD 6-15. Delayed
parturition was observed at doses > 24 mg/kg/day, but the toxicological significance of this effect
is unclear. Increased postnatal mortality, decreased pup weight, and tail defects were observed at
> 610 mg/kg/day. In the second abstract by the same authors, the ability of DBA to induce fetal
malformations was examined in pregnant CD-I mice administered DBA by oral gavage on GD
6-15. The NOAEL for renal malformations (hydronephrosis) was 50 mg/kg/day. In contrast, no
treatment-related effects on litter viability, postnatal mortality, gross malformations and a wide
array of other developmental end points were observed in a recent two-generation drinking water
EPA/OW/OST/HECD 1-11
-------
Drinking Water Criteria Document for Brominated Acetic Acids
reproductive/developmental toxicity study, conducted according to currently-accepted standard
test guidelines. Differences in findings between the two-generation study and those reported in
published abstracts may have been due to differences in internal doses associated with gavage
versus drinking water DBA administration, species differences in susceptibility to DBA toxicity,
the lower mean doses tested in the two-generation study, and/or other factors.
Subsequent studies found diminished promordial follicle populations in rabbits and
disruptions in rat estrous cyclicity and ovarian follicular steroid release in vitro.
Among the three brominated acetic acids, MBA, BCA, and DBA, unequivocal evidence
of adverse effects on spermatogenesis in rats is available for DBA. Although an adverse effect
on the mating performance in male rats treated with high doses of DBA was reported ,
reproductive function was unaffected by altered spermiation in the two-generation
reproductive/developmental toxicity study. In the single study identified in the literature, MBA
had no effect on spermatogenesis under treatment conditions similar to those that yielded positive
indications of spermatotoxicity for DBA. BCA was also found to perturb spermiogenesis and
fertility in male rats. Although a decrease in total implants per litter and live fetuses per litter
was reported for BCA in this study, both males and females were exposed to treated drinking
water, and thus these reproductive effects might have been due to female exposure. However,
adverse effects on male fertility have been reported in two other studies in which only males
were treated.
In addition to reproductive and developmental endpoints, the liver, immunotoxicity, and
neurotoxicity of DBA have been evaluated. In male mice treated with DBA in drinking water
EPA/OW/OST/HECD 1-12
-------
Drinking Water Criteria Document for Brominated Acetic Acids
for 21 days, increased liver weight was observed beginning at 125 mg/kg/day and was
accompanied by oxidative stress, as indicated by increases in hepatic cyanide-insensitive Acyl-
CoA activity and 8-OHdG levels. The NOAEL was 125 mg/kg/day.
In an immunotoxicity assay , female mice were given DBA in their drinking water for 28
days. Four independent studies were conducted and different endpoints were examined in each
study. Studies 1-3 investigated selected immunologic parameters, body-weight changes, and
selected organ weights; in Study 4, body weight, organ weights, hematology, and gross
pathology were examined. Overall, the results of Study 1 demonstrated an increase in several
measures of cellular immunity, including an increase in the total number of spleen cells and an
increase in the percent of spleen cells as macrophages, with statistically significant effects
generally occurring at doses >73 mg/kg/day. However, the toxicological significance of these
findings was unclear. In Study 2, the spleen IgM antibody-forming cell response was
significantly decreased at > 70 mg/kg/day, but no change was observed in serum IgM titer (a more
generalized measure of immune function, encompassing immune activity in the bone marrow and
lymph nodes as well as the spleen). No change in macrophage activation was observed when
tested for in Study 3. In Study 4, the authors reported decreased body weight; decreased thymus-
gland weight; increased liver, kidney, and spleen weights; and increased reticulocyte counts. For
this group of studies, spleen IgM antibody-forming cell response was chosen as the critical effect
because it represented a clear decrease in spleen immune-system function. The NOAEL for this
end point was 38 mg/kg/day. Changes in body weight, spleen and thymus weights, and
reticulocyte counts occurred at the same or higher doses than the critical effect; changes in liver
EPA/OW/OST/HECD 1-13
-------
Drinking Water Criteria Document for Brominated Acetic Acids
and kidney weights were observed at lower doses, but were not considered to be lexicologically
significant in the absence of supporting clinical chemistry and histopathology data.
In a neurotoxicity study published as an abstract, male and female adolescent F344 rats
were exposed to DBA in drinking water for 6 months, and a neurobehavioral test battery was
administered to all animals at 1,2, 4, and 6 months. Dose-dependent neuromuscular effects
included mild gait abnormalities, decreased forelimb and hindlimb strength, hypotonia, decreased
sensorimotor responsiveness (as measured by responses to a tail pinch and auditory click),
decreased motor activity, and a chest-clasping response that was only observed in high-dose
females. Sensorimotor responsiveness did not progress with continued exposure.
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. Histological evidence of neuropathology was observed in the mid- and high-dose, and
was not evaluated in the low-dose group. The LOAEL for this study was 20 mg/kg/day, and a
NOAEL could not be determined.
No long-term systemic toxicity studies for any exposure route were identified in the peer-
reviewed literature. However, DBA is currently undergoing 90-day subchronic and 2-year
chronic bioassays.
Although the genotoxicity database is limited, DBA is mutagenic in Salmonella
typhimurium assays and tests for DNA-damage repair, and has increased the DNA adduct, 8-
OHdG, in hepatic DNA of mice exposed via drinking water.. On the other hand, no induction of
EPA/OW/OST/HECD 1-14
-------
Drinking Water Criteria Document for Brominated Acetic Acids
micronuclei was reported in a newt-larvae system, suggesting that DBA was not clastogenic in
the newt test system. The clastogenicity of DBA has not been reported in other assays using a
standard test protocol, although DBA has been reported to be co-clastogenic. A standard mouse-
micronucleus assay has not been conducted. These data support the conclusion that DBA is
mutagenic and genotoxic, although the nature of the DNA damage induced by DBA remains
unclear. In published abstracts, DBA was reported to induce aberrant crypt foci in the colon of
rats and liver tumors in mice. However, complete reports of these bioassays have not been
published, limiting the utility of these data in assessing the potential carcinogenicity of DBA.
No studies were identified that directly evaluated human-health effects of exposure to
MBA, BCA, or DBA via any route
MBA is more toxic than DBA in acute toxicity studies. 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 studies using purified human
enzymes and by evidence for DNA alkylation, but a direct relationship between these reactions
with cellular macromolecules in vivo and the observed toxic effects of MBA has not yet been
established.
DBA and BCA have been associated with liver, kidney, and reproductive and
developmental toxicity in a variety of toxicity studies. Potential mechanisms for the induction
of adverse liver effects include perturbations of carbohydrate homeostasis or toxicity due to the
formation of reactive metabolites from haloacetic acid or tyrosine- metabolism pathways. The
kidney may also be a target for brominated acetic acids, possibly reflecting direct toxicity
EPA/OW/OST/HECD 1-15
-------
Drinking Water Criteria Document for Brominated Acetic Acids
associated with the formation of reactive metabolites, or toxicity secondary to oxalate formation
although it appears unlikely that sufficient oxalate is formed during brominated acetic acid
metabolism to adversely affect the kidney.
Differences in the spermatotoxicity of these brominated compounds are also apparent.
DBA, but not MBA, induced spermatotoxicity at gavage doses that also induced overt toxicity in
an acute toxicity study, but effects on reproductive function were not observed in a two-
generation reproductive toxicity drinking-water study in which animals were administered
spermatotoxic doses of DBA. BCA was reported to affect spermiogenesis and reduce fertility in
male rats. A published abstract, also, reported impaired sperm quality and spermiation, and
reduced male fertility as assessed by in utero insemination of untreated females with the sperm of
males treated with BCA for 14 days. In another published abstract, BCA was also reported to
decrease male fertility in mice. The weight-of-evidence suggests that both DBA and BCA are
male reproductive-tract toxicants. One hypothesized target for observed spermatotoxicity is the
Sertoli cells. 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 and/or its reactive metabolites have been suggested as possible
mechanisms. Another potential mechanism of spermatotoxicity is haloacetic acid-mediated
disruption of the early stages of steroidogenesis, possibly by interfering with the steroidogenic
acute regulatory protein (StAR)-mediated transport of cholesterol within the mitochondrial
membrane and thereby affecting the synthesis of pregnenolone, the precursor of progesterone.
Other studies suggest that brominated acetic acids may interfere with spermatogenesis by altering
EPA/OW/OST/HECD 1-16
-------
Drinking Water Criteria Document for Brominated Acetic Acids
sperm proteins (most notably SP22) that play an important role in the fertilization process,
possibly by regulating the androgen receptor. It has also been suggested that haloacetic acids
acting on SP22 and other sperm proteins may indirectly compromise androgen-dependent
maintenance of spermatogenesis.
All three brominated acetic acids have been reported to induce developmental effects.
Although the spectrum of developmental endpoints affected by in vivo treatment does not
implicate a common mode of action, the results of whole-embryo culture (WEC) testing have
suggested the mechanisms of developmental toxicity among haloacetic acids are similar. A
QSAR model using WEC data was able to adequately describe the rank-order potency of a series
of haloacetic acids. The results of a WEC study testing mixtures of haloacetic acids were
consistent with the QSAR model predictions of dose-additivity. Brominated acetic acids also
induced dysmorphogenesis in WEC at doses lower than the doses of known metabolites,
suggesting that either the parent compound or other unidentified metabolites are responsible for
these developmental effects. Apoptosis induction has been proposed as having a role in the
mechanism of onset of in vivo developmental toxicity based on the results of WEC testing, but
this hypothesis has not been confirmed in vivo
No data are available for identifying susceptible populations. In addition, no data on age-
dependent changes in the expression of genes involved in brominated acetic acid were located.
Based on the results of some in vivo developmental toxicity studies reported in abstracts, DBA,
but not MBA or BCA, induced fetal toxicity at lower doses than those associated with maternal
effects, suggesting that, at least for DBA, the fetus might be more susceptible than the adult.
EPA/OW/OST/HECD 1-17
-------
Drinking Water Criteria Document for Brominated Acetic Acids
However, these preliminary developmental studies found fetal and maternal effects only at doses
well above those causing effects on spermiation, indicating that protection against the latter
effect would also provide adequate protection to children and fetuses. Additionally, DBA was
administered by gavage in these studies and observed fetal toxicity might have been route-
dependent, because increased fetal susceptibility is not supported by the results of the two-
generation drinking water reproductive toxicity study with DBA.
There are also inadequate data on potentially susceptible subpopulations based on genetic
variability. Human polymorphisms in GST-Zeta, the enzyme that metabolizes DBA and BCA to
glyoxylate, have been characterized by several investigators. However, in the absence of data on
whether the parent compound or a metabolite is the active moiety, the functional consequences of
this polymorphism are not clear. Individuals having underlying defects in glycogen storage may
be susceptible to liver effects induced by brominated acetic acids, and individuals lacking certain
enzymes for 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 might also represent
a source of increased susceptibility. Another potentially susceptible population to DBA are
individuals with hereditary tryosinemia II (a disease involving a deficit in tyrosine metabolism);
its chlorinated analog, DCA, has been shown to alter tyrosine metabolism as a consequence of
its inhibitory effects on GST-Zeta. None of the possibilities has been examined directly in in
vivo studies, and potentially susceptible populations have not been identified.
EPA/OW/OST/HECD 1-18
-------
Drinking Water Criteria Document for Brominated Acetic Acids
No suitable studies for the derivation of drinking-water health advisories (HA) were
identified in the literature for MBA, BCA, or DBA. Neither subchronic nor chronic toxicity
studies have been conducted with any of these compounds, although both subchronic and chronic
toxicity testing of BCA and DBA is planned or in progress. A number of additional studies are
currently ongoing.
There are no human epidemiology studies or full animal-cancer bioassays for MBA,
BCA, or DBA, although both BCA and DBA are slated for full testing (NTP, 2000b; NTP,
2000c). Under the 1999 Draft Guidelines for Cancer Risk Assessment, the data are inadequate
for an assessment of human carcinogenic potential of MBA and BCA.
There is concern for the potential carcinogenicity of DBA based on preliminary findings
reported in published abstracts,, and analogy to DCA, a known high-dose rodent-liver
carcinogen. However, insufficient data are available to assess DBA carcinogenic hazard.
Under the 1999 Proposed Guidelines for Cancer Risk Assessment, the data are inadequate for an
assessment of human carcinogenic potential of DBA.
EPA/OW/OST/HECD 1-19
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Chapter II. Physical and Chemical Properties
Three brominated acetic acids have been selected for consideration in this document.
These are monobromoacetic acid (MBA), dibromoacetic acid (DBA), and bromochloroacetic acid
(BCA). Brominated acetic acids are formed during ozonation or chlorination of water containing
bromide ions (Jacangelo et al., 1989; Pourmoghaddas et al., 1993) and organic matter, primarily
humic and fulvic acids. Formation of chlorinated acetic acids is higher in the presence of humic
acid than in the presence of fulvic acid (WHO, 2000), suggesting that a similar relationship may
hold for brominated acetic acids. Bromide ions occur naturally in surface and ground water.
However, seasonal fluctuations in bromide-ion levels can occur. In addition, bromide-ion levels
can increase due to saltwater intrusion resulting from drought conditions, or due to pollution
(WHO, 2000).
The bromide-ion concentration is an important determinant of the spectrum of haloacetic
acids formed from the reaction of disinfectants with organic material. In the presence of sufficient
concentrations of bromide ion, the formation of brominated compounds maybe favored over
formation of chlorinated compounds (Pourmoghaddas et al., 1993). Brominated acetic-acid
concentrations in drinking water are typically in the order of BCA>DBA>MBA (Jacangelo et al.,
1989; Krasner et al., 1989; Boorman et al., 1999; U.S. EPA, 2000a). The occurrence of these
compounds is discussed more fully in Chapter IV.
EPA/OW/OST/HECD II-l
-------
Drinking Water Criteria Document for Brominated Acetic Acids
The chemical reactions resulting in the formation of brominated acetic acids have been
described in detail for chlorinated water (WHO, 2000). When chlorine gas is added to water (e.g.,
as a disinfectant), it hydrolyzes almost immediately to form hypochlorous acid (HOC1):
ci2 + H2o -> HOCI + H++ cr
Hypochlorous acid can then dissociate into the hydrogen ion and hypochlorite in a reversible
reaction:
HOCI » H+ + ocr
In the presence of bromide ion, hypobromous acid (HOBr) is formed from hypochlorous acid in
the following irreversible reaction:
HOCI + Br -> HOBr + Cl'
Similar reactions occur to form HOBr from other drinking-water disinfectants, and the resulting
HOBr reacts with organic material to form brominated acetic acids, as shown below (Jacangelo et
al., 1989; Pourmoghaddas et al., 1993). In summary, brominated acetic acids are formed from the
following reactions:
Bromide ion + Ozone or HOCI ->• HOBr
HOBr + Organic acid (e.g., humic acid) ->• Brominated acetic acid
Figure II-1 shows the structure of monobromoacetic acid (MBA), bromochloroacetic acid
(BCA), and dibromoacetic acid (DBA), and Table II-1 summarizes key physical and chemical
properties of these compounds. The data contained in Table II-1 apply to the pure form of the
selected chemicals. These chemicals exist in the environment in a dissolved form.
EPA/OW/OST/HECD II-2
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Various synonyms exist for the selected chemicals addressed in the document. Some
common synonyms for MBA are bromoacetic acid, 2-bromoacetic acid, a-bromoacetic acid,
bromoacetate ion, bromoethanoic acid, carboxymethyl bromide, and acetic acid, bromo-.
Synonyms for DBA are dibromoacetate and acetic acid, dibromo-. Likewise for BCA, the known
synonyms are acetic acid, bromochloro-, and chlorobromoacetic acid.
Figure II-l. The Chemical Structures of MBA, BCA, and DBA
MBA
H 0
Br C C OH
H
BCA
Cl 0
DBA
Br 0
Br C C OH
H
Br C C OH
H
EPA/OW/OST/HECD
II-3
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table II-l. Physical and Chemical Properties of Brominated Acetic Acids."
Property
Chemical Abstracts
Registry Services No.
Formula
Molecular weight
Appearance
Density (g/mL)
Melting point (°C)
Boiling point (°C)
Solubility
Water
Alcohol
Log;?
LogŁ>e
Monobromo-
acetic acid
(MBA)
79-08-3
BrCH2COOH
138.95
hygroscopic
crystal
1.93
49-51
208
miscible
miscible
0.41°
nd
Dibromo-
acetic acid
(DBA)
631-64-1
Br2CHCOOH
217.84
hygroscopic crystal
-
49
218
very soluble
very soluble
1.22d
-1.69
Bromochloro-
acetic acid
(BCA)
5589-96-8
BrClCHCOOH
173.39
hygroscopic crystal
1.98
27.5-31.5
210-215
ndb
nd
1.08d
-1.77
a. Adapted from the CRC Handbook of Chemistry and Physics (1999), The Merck Index (1996), and
Sigma-Aldrich (2000).
b. nd: no data
c. Log/? is the value derived experimentally as presented in Hansch et al. (1995).
d. Log/? is the calculated octanol - water partition coefficient in the un-ionized form as
presented in Schultz et al. (1999).
e. Log D is the distribution coefficient between n-octanol and buffer at pH 7.4, as presented in
Schultz etal. (1999).
PKa values were not identified for these compounds.
EPA/OW/OST/HECD
II-4
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Chapter III. Toxicokinetics
Although quantitative toxicokinetic data for brominated acetic acids are limited, the
available information suggests that DBA and BCA are well absorbed, almost completely
metabolized, and minimally excreted in the urine and feces. Most of these quantitative data come
from a single, preliminary study conducted by Schultz et al. (1999).
A. Absorption
Monobromoacetic acid
No studies investigating the quantitative parameters of MBA absorption were identified
for any route of exposure. Adverse target-organ effects observed in short-term toxicity studies
(described in detail in Section V) show that MBA is absorbed following exposure by the oral
route; however, the kinetics of absorption are not currently known.
Bromochloroacetic acid
BCA is systemically absorbed following oral dosing (Table ni-1). Schultz et al. (1999)
administered a single oral gavage or intravenous (IV) doses of 500 |_imol/kg (87 mg/kg ) to male
F344 rats. Following dosing, BCA venous blood concentrations were measured at 0, 5, 10, 20,
EPA/OW/OST/HECD III-l
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table III-l. Toxicokinetic Data for BCA and DBA in F344 Rats"
Parameters determined following
IV dosing with 500 |j,mol/kg
Area under blood concentration-time curve
AUC ftiM-h)
Amount excreted in urine in 24 h (% Dose)
Steady-state apparent volume of distribution
(mL/kg)
Total body clearance (mL/h-kg)
Renal clearance (mL/h-kg)
Mean residence time (h)
Elimination half-life (h)
Unbound fraction in plasma (fj
Blood/plasma concentration ratio
Parameters determined following oral
dosing with 500 |j,mol/kg
Area under blood concentration-time curve
AUC OiM-h)
Mean residence time (h)
Time to peak blood concentration (h)
Mean absorption time (h)c
Oral Bioavailability (%)d
BCA
(87 mg/kg)
576±286b
2.16 ±1.07
881 ±373
1037 ±453
36.9 ±20.8
0.92 ±0.41
3.93 ±1.50
0.93 ±0.07
0.98 ±0.12
BCA
(87 mg/kg)
270 ± 38
2.12±0.81
1.5
1.20
>47e
DBA
(109 mg/kg)
1120 ±362
2.67 ±1.09
400 ±112
491 ±116
12.9 ±4.0
0.93 ±0.50
0.72 ±0.12
0.89 ±0.05
0.91 ±0.05
DBA
(109 mg/kg)
333 ±70
2.10 ±0.70
1
1.17
>30e
Notes:
a. Adapted from Schultz et al, 1999
b. Mean ± standard deviation
c. Calculated as the difference between the mean residence time following IV versus oral dosing.
EPA/OW/OST/HECD
III-2
-------
Drinking Water Criteria Document for Brominated Acetic Acids
d. The ratio of the blood concentration AUC for oral versus IV dosing x 100%.
e. Although the study authors reported estimated values of 47% and 30% for BCA and DBA, respectively, the oral
bioavailability is likely to be underestimated, based on first-pass metabolism considerations associated with oral
and not IV dosing.
30, 60 and 90 minutes, and 3, 4, 6, 8 and 12 hours; concentrations of BCA in the urine and feces
were measured in samples collected for 24 hours after dosing. The oral bioavailability - the ratio
of the averaged values for the area under the curve for the oral and i.v. doses - was estimated as
47% for BCA. However the oral bioavailability of BCA might be higher than indicated due to
more extensive first pass metabolism via this route than via the intraveneous route.
To measure the absorption rate, the time-to-peak blood concentration was determined.
The peak concentration of BCA was observed 1.5 hours following oral dosing. Rapid absorption
of BCA was confirmed by the mean absorption time, which was determined by measuring the
difference in the mean residence time in blood between oral and IV dosing. The mean absorption
time was reported as 1.2 hours. These data show that BCA is readily absorbed following a single
bolus dose. However, no quantitative data are available to assess whether the degree of
absorption would be different under ingestion conditions more closely resembling human
exposure conditions (i.e., temporally dispersed and at much lower doses).
No studies on the absorption of BCA were identified following exposure by the inhalation
or dermal routes.
Dibromoacetic acid
EPA/OW/OST/HECD III-3
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Shultz et al. (1999) also studied the systemic absorption of DBA following oral dosing, as
described above for BCA (Table III-l). Male F344 rats were given single oral gavage or IV doses
of 500 |-imol/kg (109 mg/kg) DBA. Following dosing, the DBA venous-blood concentrations
were measured at 0, 5, 10, 20, 30, 60 and 90 minutes, and 4, 6, and 8 hours; concentrations of
DBA in the urine and feces were measured in samples collected 24 hours after dosing. The oral
bioavailability of DBA was estimated to be 30%. However, this value is likely to be an
underestimate because it probably reflects first pass-metabolism following oral dosing and, thus,
underestimates the degree of oral absorption. The peak blood concentration was observed 1 hour
following oral dosing and the mean absorption time (the difference in the mean residence time in
blood following dosing via oral and IV routes) was 1.17 hours, indicating rapid absorption.
However, no quantitative data were provided on the degree of absorption under conditions of low-
dose repeated oral dosing, a regimen which would more closely resemble human-exposure
conditions.
Blood-level measurements for brominated acetic acids following oral dosing have been
reported in other toxicity studies, and provide some additional information on DBA absorption.
As part of a study on the effects of DBA on pubertal development and adult reproductive function,
Klinefelter et al. (2000, abstract only) reported blood-serum and milk concentrations in Sprague-
Dawley rats (3 litters/dose) administered drinking water containing 0, 400, 600, or 800 ppm DBA
from gestation day (GD) 15 through postnatal day (PND) 98. Estimated doses resulting from
these treatments were 0, 50, 75, and 100 mg/kg/day, respectively (personal communication).
DBA levels were also assayed in dams' milk and blood serum, and in the serum of suckling males
on PND 20 (personal communication). Only data for the 800 ppm treatment group were
presented in the abstract. In this group, DBA concentrations were 5.2-14.1 |j,g/mL in dams' milk,
3.0- 6.9 i-ig/mL in dams' serum, and 0.01-0.24 |j,g/mL in the serum of male offspring. The limited
amount of information in the abstract precluded quantification of the degree of absorption.
EPA/OW/OST/HECD III-4
-------
Drinking Water Criteria Document for Brominated Acetic Acids
However, the data demonstrate that DBA was absorbed following drinking-water exposure. In
contrast, no detectable levels of DBA were observed in the plasma of female B6C3F1 mice
following 28 days of exposure to DBA-treated drinking water at concentrations up to 2000 mg/L,
corresponding to estimated doses of 229 to 285 mg/kg/day depending on the sub-study (NTP,
1999). The reasons for the differences in findings between the Klinefelter (2000) and NTP (1999)
studies are unclear, and may have been due to differences in analytical methods. No details on the
post-dosing sampling schedules for blood and/or milk were presented in either study. The
absence of measurable DBA in plasma in the NTP (1999) study might reflect extensive
metabolism and rapid excretion, rather than limited absorption. Alternately, species differences
and/or differences in hormonal status might have affected the kinetics of DBA absorption,
distribution, metabolism and/or excretion in these studies. Christian et al. (1999) reported the
detection of DBA in the plasma of male and female rats exposed to DBA in drinking water
concentrations for 16 hours, and following 14 days of treatment, also demonstrating that
gastrointestinal absorption occurred.
No studies on the absorption of DBA were identified following exposure by the inhalation
or dermal routes.
B. Distribution
Monobromoacetic acid
No studies of MBA tissue distribution following dosing by any route were identified.
Bromochloroacetic acid
EPA/OW/OST/HECD III-5
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Schultz et al. (1999) administered male F344 rats single oral gavage or IV doses of 500
|j,mol/kg (87 mg/kg) BCA. BCA venous-blood concentrations were measured at 0, 5, 10, 20, 30,
60 and 90 minutes, and at 3, 4, 6, 8 and 12 hours post-dosing. No concentrations in tissues other
than blood were assessed. Following intravenous dosing, BCA did not appear to bind
significantly to plasma proteins or accumulate in blood cells. The unbound fraction in plasma was
0.93, and the ratio between plasma and blood concentrations was close to unity (i.e., 0.98). Tissue
concentrations were not measured directly. However, following intravenous dosing, the apparent
volume of distribution (881 mL/kg) was similar to the total body-water volume for rats
(approximately 660 mL/kg) (Reinoso et al., 1997). This similarity suggested to the study authors
that BCA distributed uniformly outside the vascular system and was unlikely to sequester
significantly in peripheral tissues. The octanol-buffer partition coefficient (Log D), considered to
be a reasonable predictor of lipophilicity (Schultz et al., 1999) was reported to be -1.77 at pH of
7.4. This low value also suggested that at physiological pH values, BCA has little propensity to
accumulate in fat tissue. However, in the absence of direct tissue measurements, the distribution
of BCA cannot be ascertained. Plasma binding, plasma:blood concentrations, and the apparent
volume of distribution (Vss) were only measured following intravenous dosing and, thus, the
effects of first-pass metabolism are not known. Further, the high dose employed in this study
might have resulted in metabolic saturation which could have led to a wider distribution of BCA
than would have occurred at lower doses where metabolism was not saturated.
Short-term animal studies suggest that oral administration of high doses of BCA results in
liver, reproductive, and developmental toxicity (NTP, 1998, Parrish et al., 1996; see Chapter 5 for
more details), indicating that BCA does distribute to the liver, the reproductive organs, and the
fetus under the conditions of these studies. The low protein-binding capacity of BCA suggests
that the potential for distribution across the placenta to the fetus may be significant at high
maternal doses administered during pregnancy.
EPA/OW/OST/HECD III-6
-------
Drinking Water Criteria Document for Brominated Acetic Acids
No data on tissue distribution of BCA following exposure by the inhalation or dermal
routes were identified.
Dibromoacetic acid
The systemic distribution of DBA has also been studied by Schultz et al. (1999). Male
F344 rats were given single oral gavage or IV doses of 500 |_imol/kg (109 mg/kg) DBA. DBA
venous-blood concentrations were measured at 0, 5, 10, 20, 30, 60 and 90 minutes, and at 4, 6,
and 8 hours post-dosing. No tissue concentrations other than blood were measured. Following
intravenous dosing, DBA did not bind significantly to plasma proteins or accumulate in blood
cells. The unbound fraction in plasma was 0.89 for DBA, and the ratio between plasma and blood
levels of DBA was close to unity (i.e., 0.91). Tissue concentrations were not measured directly,
but the apparent volume of distribution and the total body-water volume in rats were similar,
suggesting to the study authors that DBA was widely and uniformly distributed outside the
vascular system and was unlikely to sequester significantly in peripheral tissues. The octanol-
buffer partition coefficient (Log D), considered to be a reasonable predictor of lipophilicity
(Schultz et al., 1999), was reported to be -1.69 at pH of 7.4. This low value suggested that at
physiological pH values, DBA has little propensity to accumulate in fat tissue. However, in the
absence of direct tissue measurements, the distribution of DBA is not known. Plasma binding,
plasma:blood concentrations, and the apparent volume of distribution (Vss) were only measured
following intravenous dosing and, thus, the effects of first-pass metabolism cannot be ascertained.
Further, the high dose employed in this study might have resulted in metabolic saturation, which
could have led to a wider distribution of DBA than would have occurred at lower doses where
metabolism was not saturated.
EPA/OW/OST/HECD III-7
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Short-term animal studies suggest that oral administration of DBA results in toxicity to the
liver, kidney, spleen, and male reproductive system.(Linder et al., 1994a; Linder et al., 1994b;
Linder et al., 1995; Parrish et al., 1996; Linder et al., 1997; Cummings and Hedge, 1998; Vetter et
al., 1998; NTP, 1999; more details in Chapter 5), indicating that DBA does distribute to these
organs under the dosing conditions of these studies. Further, the study by Klinefelter et al. (2000,
abstract), described in the previous section on DBA absorption, showed that the milk of lactating
Sprague-Dawley females treated with 800 ppm (100 mg/kg/day) of DBA in drinking water
contained elevated levels of DBA relative to the females' blood serum (5.2-14.1 |j,g/mL in milk,
3.0- 6.9 |j,g/mL in serum). Thus, lactational distribution to nursing pups may be significant at high
maternal doses.
As part of a series of range-finding reproductive and developmental toxicity studies,
Christian et al. (2001) evaluated the distribution of DBA in adult Sprague-Dawley rats, and in
pregnant females and their fetuses. Male and female rats (10/sex/group) were given DBA in
deionized drinking water at concentrations of 0, 125, 250, 500 or 1000 ppm, beginning 14 days
prior to cohabitation and continuing through gestation and lactation (63-70 days of treatment).
The average daily doses (based on measured water consumption and body weights) varied,
depending on the phase of reproduction. For parental males throughout the study (SD 1-70),
equivalent mean daily doses were 10.2, 20.4, 35.7, and 66.1 mg/kg/day, respectively. For females
on SD 1-15, equivalent mean daily doses were 13.3, 26.2, 41.8 and 60.2 mg/kg/day, respectively;
and 14.8, 30.3, 48.5 and 81.6 mg/kg/day, respectively, on gestation day (GD) 0-21. During
lactation (LD 1-29), the estimated doses were -was 43.5, 86.6, 150.7 and 211.7 mg/kg/day for the
0, 125, 250, 500, and 1000 ppm groups, respectively; however, these doses included consumption
of water by the pups and thus overestimated the mean daily intake for lactating females.
EPA/OW/OST/HECD III-8
-------
Drinking Water Criteria Document for Brominated Acetic Acids
An additional 6 male and 17 female rats/group were used for collecting bioanalytical
samples. Blood plasma levels of DBA in parental male and female rats were taken on SD 1 and
14, in mated females on GD 20 and LD 15, and in weanling male and female rats on LD 29
(which was also postweaning day 1). Tissue levels of DBA in placenta and amniotic fluid were
assessed on GD 21 and in milk on LD 15. During the designated collection days, three plasma
samples were collected from male and female rats at approximately 8-hour intervals in order to
assess the potential differences in tissue concentrations associated with diurnal rhythms of water
consumption by animals.
In analyzing biodisposition samples, numerous values were measured at concentrations
below the limit of detection (LOD) of the methodology used. These values were quantified but
were not considered by the study authors to be reliable. In blood plasma, DBA was detected in
male rats at 125 ppm and reliably quantified at >250 ppm after 24 hours of continual access to
DBA in drinking water. Blood plasma levels in males exposed to 125 ppm did not exceed the
level of detection even after 14 days of exposure. The overall pattern of DBA detection in plasma
in female rats was similar but more variable; however, quantifiable DBA was increased in a dose-
dependent fashion at all exposure concentrations on SD 14, and was observed at concentrations of
4.4-6.7 |-ig/g at the end of the dark period. The higher plasma levels at this time were attributed to
nocturnal drinking habits and not to accumulation of DBA in plasma. During gestation, the
calculated daily dose of DBA was increased, and was reflected in increased plasma concentrations
at all DBA exposure levels on GD 20 (ranging from 4.2-18.0 |-ig/g), but not in an exposure-
dependent manner. During lactation, DBA could be quantified in the maternal plasma at all doses
when measured on LD 15. Measurements ranged from 1.9 to 26.2 |j,g/g and varied, depending on
time of sample; a dose-response was only observed at one time point. Analysis of milk samples
collected from three lactating dams/group during the light period on LD 15 indicated that DBA
was not detected in milk, in contrast to the findings of Klinefelter et al (2000, abstract).
EPA/OW/OST/HECD III-9
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Pooled fetal blood, amniotic fluid, and placentas were collected on GD 21 from three
litters/exposure group and analyzed for DBA. In the placenta, DBA was detected at all exposure
levels but could only be reliably quantified in the 1000-ppm group (8.1 |_ig/g at a mean maternal
daily dose of 8.16 mg/kg/day). In fetal plasma, DBA was detected in a dose-dependent manner in
all groups; however, levels in the 125- and 250-ppm groups were below the limit of detection and
reliable quantitation was only possible in the 500 and 1000 ppm groups (2.4 and 9.2 |j,g/g at
maternal doses of 48.5 and 81.6 mg/kg/day, respectively). In contrast, quantifiable amounts of
DBA were noted in the amniotic fluid (3.9, 5.3, 3.0, and 5.8 jig/g at 14.8, 30.3, 48.5, and 81.6
mg/kg/day, respectively) and levels were comparable to those observed in the plasma of maternal
rats on GD 20 at the start of the dark period, demonstrated a dose-dependent distribution of DBA
in amniotic fluid (3.9-5.8 |-ig/g) on GD 21 at DBA drinking water concentrations of 125-1000 ppm
(estimated mean daily doses of 14.8-81.6 mg/kg/day). Equivalent concentrations of serum plasma
DBA in pregnant females on GD 20, taken at the same time during the dark cycle as the amniotic
fluid samples, ranged from 3.6 to 6.4 |-ig/g. These results demonstrate that at high drinking water
concentrations, DBA can cross the placenta and distribute to fetal tissue.
No data on tissue distribution of DBA following exposure by the inhalation or dermal
routes were identified.
C. Metabolism
Monobromoacetic acid
No studies were identified that described the metabolism of MBA following exposure by
any route.
EPA/OW/OST/HECD III-10
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Bromochloroacetic acid
Most of the data on the metabolism of BCA are indirect or based on analogy to the
chlorinated acetic acids. These indirect data suggest that dihalogenated acetic acids are rapidly
metabolized. Schultz et al. (1999) did not directly measure the metabolism of BCA. However,
comparisons of renal and blood clearance following IV administration suggested that metabolism
was likely to be the major contributor to BCA removal from the blood. Only 2% of BCA blood
clearance was accounted for by renal clearance and excretion in the feces was negligible. Thus,
non-renal clearance (e.g., through metabolism) accounted for most of the removal of the parent
compound from the blood. It should be noted that these data were obtained following IV
administration of the test compound and, thus, the degree of metabolism in different tissues could
not be determined. The relatively low blood concentrations of BCA following oral dosing,
compared to the blood concentrations following IV dosing, suggest that the liver may be an
important site for first-pass metabolism.
Shultz et al. (1999) suggested that the similarities between the toxicokinetics of BCA and
dichloroacetic acid (DCA), coupled with limited in vitro metabolism data (Schultz, et al., 1998;
Tong et al., 1998a), support the hypothesis that metabolism of BCA is similar to that of DCA.
However, while the metabolic pathways for BCA and DCA might be similar, the rate of
metabolism seems to be greater for BCA than DCA based on area-under-the-curve plasma data.
In a review of DCA metabolism, Stacpoole et al. (1998) described several potential
mechanisms for the dehalogenation of DCA, but evidence for any single pathway was limited.
Recent evidence suggests that DCA is metabolized to glyoxylic acid through a glutathione-S-
transferase-dependent mechanism by a novel GST isozyme, GST-Zeta (Tong et al., 1998b); in
EPA/OW/OST/HECD III-11
-------
Drinking Water Criteria Document for Brominated Acetic Acids
vitro studies have demonstrated that this enzyme can also catalyze the metabolism of BCA to
glyoxylic acid (Tong et al., 1998a). In comparative studies using DCA, Tong et al. (1998a) found
that GST-Zeta is expressed in mouse, rat and human-liver cytosol. For DCA, the relative
biotransformation rates were mouse>rat>human. Although glutathione was a required cofactor, it
was not consumed or oxidized during the conversion of DCA to glyoxylate.
EPA/OW/OST/HECD 111-12
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Glycolate
OH
H
H
Glycine
H
H
0
N —
H
CO,
I
H
Dibromoaceticacid
Br
OH
AGT
Glyoxylic acid
O LDH Q^ ^O
GS^Zeta
OH
C
H
HAOX
o.
OH OH
Oxalic acid
Br
OH
OH
OH
O.
,O
Br
OH
Figure III-l. Proposed Metabolism of DBA3
AGT = alanine:glyoxylate aminotransferase
GR = glyoxylate reductase
GSTZ = glutathione-S-transferase-Zeta
HAOX = glycolate oxidase (2-hydroxyacid oxidase)
LDH = lactate dehydrogenase
OGC = 2-oxoglutarate:gloxylate carboligase
a. Adapted largely from Kennedy et al. (1993), Stacpoole et al. (1998)
EPA/OW/OST/HECD
III-13
-------
Drinking Water Criteria Document for Brominated Acetic Acids
In addition to glyoxylate, DCA metabolites detected in vivo included monochloroacetic
acid and a series of downstream metabolites of glyoxylate (Stacpoole et al., 1998). Thus, the
formation of glyoxylate metabolites maybe of toxicologic importance in BCA metabolism and is
described here briefly. Glyoxylate may be metabolized through competing pathways to form
glycine, glycolate, oxalate, and CO2 (Stacpoole et al., 1998).. Glyoxylate can also be converted to
glycolate by glyoxylate reductase (Michal, 1999). Glycine may be formed through the activity of
glycolate aminotransferase (formally known as alanine: glyoxylate aminotransferase). Glycine,
which can be incorporated into proteins, used in the synthesis of serine, or degraded releasing
carbon dioxide (Michal, 1999). Conversion to oxalate may occur via a (S)-2-hydroxyacid
dehydrogenase such as lactate dehydrogenase
Few data are available regarding the kinetics of brominated acetic acid metabolism. In the
single toxicokinetics study identified in the literature that examined absorption, distribution,
metabolism, and excretion (Schultz et al., 1999), only one dose of BCA was used and analysis of
metabolic saturation could not be conducted. Gonzalez-Leon et al. (1999, published abstract)
used microsomes to study the effect of BCA pre-treatment on metabolic inhibition following
administration of subsequent BCA doses. Microsomal fractions were prepared from the livers of
male F344 rats given 2000 mg/L BCA in drinking water for 2 weeks and in vitro metabolism was
assessed. Pretreatment reduced the Vmax by 50% to 75%, while the Iv remained unchanged,
indicating possible noncompetitive inhibition of metabolism. However, in the absence of
additional detail on laboratory and analytic methods, it is not clear whether metabolism actually
occurred in the microsomes or whether the microsomal samples were contaminated with cytosol,
and metabolism occurred in the cytosol. Thus, an alternative interpretation of the data is that the
observed metabolism was due to cytosolic contamination of the microsomal fraction, and that
pretreatment with BCA inhibited the cytosolic enzyme. Anderson et al. (1999) administered i.p.
injections of 0.3 mmol BCA to male F344 rats (3/dose) and measured GST-Zeta activity in liver
EPA/OW/OST/HECD III-14
-------
Drinking Water Criteria Document for Brominated Acetic Acids
cytosol 12 hours later. BCA reduced GST-Zeta activity to 19% of that in saline-treated controls,
indicating a possible mechanism for auto-inhibition of metabolism. The toxicologic implications
of these findings are not yet clear because it is not known whether the toxic moiety is the parent
compound or a metabolite. However, this finding is of interest because DCA is known to inhibit
its own metabolism through inhibition in both rodents (Line et al., 1993) and humans (Stacpoole
et al., 1998), and BCA may also exhibit similar metabolic activity. In the case of DCA inhibition
is the result of DCA reaction with and subsequent modification of GST zeta.(Tong et al., 1998a).
Dibromoacetic acid
As for BCA, most of the information on the metabolism of DBA is indirect or based on
analogy to chlorinated acetic acids. These indirect data suggest that DBA is likely to be rapidly
metabolized. Comparisons of DBA renal and blood clearance in the Schultz et al. (1999) study
revealed that less than 1% of blood clearance of the parent compound was accounted for by
urinary excretion, suggesting that metabolism is the major contributor to DBA removal from the
blood. However, these data were obtained following IV administration of the test compound and,
thus, the degree of metabolism in different tissues cannot be determined from the data provided.
The limited oral bioavailability of DBA suggests that the liver maybe an important site for first-
pass metabolism. The proposed metabolic pathway for DBA, based on the data presented in the
following paragraphs, is shown in Figure III-l.
Shultz et al. (1999) proposed that similarities between the toxicokinetics of DCA and
those of DBA and BCA indicate that these compounds are likely to share metabolic pathways.
This hypothesis is supported by several comparative in vitro metabolic studies (Schultz et al.,
1998; Tong et al., 1998a) that demonstrated that GST-Zeta can catalyze the metabolism of DBA,
as well as BCA and DCA, to glyoxylic acid (Tong et al., 1998a)
EPA/OW/OST/HECD III-l 5
-------
Drinking Water Criteria Document for Brominated Acetic Acids
As described in the previous section for BCA, very little data are available regarding the
kinetics of brominated acetic acid metabolism. In the single toxicokinetics study identified in the
literature that examined absorption distribution, metabolism, and excretion (Schultz et al., 1999),
only one dose of DBA was used. Therefore, no conclusions regarding metabolic saturation can be
made. In a study similar to that reported in the previous section for BCA, Gonzalez-Leon et al.
(1999, published abstract) studied the effects of DBA pre-treatment on in vitro metabolic
inhibition in the liver microsomes of male F344 rat liver following in vivo administration of 2000
mg/L DBA in drinking water for 2 weeks. Similar to BCA, pretreatment reduced the Vmax by 50%
to 75%, while the K^ remained unchanged, suggesting metabolic inhibition; however it was not
clear from the abstract whether metabolism occurred in the microsomes or in the cytosol.
Anderson et al. (1999) administered i.p. injections of 0.3 mmol DBA to male F344 rats (3/dose)
and measured GST-Zeta activity in liver-cytosol preparations 12 hours later. DBA administration
reduced GST-Zeta activity to 17% of that of saline-treated controls, indicating a possible
mechanism for auto-inhibition of metabolism. The toxicologic implications of these findings are
not yet clear because it is not known whether the toxic moiety is the parent compound or a
metabolite. However, it is of interest that similar auto-inhibition was seen for DCA metabolism
(TongetaL, 1998a).
D. Excretion
Monobromoacetic acid
No studies on the excretion of MBA following exposure by any route were identified.
Bromochloroacetic acid
EPA/OW/OST/HECD III-16
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Schultz et al. (1999) measured parent compound concentrations in the blood, urine, and
feces after oral or i.v. dosing of male F344 rats with 500 |_imol/kg (87 mg/kg) BCA. Blood
measurements were taken at 0, 5, 10, 20, 30, 60 and 90 minutes, and at 3, 4, 6, 8 and 12 (i.v. only)
hours post-dosing; urine and feces were collected 24 hours following dosing. BCA was rapidly
cleared from the blood, with apparent bi-exponential elimination following i.v. administration.
There was an initial, rapid decline in blood concentration, corresponding to a short distributive
phase, followed by a long linear decline. In the concentration-time profile, peculiarities were
noted in the profile which suggested that physiological mechanisms or processes were involved
other than multiple distribution phases (e.g., 2- 3 compartments or distribution phases).1 As a
result, the authors were doubtful that the unique appearance of the profiles was due to a prolonged
distribution phase(s), and chose to analyze the data using simple non-compartmental methods,
which require fewer assumptions than compartmental models with regard to distribution within
the animal. Therefore, they provided two estimates of half-life (t1/2): one relying on the initial
decline in the profiles (0-4 hours) and another using the full or complete profile. Truncating the
concentration-time profiles had no significant effect on the AUC; however, the elimination half-
life was markedly altered by more than five-fold; t1/2 was 3.93 hours for the complete profile versus
0.74 hours for the truncated profile. After oral administration, blood levels reached a maximum at
1.5 hours following dosing and declined rapidly during the next 6 hours; t max was 1.5 hours.
Removal of parent compound from the blood appeared to be rapid, due mainly to
biotransformation (Schultz et al., 1999). The urine and feces were minimal contributors to overall
blood clearance. Urinary clearance of the parent compound accounted for 2% of the total
clearance and feces contained negligible amounts of BCA. The study authors did not measure
either putative metabolites or expired CO2. Therefore, it was not possible to determine the
contribution of each route of excretion to the total administered dose of the parent compound.
Personal communication, I. Schultz, Battelle Laboratories, Washington
EPA/OW/OST/HECD 111-17
-------
Drinking Water Criteria Document for Brominated Acetic Acids
No data on the excretion of BCA following exposure by the inhalation or dermal routes
were identified.
Dibromoacetic acid
Schultz et al. (1999) also measured DBA concentrations in the blood, urine, and feces 24
hours after oral or i.v. dosing of male F344 rats with 500 |_imol/kg DBA (109 mg/kg). Blood
measurements were taken at 0, 5, 10, 20, 30, 60 and 90 minutes, and at 3, 4, 6 and 8 hours post-
dosing; urine and feces were collected 24 hours following dosing. DBA was rapidly cleared from
the blood, with apparent bi-exponential elimination following i.v. administration. After oral
administration, blood levels reached a maximum about one hour following dosing and declined
rapidly during the next six hours. Similar to BCA, features were noted in the DBA concentration-
time profile which suggested physiological mechanisms or processes were involved other than
multiple distribution phases (e.g., 2- 3 compartments or distribution phases), and the study authors
chose to analyze the data using simple non-compartmental methods, which require fewer
assumptions than compartmental models with regard to distribution within the animal. As with
BCA, they provided two estimates of t1/2: one relying on the initial decline in the profiles (0-4
hours) and another using the full or complete profile. Unlike BCA, DBA elimination half-lives
were similar for both the complete and truncated profiles (0.72 versus 0.52 hours). The urine and
feces were minimal contributors to overall blood clearance. Urinary clearance of DBA accounted
for only a small fraction of the total clearance and was less than 1% of total clearance; negligible
amounts of DBA were found in the feces. The study authors did not measure either putative
metabolites or expired CO2. Therefore, it was not possible to determine the contribution of each
route of excretion to the total administered dose of the parent compound.
EPA/OW/OST/HECD III-18
-------
Drinking Water Criteria Document for Brominated Acetic Acids
No data on excretion of DBA following exposure by the inhalation or dermal routes were
identified.
E. Bioaccumulation and Retention
Monobromoacetic acid
No studies on the bioaccumulation and retention of MBA following exposure by any route
were identified.
Bromochloroacetic acid
The available data on the potential for bioaccumulation or retention of BCA are very
limited. Schultz et al. (1999) demonstrated that a single dose of BCA is rapidly eliminated from
the blood. Following intravenous dosing, BCA did not appear to bind significantly to plasma
proteins or accumulate in blood cells. The unbound fraction in plasma and the plasma to blood
concentration approached unity, and the apparent volume of distribution was similar to the total
body water volume for rats (Reinoso et al., 1997), suggesting that BCA distributed uniformly
outside the vascular system and was unlikely to accumulate significantly in peripheral tissues.
However, only a single dose was administered and no measurements were collected in tissues
other than blood. Therefore, the extent of bioaccumulation or retention cannot be determined.
BCA at physiologic pH is not lipophilic (Schultz et al., 1999), suggesting little affinity for
accumulation in adipose tissue.
No studies were identified on the bioaccumulation and retention of BCA following
exposure by the inhalation or dermal routes.
EPA/OW/OST/HECD 111-19
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Dibromoacetic acid
Similar to BCA, DBA was rapidly eliminated from the blood in the toxicokinetic study by
Schultz et al. (1999), and did not bind significantly to plasma proteins or accumulate in blood
cells. Both the unbound fraction in plasma and the plasma:blood concentrations were close to
unity and the apparent volume of distribution was similar to the total body-water volume for rats
(Reinoso et al., 1997), suggesting uniform distribution outside the vascular system and little
affinity for accumulation in peripheral tissues. However, as with BCA, only a single dose was
administered and no measurements were collected in tissues other than blood. Rapid blood
clearance of DBA was also suggested by the absence of detectable blood levels in an NTP (1999)
immunotoxicity study. Kennedy et al. (1993) reported on the tissue retention of radiolabel
following oral dosing of rats with [14 C]l,l,2,2-tetrabromoethane, a compound whose major
urinary metabolite is DBA. After 96 hours, the percent of administered dose retained in the body
ranged from 14% to 22% and was not dose-dependent. The largest percentage of dose was in the
carcass, followed by the liver, blood, and gastrointestinal tract, with lesser amounts found in the
kidney and fat. These data suggest that 1,1,2,2-tetrabromoethane and/or one (or more) of its
metabolites were widely distributed; however, it is not known whether the observed tissue-
distribution pattern would be similar following direct oral dosing with DBA. Further, the results
of total [14 C] distribution did not identify whether the tissue distribution represented the parent
compound or its metabolites, and the specific metabolites were not identified.
DBA at physiologic pH is not lipophilic (Schultz et al., 1999), suggesting little propensity
for retention or accumulation in adipose tissue. However, Klinefelter et al. (2000, abstract)
reported the presence of DBA in the milk of Sprague-Dawley females, following high-dose
exposure during pregnancy and lactation, at concentrations greater than those measured in
females' blood serum (Klinefelter et al., 2000, abstract), suggesting that retention or accumulation
EPA/OW/OST/HECD 111-20
-------
Drinking Water Criteria Document for Brominated Acetic Acids
may be possible under certain physiologic conditions. A full report of this study has not been
published and thus these findings cannot be comprehensively evaluated. In contrast, Christian et
al. (2001) did not detect levels of DBA in the milk of lactating Sprague-Dawley rats at drinking
water concentrations up to 1000 ppm (estimation of daily doses to lactating dams was confounded
by concomitant pup water consumption). DBA was reliably detected on GD 21 in placental tissue
at 1000 ppm (81.6 mg/kg/day), and in fetal plasma at 500 and 1000 ppm (maternal doses of 48.5
and 81.6 mg/kg/day, respectively), at concentrations which were generally similar to those
measured in the plasma of pregnant females on GD 20. Higher concentrations of DBA in the
plasma of male and female rats noted on SD 14 as compared with SD 1 were attributed by the
authors to variability and not to accumulation. Placental DBA levels on GD 21 were lower than
those observed in maternal serum plasma on GD 20 except at 1000 ppm. Christian et al. (2001)
concluded that although DBA freely crossed the placenta and distributed to the fetus during
gestation, it did not appear to bioaccumulate.
No studies were identified on the bioaccumulation and retention of DBA following
exposure by the inhalation or dermal routes.
F. Summary
No toxicokinetic studies of MBA have been identified in the literature. Brominated acetic
acids appear to be rapidly absorbed from the gastrointestinal tract (Schultz et al., 1999). Key data
from this study are summarized in Table III-l. Following single-dose intravenous or oral-gavage
exposure, both BCA and DBA were rapidly cleared from blood and had short plasma elimination
half-lives (Schultz et al., 1999). However, the extent of tissue distribution is not known because
tissue distribution studies of these compounds have not been conducted. Neither BCA nor DBA
are lipophilic at physiologic pH (Schultz et al., 1999), suggesting a low propensity to accumulate
EPA/OW/OST/HECD 111-21
-------
Drinking Water Criteria Document for Brominated Acetic Acids
in fat. Following repeated exposure in drinking water, DBA was detected in the blood serum and
milk of lactating female Sprague-Dawley rats exposed from gestation day 15 to postnatal day 20
(Klinefelter, 2000, abstract), but was not detected in the blood plasma of nonlactating female
B6C3F1 mice after 28 days of exposure (NTP, 1999). The discrepancy in these findings might be
due to sampling differences, species differences, and/or differences in physiologic status.
The metabolism of BCA and DBA has also not been thoroughly investigated. Limited in
vitro data (Schultz et al., 1998; Tong et al., 1998a, 1998b) and the single comparative
toxicokinetics study by Schultz et al. (1999) suggest that both BCA and DBA are metabolized in a
manner similar to DCA. Potential pathways of brominated acetic-acid metabolism to glyoxylic
acid have been proposed based on analogy to chlorinated acetic acids (reviewed in Stacpoole et
al., 1998); these pathways are mediated through a recently-identified class of GST isoenzymes,
GST-Zeta. It is not clear whether the effective toxicologic moiety is the parent compound or an
active metabolite. Overall, the data are consistent with DBA and DCA being rapidly excreted and
having little propensity for bioaccumulation. Lactational exposure may be a route of concern
because of the presence of DBA in the milk of lactating Sprague-Dawley females at
concentrations greater than those measured in the females' blood serum (Klinefelter et al., 2000,
abstract). In contrast, Christian et al. (2001) did not detect DBA in the milk samples of lactating
Sprague-Dawley rats exposed to high DBA drinking water concentrations, although DBA was
measurable in fetal plasma on gestation day 21. Christian et al. (2001) observed that, although
DBA freely crossed the placenta in pregnant Sprague-Dawley rats, the attained maternal and fetal
plasma levels were associated with the amount and timing of water consumption and did not
appear to accumulate.
EPA/OW/OST/HECD 111-22
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Chapter IV. Human Exposure
A. Drinking Water Exposure
MBA, BCA, and DBA have been identified as drinking-water disinfection byproducts
under the Information Collection Rule (U.S. EPA, 1994) and are being assessed for regulatory
consideration in the Stage 2 Disinfectants/Disinfection Byproducts Rule to be promulgated.
Therefore, this section will examine the occurrence of these compounds in drinking water.
A.I National Occurrence Data for MBA, BCA, and DBA
This section presents the data collected from the Information Collection Rule (ICR)
databases, which provide data from surface- and ground-water systems serving at least 100,000
persons. This data base includes information gathered for 18 months from July 1997 to December
1998.
Section A. 1.1 describes the ICR data set and analysis techniques used to present the data
for the plants that submitted data under the ICR. The data in Sections A.I and A.2 were taken
from the online version of the ICR database (U.S. EPA, 2000a), and the explanation of the
EPA/OW/OST/HECD IV-1
-------
Drinking Water Criteria Document for Brominated Acetic Acids
methods used was taken from the Draft EPA Document on Stage 2 Occurrence and Exposure
Assessment for Disinfectants and Disinfection Byproducts (D/DBPs) in Public Drinking Water
(U.S. EPA, 2000b).
A.1.1 ICR Plants
The ICR generated plant-level sets of data that link water quality and treatment from
source to tap, and aid in understanding the seasonal variability in these relationships. The database
contains information from 18 monthly or 6 quarterly samples from 7/97 to 12/98 from
approximately 300 large systems covering approximately 500 plants. These samples were tested
for influent and finished water-quality parameters (e.g., TOC, temperature, pH, alkalinity), DBP
levels, and disinfectant residuals. Samples were collected at several locations throughout the
distribution system to cover the entire range of residence times during which DBFs can form in
the finished water. Over the 18-month period, approximately 1470 samples were taken from 305
plants with surface water as their source, and approximately 580 samples were taken from 123
plants with groundwater as their source. For more detailed information, such as sampling
locations and frequencies, refer to the ICR Data Analysis Plan (U.S. EPA, 2000c).
EPA/OW/OST/HECD IV-2
-------
Drinking Water Criteria Document for Brominated Acetic Acids
A. 1.2 Quarterly Distribution System Average and Highest Value for MBA, BCA, and DBA
This section describes the data-analysis techniques employed for the analysis of observed
data for water-quality parameters, and for MBA, BCA, and DBA concentrations. All data are
categorized according to the types of source water - surface or ground. Plants having both
surface- and ground-water sources (mixed) or that purchase water are included in the surface-
water category. Quarterly Distribution System Average and Highest Value for the brominated
acetic acids are presented in Table IV-1.
The quarterly distribution-system average is an average of the following four distinct
locations in the distribution system.
• Distribution System Equivalent (DSE) location;
• Average 1 (AVG 1) and Average 2 (AVG 2) locations: Two sample points in the
distribution system representing the approximate average residence time as designated
by the water system; and
• Distribution System Maximum: Sample point in the distribution system having the
highest residence time (or approaching the longest time) as designated by the water
system
EPA/OW/OST/HECD IV-3
-------
Drinking Water Criteria Document for Brominated Acetic Acids
The quarterly distribution-system highest value is the highest of the four distribution-
system samples collected by a plant in a given quarter.
EPA/OW/OST/HECD IV-4
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-1. Bromoacetic Acids
Quarterly Distribution System Average and Highest Value1
Source
Quarterly
Dist. Sys.
Plants
N
PctND
%
Mean
fig/L
Median
fig/L
STD
fig/L
Min
fig/L
Max
fig/L
plO
fig/L
p90
fig/L
MBA
SW
GW
Average
High
Average
High
305
305
123
123
1470
1470
581
581
83.06
83.06
86.57
86.57
0.25
0.41
0.16
0.27
0.00
0.00
0.00
0.00
0.98
1.84
0.54
0.92
ND
ND
ND
ND
26.00
58.10
6.68
12.00
0.00
0.00
0.00
0.00
0.96
1.50
0.38
1.30
BCA
SW
GW
Average
High
Average
High
305
305
123
123
1474
1474
584
584
9.70
9.70
48.63
48.63
3.61
4.45
1.47
2.18
2.88
3.50
0.28
1.10
3.08
3.86
2.15
3.40
ND
ND
ND
ND
24.18
41.90
11.28
41.00
0.25
1.00
0.00
0.00
7.70
9.00
4.50
6.40
DBA
SW
GW
Average
High
Average
High
305
305
123
123
1484
1484
584
584
60.58
60.58
56.16
56.16
1.09
1.39
0.82
1.20
0.00
0.00
0.00
0.00
2.08
2.52
1.48
1.88
ND
ND
ND
ND
14.25
19.00
13.00
16.00
0.00
0.00
0.00
0.00
3.68
4.30
2.58
3.70
1 Nondetects are treated as zero.
Source: SW - Surface Water, GW - Groundwater
Quarterly Dist. Sys: Quarterly Distribution System Average
Plants:
N:
PctND:
Mean:
Median:
STD:
Number of plants sampled
Number of samples
Percent samples nondetect
Arithmetic mean of all samples
Median value of all samples
Standard deviation
EPA/OW/OST/HECD
IV-5
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-1. Bromoacetic Acids
Quarterly Distribution System Average and Highest Value1
Source
Quarterly
Dist. Sys.
Plants
N
PctND
%
Mean
fig/L
Median
fig/L
STD
fig/L
Min
fig/L
Max
fig/L
plO
fig/L
p90
fig/L
Min:
Max:
plO:
p90:
ND:
Minimum Value
Maximum Value
10th perc entile
90th perc entile
Nondetected
EPA/OW/OST/HECD
IV-6
-------
Drinking Water Criteria Document for Brominated Acetic Acids
The mean concentrations of MBA (averaged across the four sampling locations) were
0.16 and 0.25 ug/L in groundwater and surface water, respectively. The mean concentrations of
BCA (averaged across the four sampling locations) were 1.47 and 3.61 ug/L in groundwater and
surface water, respectively. The mean concentrations of DBA (averaged across the four sampling
locations) were 0.82 and 1.09 ug/L in groundwater and surface water, respectively. Examination
of the data using the Student's t-test indicates that the mean concentrations of MBA, BCA, and
DBA in surface water was significantly higher that the mean concentrations of these chemicals in
ground water. The mean concentrations of BCA were statistically significantly higher (p = 0.05)
than the mean concentrations of DBA, which were significantly higher (p = 0.05), that the mean
MBA concentrations in both surface- and ground-water. The lowest mean concentrations are
associated with the highest percentage of nondetects, which are treated as 0 in the calculation of
the mean, median, standard deviation, and plO values (U.S. EPA, 2000a).
A.2 Factors Affecting the Relative Concentrations of MBA, BCA, and DBA
Sections A.2.1 - A.2.4 contain investigational information and ICR data on the effects of
disinfection chemicals, influent bromide concentration, influent total organic carbon (TOC)
concentration, and seasonal shifts, respectively in MBA, BCA, and DBA concentrations.
EPA/OW/OST/HECD IV-7
-------
Drinking Water Criteria Document for Brominated Acetic Acids
A.2.1 Disinfection Treatment
Chlorination has been the predominant water-disinfection method in the United States.
However, water utilities are considering a shift to alternative disinfectants. Therefore, there is a
need to understand the occurrence of DBFs in drinking water and the factors that may influence
their formation. Several published studies (Boorman et al., 1999; Richardson, 1998; Lykins et al.,
1994; Jacangelo et al., 1989) reported on the formation of brominated acetic acids and other DBFs
under different disinfection conditions.
In a review on drinking-water disinfection byproducts, Boorman et al. (1999) compared
the concentrations of different drinking-water disinfection byproducts, including MBA, BCA, and
DBA, formed by chlorination, ozonation, chlorine dioxide, and chloramination. Most of the data
were available for surface-water systems that used chlorination. For the systems using
chlorination, BCA, with a median and a maximum concentration of 3.2 and 49 |-ig/L, respectively,
was present at the highest concentrations. The median values of both MBA and DBA in
chlorinated water were less than 0.5 |-ig/L, with maximum values of 1.7 and 7.4 |_ig/L,
EPA/OW/OST/HECD IV-8
-------
Drinking Water Criteria Document for Brominated Acetic Acids
respectively. The principal products formed by chloramination were similar to those formed by
chlorination; additional information was not provided. Ozonation of water containing bromide
may produce DBFs such as bromoform, dibromoacetic acid, cyanogen bromide, and bromate.
The total concentration of brominated acetic acids formed by the use of ozonation ranged from 1
to 50 |-ig/L; concentrations for individual compounds were not provided. Chlorine dioxide formed
oxidation by-products similar to those formed by ozonation; additional details were not provided.
Richardson (1998) compared the relative concentrations of DBFs in drinking water using
different treatment methods, and found that chlorination produced the highest concentration of
DBFs, including MBA, BCA, and DBA. Chlorine dioxide and chloramine, when compared to
chlorine, produced fewer and lower concentrations of DBFs. MBA, BCA, and DBA were not
produced by chlorine dioxide in measurable quantities. Compared to chlorine treatment,
chloramine produced 3% to 20% lower levels of by-products, including haloacetic acids.
Ozonation produced insignificant levels of trihalomethanes (THMs). However, when elevated
levels of bromide ion were in the raw water, MBA and DBA were detected following ozonation.
When ozone was the primary disinfectant followed by chloramine, the levels of most DBFs,
including haloacetic acids, were lower than when chloramine was used solely. However, there
EPA/OW/OST/HECD IV-9
-------
Drinking Water Criteria Document for Brominated Acetic Acids
was an observed shift to more brominated species of THMs and haloacetic acids when ozone was
followed by chlorine, than when chorine was used solely.
Lykins et al. (1994) investigated the formation of halogenated DBFs in the water-
distribution system, by predisinfecting and postdisinfecting the water with either chlorine or
chloramine and holding the water for five days. They found that the use of chlorine produced the
highest concentration of halogenated DBFs and that, in general, the concentrations could be
reduced by adding ozone as a predisinfectant with postchlorination. Lykins et al. (1994) also
found that the highest average concentrations of MBA (1.2 ug/L) and BCA (18 ug/L) were
formed when chlorine was the sole treatment method. The average DBA concentration was 0.6
ug/L when chlorine was the sole treatment method. The next highest concentrations of MBA (1.0
ug/L) and BCA (14 ug/L) were observed with ozone treatment followed by chlorine. In contrast,
the average DBA concentration (1.0 ug/L) was slightly higher with ozone treatment followed by
chlorine than when chlorine was the sole treatment method. Chloramine treatment, and ozone
treatment followed by chloramine, resulted in the lowest concentrations of MBA and DBA, with
both treatment methods resulting in 0.1 ug/L MBA and 0.1 ug/L DBA. BCA concentrations were
lower with ozone treatment followed by chloramine (1.0 ug/L BCA) than when chloramine was
the sole treatment method (1.9 ug/L).
EPA/OW/OST/HECD IV-10
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Jacangelo et al. (1989) examined the impact of ozonation on the formation and control of
DBFs in drinking water at four utilities. Treatment modifications were made on the process train
at each full or pilot-scale plant to incorporate ozone in the treatment process. For two of the
utilities in the Jacangelo et al. (1989) study, only total haloacetic acids (HAAs) were measured,
and no measurements were made of individual HAAs. The disinfection schemes that employed
ozonation followed by chloramines as a disinfectant resulted in large decreases in HAAs relative
to chlorination. However, the sample size did not allow for statistical analysis of the data
(Jacangelo et al., 1989). For two other utilities that measured individual HAAs, preozonation
followed by chlorination decreased the total HAAs by 14 to 50%, when compared with
chlorination only. The concentration of MBA was essentially the same with and without
preozonation, while DBA increased with ozonation. BCA was not measured in this study. The
authors suggested that ozone reacts with bromide ions in the source water, resulting in the
formation of hypobromous acid. Reaction of hypobromous acid and natural organic matter can
produce brominated HAAs. When preozonation and postchlorination are practiced, competition
exists between hypochlorous acid and hypobromous acid for organic matter, leading to varying
concentrations of chlorinated and brominated HAAs (Jacandgelo et al., 1989). In addition,
EPA/OW/OST/HECD IV-11
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Jacangelo et al. (1989) noted that the concentrations of brominated acetic acids increased with
increasing bromide concentrations in the source water.
Miltner et al. (199) studied DBF formation and control in three surface water pilot plants
employing three different disinfectant methods (chlorine, ozone followed by chlorine, and ozone
followed by chloramine). In an examination of the data using the Student's t-test, the authors
found that the amount of BCA measured in finished water and in simulated distribution waters
was lower (p = 0.05) when ozonation was combined with chlorination or with cloramination than
when chlorination was used a lone. However, ozonation had no effect (p = 0.05) on the formation
of MBA, and the formation of DBA was higher (p = 0.05) when ozonation was followed by
chlorination than when chlorination alone was used.
A.2.1.1 Disinfection Treatment in ICR Data Base
EPA/OW/OST/HECD IV-12
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Data on the concentrations of MBA, BCA, and DBA were gathered from plants using
several disinfection treatments. Those chemical-disinfection treatments most commonly used
(used by 10% or more of the plants evaluated), along with the ozonation treatments, are presented
in Tables IV-2, IV-3, and IV-4 for MBA, BCA, and DBA, respectively.
Examination of the data using the Student's t-test indicates that, for all chemical-
disinfection treatments used for surface water, the mean concentrations of BCA were statistically
significantly higher (p = 0.05) than the mean concentrations of DBA which were significantly
higher (p = 0.05) than the mean concentrations (Tables IV-2 and IV-3). For all chemical-
disinfection treatments used for ground water, the mean concentrations of BCA were significantly
higher (p = 0.05) than the mean concentrations of MBA. The mean concentrations of BCA in
groundwater were significantly higher ( p = 0.05)than the mean DBA concentrations only in plants
using ozone and choramine.
EPA/OW/OST/HECD IV-13
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-2. MBA by Disinfection Method
(Quarterly Distribution System Average)1
Source
SW
GW
Disinfection
Chemicals
cyci2
C12_CLM/CLM
O3/CL2
O3/CLM
/C12
C12/C12
O3/CLM
Plants
180
66
7
10
67
39
1
N
814
307
25
49
299
170
6
PctND
%
85.63
80.78
88.00
85.71
87.29
85.88
100.00
Mean
fig/L
0.24
0.27
0.11
0.10
0.18
0.15
0.00
Median
fig/L
0.00
0.00
0.00
0.00
0.00
0.00
0.00
STD
fig/L
1.14
0.72
0.33
0.30
0.64
0.45
0.00
Min
fig/L
ND
ND
0.00
0.00
ND
ND
0.00
Max
fig/L
26.00
5.73
1.28
1.48
6.68
2.53
0.00
plO
fig/L
0.00
0.00
0.00
0.00
0.00
0.00
0.00
p90
fig/L
0.75
1.10
0.28
0.58
0.35
0.51
0.00
1 Nondetects are treated as zero.
Source: SW - Surface Water, GW - Groundwater
C12/C12: Free chlorine in Water Treatment Plant (WTP) and Distribution System (DS).
C12_CLM/CLM: Free chlorine followed by chloramine in WTP and chloramine in DS.
/C12: No disinfectant in WTP and free chlorine in DS.
O3/C12: Ozone in WTP and free chlorine in DS.
O3/CLM: Ozone in WTP and chloramine in DS.
Plants: Number of plants sampled
N: Number of samples
PctND: Percent samples nondetect
Mean: Arithmetic mean of all samples
Median: Median value of all samples
STD: Standard deviation
Min: Minimum Value
Max: Maximum Value
plO: 10th percentile
p90: 90th percentile
EPA/OW/OST/HECD
IV-14
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-3. BCA Acid by Disinfection Method
(Quarterly Distribution System Average) *
Source
SW
GW
Disinfection
Chemicals
cyci2
C12_CLM/CLM
O3/CL2
O3/CLM
/C12
C12/C12
O3/CLM
Plants
180
66
7
10
67
39
1
N
816
306
25
49
301
170
6
PctND
%
11.40
1.63
4.00
22.45
65.78
45.53
0.00
Mean
fig/L
3.14
4.67
2.54
2.19
0.87
1.27
2.68
Median
fig/L
2.54
4.01
2.13
2.15
0.00
0.31
2.55
STD
fig/L
2.77
3.28
1.77
1.70
1.77
1.80
0.69
Min
fig/L
ND
ND
0.00
0.00
ND
ND
2.05
Max
fig/L
18.73
23.80
5.68
6.60
10.70
7.88
3.88
plO
fig/L
0.00
1.35
0.33
0.00
0.00
0.00
2.05
p90
fig/L
6.83
8.83
4.90
4.55
2.70
3.78
3.88
Nondetects are treated as zero.
Source: SW - Surface Water, GW - Groundwater
C12/C12: Free chlorine in Water Treatment Plant (WTP) and Distribution System
C12_CLM/CLM: Free chlorine followed by chloramine in WTP and chloramine in DS.
/C12: No disinfectant in WTP and free chlorine in DS.
O3/C12: Ozone in WTP and free chlorine in DS.
O3/CLM: Ozone in WTP and chloramine in DS.
Plants: Number of plants sampled
N: Number of samples
PctND: Percent samples nondetect
Mean: Arithmetic mean of all samples
Median: Median value of all samples
STD: Standard deviation
Min: Minimum Value
Max: Maximum Value
plO: 10th percentile
p90: 90th percentile
(DS).
EPA/OW/OST/HECD
IV-15
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-4. DBA by Disinfection Method
(Quarterly Distribution System Average) *
Source
SW
GW
Disinfection
Chemicals
cyci2
C12_CLM/CLM
O3/CL2
O3/CLM
/C12
C12/C12
O3/CLM
Plants
180
66
7
10
67
39
1
N
823
308
25
49
303
169
6
PctND
%
68.53
51.62
68.00
46.94
67.66
46.15
50.00
Mean
Hg/L
0.68
1.36
1.02
1.01
0.68
1.01
0.41
Median
Hg/L
0.00
0.00
0.00
0.25
0.00
0.25
0.14
STD
Hg/L
1.55
2.11
2.27
1.50
1.52
1.62
0.55
Min
Hg/L
ND
ND
0.00
0.00
ND
ND
0.00
Max
Hg/L
12.50
14.25
8.15
7.30
13.00
7.25
1.25
plO
Hg/L
0.00
0.00
0.00
0.00
0.00
0.00
0.00
p90
Hg/L
2.20
4.83
2.83
2.80
2.53
3.15
1.25
1 Nondetects are treated as zero.
Source: SW - Surface Water, GW - Groundwater
C12/C12: Free chlorine in Water Treatment Plant (WTP) and Distribution System (DS).
C12_CLM/CLM: Free chlorine followed by chloramine in WTP and chloramine in DS.
/C12:
O3/C12:
O3/CLM:
Plants:
N:
PctND:
Mean:
Median:
STD:
Min:
Max:
plO:
p90:
No disinfectant in WTP and free chlorine in Distribution System.
Ozone in WTP and free chlorine in DS.
Ozone in WTP and chloramine in DS.
Number of plants sampled
Number of samples
Percent samples nondetect
Arithmetic mean of all samples
Median value of all samples
Standard deviation
Minimum Value
Maximum Value
10th perc entile
90th perc entile
EPA/OW/OST/HECD
IV-16
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Examination of the ICR data using the Student's t-test indicates that the mean
concentrations of BCA and DBA were significantly higher (p = 0.05) in surface water plants using
chlorine followed by chloramine than in those using free chlorine alone. There were no
significant differences (p = 0.05) between the mean concentrationof MBA in surface water plants
using chlorine followed by chloramine and the concentration those solely using chlorine.
In plants with groundwater as a source, the mean concentrtions of BCA and DBA in plants
with no disinfectant in the treatment plant and with free chlorine in the distribution system were
significantly lower (p = 0.05) than the mean concentrations of the same chemicals in plants with
free chlorine in both the treatment plant and the distribution system. There were no significant
differences (p = 0.05) between the mean concentrtions of MBA in plants with no disinfectant in
the treatment plant and with free chlorine in the distribution system and those with free chlorine in
both the treatment plant and the distribution system.
Only a very limited number of plants used ozonation in combination with either chlorine
or chloramine in the distribution system. An examination of the ICR data in Tables IV-2, IV-3
and IV-4 using the Student's t-test indicates that, with one exception, the mean concentrations of
BCA were significantly lower in surface-water plants that use ozone in the water-treatment plant
EPA/OW/OST/HECD IV-17
-------
Drinking Water Criteria Document for Brominated Acetic Acids
and free chlorine or free chloramine in the distribution system than in plants using non-ozonation
disinfection processes. This finding was also presented by Lykins et al. (1994). However , there
were no significant differences ( p = 0.05) between the mean concentrations of BCA, in surface-
water plants when free chlorine was used solely, and the BCA concentrtions when both ozone and
chlorine were used. There were no significant differences (p = 0.05) between the concentrations
of MBA and DBA in surface-water plants using common (non-ozonation) disinfection processes
and the concentrations of the same chemicals pin plants with ozone in the water-treatment plant
and free chlorine or free chloramine in the distribution system. In addition, there were no
significant differences ( p = 0.05) for MBA, BCA, or DBA between the two treatments using
ozonation in treating surface water.
There was only 1 groundwater plant that used the ozone/chloramine disinfection method
and statistical analysis was not conducted.
In summary, an analysis of the ICR data suggest that although BCA concentrations in
surface water treated with chlorine are similar to those treated with ozone and chlorine, surface
water plants using ozonation had lower BCA concentrations than those using most common (non-
EPA/OW/OST/HECD IV-18
-------
Drinking Water Criteria Document for Brominated Acetic Acids
ozonation) disinfection processes. In addition, ozonation appeared to have no effect on the
formation of MBA and DBA.
A.2.2 Bromide Concentration
Pourmoghaddas et al. (1993) examined the effects of source water and treatment
characteristics, such as pH, reaction time, chlorine dosage, and bromide-ion concentration, on the
formation of HAAs.
The study quantified nine HAA species in the presence of bromide ion at low, neutral, and
high pH over time at two chlorine dosages. This study found a shift in the distribution of
HAAs from chlorinated to brominated and mixed (bromochlorinated) halogenated species with
increased bromide-ion concentration. Chloride-ion concentration had no observed effect on the
formation of brominated HAAs.
EPA/OW/OST/HECD IV-19
-------
Drinking Water Criteria Document for Brominated Acetic Acids
At the low chlorine dose of 11.5 mg/L, MBA showed a consistent trend toward higher
concentrations as reaction time and bromide-ion concentrations increased. The only apparent
effect of pH was to increase the amount of MBA at the highest bromide concentration (4.5 mg/L).
The highest measured concentration of MBA was at a pH of 5 and a bromide concentration of 4.5
mg/L (Pourmoghaddas et al, 1993). DBA concentrations increased with increasing bromide-ion
concentration and reaction time. Changes in pH had little influence on the formation of DBA
(Pourmoghaddas et al., 1993).
BCA was formed only if the bromide ion was present. BCA concentrations increased with
reaction time and were significantly lower at pH 9.4. The highest observed concentration of BCA
was at 1.5 mg/L bromide-ion concentration, while BCA levels decreased at the highest bromide
concentration of 4.5 mg/L (Pourmoghaddas et al., 1993). It is possible that the decreased BCA
levels at high bromide concentrations reflect the preferential formation of DBA over BCA under
such conditions, but the data provided are insufficient to test this hypothesis.
A.2.2.1 Bromide Concentration in ICR Data Base
EPA/OW/OST/HECD IV-20
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Tables IV-5, IV-6, and IV-7 present the formation of MBA, BCA, and DBA, respectively,
as a function of influent bromide concentrations.
Bromide concentrations tended to be lower in plants using surface water as a source than
in those using groundwater as a source. For example, 114 of the 305 plants using surface water as
the source (37%) had influent bromide levels below the minimal reporting limit (MRL) of 20 ppb,
while only 13 of the 123 plants using groundwater as the source (11%) had influent bromide
levels below the MRL.
EPA/OW/OST/HECD IV-21
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-5. MBA by Influent Bromide Concentration
(Quarterly Distribution System Average) 1
Source
SW
GW
Influent
Bromide
Cone, (ppb)
100
100
Plants
114
41
48
59
39
13
11
26
32
41
N
556
200
221
282
192
66
50
109
148
208
PctND
%
90.83
88.50
83.26
77.30
64.06
93.94
98.00
96.33
83.11
78.85
Mean
fig/L
0.14
0.21
0.45
0.24
0.41
0.07
0.01
0.03
0.18
0.27
Median
fig/L
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
STD
fig/L
0.59
0.72
2.02
0.57
1.12
0.30
0.04
0.18
0.67
0.66
Min
fig/L
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
Max
fig/L
6.63
4.03
26.00
3.95
7.20
1.60
0.30
1.45
6.68
3.65
plO
fig/L
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
p90
fig/L
0.00
0.38
1.13
1.03
1.90
0.00
0.00
0.00
0.63
1.20
1 Nondetects are treated as zero.
Source: SW - Surface Water, GW - Groundwater
Plants: Number of plants sampled
N: Number of samples
PctND: Percent samples nondetect
Mean: Arithmetic mean of all samples
Median: Median value of all samples
STD: Standard deviation
Min: Minimum Value
Max: Maximum Value
plO: 10th percentile
EPA/OW/OST/HECD
TV-22
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-5. MBA by Influent Bromide Concentration
(Quarterly Distribution System Average) 1
Source
Influent
Bromide
Cone, (ppb)
Plants
N
PctND
/o
Mean
"§/L
Median
"§/L
STD
"§/L
Min
"§/L
Max
fig/L
plO
"§/L
p90
"§/L
p90: 90th percentile
MRL: Minimum Reporting Limit
ND: Nondetect
EPA/OW/OST/HECD
IV-23
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-6. BCA by Influent Bromide Concentration
(Quarterly Distribution System Average) 1
Source
SW
GW
Influent
Bromide
Cone, (ppb)
100
100
Plants
114
41
48
59
39
13
11
26
32
41
N
559
197
225
282
192
66
51
108
150
209
PctND
%
22.00
0.51
3.11
3.90
0.52
78.79
58.52
49.07
48.67
36.36
Mean
fig/L
1.76
3.32
3.88
5.43
6.29
0.35
0.69
1.16
1.22
2.35
Median
fig/L
1.65
3.03
3.40
4.96
6.01
0.00
0.00
0.25
0.26
1.83
STD
fig/L
1.79
1.67
2.94
3.26
3.52
1.34
1.19
1.61
1.86
2.60
Min
fig/L
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
Max
fig/L
23.80
12.28
22.50
15.50
24.18
10.25
6.00
7.05
7.88
11.28
plO
fig/L
0.00
1.63
0.93
1.48
2.60
0.00
0.00
0.00
0.00
0.00
p90
fig/L
3.43
5.30
7.18
9.68
11.20
0.75
2.25
3.58
4.23
6.25
1 Nondetects are treated as zero.
Source: SW - Surface Water, GW - Groundwater
Plants: Number of plants sampled
N: Number of samples
PctND: Percent samples nondetect
Mean: Arithmetic mean of all samples
Median: Median value of all samples
STD: Standard deviation
Min: Minimum Value
Max: Maximum Value
EPA/OW/OST/HECD
IV-24
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-6. BCA by Influent Bromide Concentration
(Quarterly Distribution System Average) 1
Source
plO:
Influent
Bromide
Cone, (ppb)
Plants
N
PctND
Mean
p.g/L
Median
fig/L
STD
fig/L
Min
p.g/L
Max
p.g/L
plO
fig/L
p90
fig/L
10th perc entile
p90: 90th perc entile
MRL: Minimum Reporting Limit
ND: Nondetect
EPA/OW/OST/HECD
IV-25
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-7. DBA by Influent Bromide Concentration
(Quarterly Distribution System Average) *
Source
SW
GW
Influent
Bromide
Cone, (ppb)
100
100
Plants
114
41
48
59
39
13
11
26
32
41
N
561
201
226
284
193
67
50
109
150
208
PctND
%
91.62
67.66
58.41
32.04
6.74
88.06
90.00
69.72
42.00
40.87
Mean
fig/L
0.08
0.29
0.66
1.70
4.37
0.09
0.12
0.27
1.00
1.38
Median
fig/L
0.00
0.00
0.00
1.33
3.78
0.00
0.00
0.00
0.40
0.60
STD
fig/L
0.32
0.59
1.10
1.84
3.22
0.31
0.49
0.54
1.27
2.00
Min
fig/L
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
Max
fig/L
3.00
3.48
9.58
7.65
14.25
1.68
2.45
2.15
5.83
13.00
plO
fig/L
0.00
0.00
0.00
0.00
0.58
0.00
0.00
0.00
0.00
0.00
p90
fig/L
0.00
1.10
2.03
4.30
8.48
0.28
0.13
1.10
3.10
3.73
1 Nondetects are treated as zero.
Source: SW - Surface Water, GW - Groundwater
Plants: Number of plants sampled
N: Number of samples
PctND: Percent samples nondetect
Mean: Arithmetic mean of all samples
Median: Median value of all samples
STD: Standard deviation
Min: Minimum Value
Max: Maximum Value
EPA/OW/OST/HECD
IV-26
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-7. DBA by Influent Bromide Concentration
(Quarterly Distribution System Average) *
Source
plO:
Influent
Bromide
Cone, (ppb)
Plants
N
PctND
%
Mean
fig/L
Median
fig/L
STD
fig/L
Min
fig/L
Max
fig/L
plO
M-g/L
p90
fig/L
10th perc entile
p90: 90th perc entile
MRL: Minimum Reporting Limit
ND: Nondetect
EPA/OW/OST/HECD
IV-27
-------
Drinking Water Criteria Document for Brominated Acetic Acids
A regression analysis of the ICR data indicates that, with the exception of MBA in surface
water, there was a significant correlation (cc = 0.05) between influent bromide concentration and
the mean concentrations of BCA and DBA in surface water and groundwater.(Tables IV-5, IV-6
and IV-7).
Tan examination of th data using the Student's t-test indicates that in plants treating
surface water, the mean concentrations of BCA were significantly higher (p = 0.05_ than the mean
MBA and DBA concentrtions, for a given bromide concentration range (Tables IV-5, IV-6, and
IV-7). In addition, the mean concentrations of DBA were significantly lower (p = 0.05) than the
mean MBA concentrations at the lowest influent bromide concentration (< minimum reporting
limit of 20 ppb) and higher than MBA concentrations at the two highest influent bromide
concentrations (ranging from 50 to > 100 ppb (Tables IV-5 and IV-7).
The mean concentrations of BCA in surface water were statistically significantly higher
(at p = 0.05) than the mean concentrations of BCA in groundwater at influent bromide
concentrations of 20 - < 30 ppb and at 50 - < 100 ppb. There were no statistically significant
differences (at p = 0.05) between the mean concentrations of DBA and MBA in surface water and
the mean concentrations of these chemicals in groundwater.
EPA/OW/OST/HECD IV-28
-------
Drinking Water Criteria Document for Brominated Acetic Acids
The mean concentrations of BCA and DBA are significantly higher in surface water than
their mean concentrations in groundwater for a given influent bromide concentration (with the
exception of DBA at influent bromide concentrations at <20ppb). However, the mean MBA
concentrations in surface water were significantly higher (p = 0.05) than the mean concentrations
of MBA in groundwater only for the influent bromide concentrations ranging from 20 ppb to < 50
ppb.
A.2.3 Total Organic Carbon (TOC) Concentration in ICR Database
Many researchers have documented that chlorine reacts with natural organic matter in
water to produce a variety of DBFs, including trihalomethanes and haloacetic acids (Reckhow and
Singer, 1990; Reckhow et al., 1990; Marhaba and Van, 2000). Natural organic matter in source
water is generally monitored as total organic carbon (TOC). Arora et al. (1997) analyzed results
of a DBF survey and a two-year DBF-monitoring study of more than 100 treatment plants of the
American Water System from 1989 to 1991, and reported no correlation between raw-water TOC
EPA/OW/OST/HECD IV-29
-------
Drinking Water Criteria Document for Brominated Acetic Acids
and the total of 5 haloacetic acid (HAAS) concentrations in finished and distributed- water
samples. A significant correlation (p < 0.01) was found between TOC and HAAS in plant effluent
and distributed water samples. However, only 11 and 15 percent of the variation in HAAS was
explained by TOC for the distributed-water samples and plant effluent, respectively. No
published studies were located that examined the effect of TOC on the concentration of
brominated acids.
Tables IV-8, IV-9 and IV-10 present data from the ICR database for the concentrations of
MBA, BCA and DBA, respectively, as a function of influent TOC concentrations.
In contrast to the data presented by Arora (1997), a regression analysis of the ICR data
indicates that there was a significant correlation (cc = 0.05) between influent TOC concentration
and the mean concentrations of MBA, BCA, and DBA in surface water (Tabes IV-5, IV-6, and
IV-7)/ This correlation with TOC levels is consistent with the formation of brominated acetic
acids from the reaction of humic acid and hypobromous acid, a compound formed by the reaction
of bromide ion with ozone and/or chlorine in the disinfection process (Chapter n). No such
correlation was observed in groundwater, which had lower overall TOC levels.
EPA/OW/OST/HECD IV-30
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Examination of the data using the Student's t-test indicates that, with a few exceptions, at
a given TOC concentration in surface water and groundwater, the mean concentrations of BCA
were significantly higher (p = 0.05) than the mean concentrations of DBA, which were
significantly higher (p = 0.05) than the mean MBA concentrations (Tables IV-8, IV-9, and IV-10).
ion.
EPA/OW/OST/HECD IV-31
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-8. MBA by Influent Total Organic Carbon (TOC) Concentration
(Quarterly Distribution System Average)
Source
SW
GW
Influent
TOC
Cone, (ppb)
<1
1 -<2
2-<3
3-<4
>4
<1
1 -<2
2-<3
3-<4
>4
Plants
12
58
100
60
69
83
13
8
3
16
N
61
271
479
306
324
405
52
37
7
80
PctND
%
98.36
95.57
84.13
76.14
74.69
89.38
65.38
81.08
85.71
88.75
Mean
Hg/L
0.00
0.04
0.22
0.28
0.50
0.14
0.38
0.30
0.10
0.06
Median
Hg/L
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
STD
Hg/L
0.04
0.26
0.72
0.61
1.75
0.55
0.65
0.74
0.26
0.24
Min
Hg/L
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
Max
Hg/L
0.28
3.20
6.63
3.33
26.00
6.68
2.18
3.08
0.70
1.95
plO
Hg/L
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
p90
Hg/L
0.00
0.00
0.83
1.33
1.55
0.28
1.45
1.23
0.70
0.28
1 Nondetects are treated as zero.
Source: SW - Surface Water, GW - Groundwater
Plants: Number of plants sampled
N: Number of samples
PctND: Percent samples nondetect
Mean: Arithmetic mean of all samples
Median: Median value of all samples
STD: Standard deviation
Min: Minimum Value
Max: Maximum Value
plO: 10th percentile
EPA/OW/OST/HECD
IV-32
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-8. MBA by Influent Total Organic Carbon (TOC) Concentration
(Quarterly Distribution System Average)
Source
Influent
TOC
Cone, (ppb)
Plants
N
PctND
Mean
Hg/L
Median
Hg/L
STD
Hg/L
Min
Hg/L
Max
^g/L
plO
^g/L
p90
^g/L
p90: 90th percentile
ND: Nondetect
EPA/OW/OST/HECD
IV-33
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-9. BCA by Influent Total Organic Carbon (TOC) Concentration
(Quarterly Distribution System Average)
Source
SW
GW
Influent
TOC
Cone, (ppb)
<1
l-<2
2-<3
3-<4
>4
<1
l-<2
2-<3
3-<4
>4
Plants
12
58
100
60
69
83
13
8
3
16
N
62
270
481
309
323
405
55
36
8
80
PctND
%
53.23
15.19
8.11
4.21
3.10
63.46
10.91
38.89
0.00
8.75
Mean
Hg/L
0.74
2.52
3.59
4.17
4.67
0.65
3.06
2.20
3.40
3.99
Median
Hg/L
0.00
1.98
2.80
3.38
3.98
0.00
2.55
1.39
2.86
3.55
STD
Hg/L
1.00
2.41
3.07
3.14
3.19
1.40
1.98
2.43
1.27
2.51
Min
Hg/L
ND
ND
ND
ND
ND
ND
ND
ND
1.78
ND
Max
Hg/L
4.10
15.75
23.80
22.50
24.18
10.70
7.35
7.83
5.08
11.28
plO
Hg/L
0.00
0.00
0.35
1.10
1.48
0.00
0.00
0.00
1.78
0.43
p90
Hg/L
2.25
5.25
7.73
8.18
8.53
2.15
5.93
5.68
5.08
7.64
1 Nondetects are treated as zero.
Source: SW - Surface Water, GW - Groundwater
Plants: Number of plants sampled
N: Number of samples
PctND: Percent samples nondetect
Mean: Arithmetic mean of all samples
Median: Median value of all samples
STD: Standard deviation
Min: Minimum Value
Max: Maximum Value
EPA/OW/OST/HECD
IV-34
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-9. BCA by Influent Total Organic Carbon (TOC) Concentration
(Quarterly Distribution System Average)
Source
Influent
TOC
Cone, (ppb)
Plants
N
PctND
Mean
Hg/L
Median
Hg/L
STD
Hg/L
Min
Hg/L
Max
^g/L
plO
^g/L
p90
^g/L
plO:
p90:
ND:
10th perc entile
90th perc entile
Nondetect
EPA/OW/OST/HECD
IV-35
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-10. DBA by Influent Total Organic Carbon (TOC) Concentration
(Quarterly Distribution System Average)
Source
SW
GW
Influent
TOC
Cone, (ppb)
<1
1 -<2
2-<3
3-<4
>4
<1
1 -<2
2-<3
3-<4
>4
Plants
12
58
100
60
69
83
13
8
3
16
N
62
272
486
309
326
406
54
36
8
80
PctND
%
74.19
72.79
62.76
55.02
48.16
60.59
31.48
61.11
100.00
43.75
Mean
Hg/L
0.45
0.46
0.87
1.69
1.49
0.69
1.54
1.71
0.00
0.67
Median
Hg/L
0.00
0.00
0.00
0.00
0.25
0.00
1.11
0.00
0.00
0.26
STD
Hg/L
0.86
1.10
1.56
2.81
2.45
1.36
1.71
2.68
0.00
0.85
Min
Hg/L
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
Max
Hg/L
2.68
8.48
7.65
12.75
14.25
13.00
6.58
7.25
0.00
3.48
plO
Hg/L
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
p90
Hg/L
2.08
1.70
3.25
6.23
4.33
2.28
3.75
6.73
0.00
2.14
1 Nondetects are treated as zero.
Source: SW - Surface Water, GW - Groundwater
Plants: Number of plants sampled
N: Number of samples
PctND: Percent samples nondetect
Mean: Arithmetic mean of all samples
Median: Median value of all samples
STD: Standard deviation
Min: Minimum Value
Max: Maximum Value
plO: 10th percentile
EPA/OW/OST/HECD
IV-36
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-10. DBA by Influent Total Organic Carbon (TOC) Concentration
(Quarterly Distribution System Average)
Source
Influent
TOC
Cone, (ppb)
Plants
N
PctND
Mean
Hg/L
Median
Hg/L
STD
Hg/L
Min
Hg/L
Max
^g/L
plO
^g/L
p90
^g/L
p90: 90th percentile
ND: Nondetect
A.2.4 Seasonal Shifts
Seasonal shifts in brominated acetic acids were investigated by Krasner et al. (1989). In
September 1987, the USEPA's Office of Drinking Water entered into a cooperative agreement
with the Association of Metropolitan Water Agencies (AMWA) to perform a study of the
occurrence and control of DBFs. The AMWA contracted with the Metropolitan Water District of
Southern California (MWD) to provide management services for the project and to perform the
DBF analysis. In addition, the State of California Department of Health Services (CDHS),
through the California Public Health Foundation (CPHF), contracted with MWD to perform a
similar study in California. Baseline data were gathered on 35 water-treatment facilities,
including 25 water utilities across the United States in the USEPA study and 10 California water
EPA/OW/OST/HECD
IV-37
-------
Drinking Water Criteria Document for Brominated Acetic Acids
utilities in the CDHS study. Levels of MBA and DBA were measured, but BCA was not
evaluated.
During the first quarter (spring 1988), a high correlation was found between DBA and the
disinfectant byproduct dibromochloromethane. In addition, Krasner et al. (1989) reported that
relatively high levels of the measured brominated DBFs were detected at some of the utilities.
These findings suggested that the influence of bromide in the raw water should be evaluated.
Therefore, chloride and bromide analyses was added to the protocol, beginning with the second
quarter (summer 1988) of sampling. Among the 35 facilities, bromide levels ranged from < 0.01
to 3.00 mg/L. At the utility with the highest bromide levels there was a shift in the distribution of
DBFs from the chlorinated DBFs to the brominated DBFs, resulting in DBA as the major
haloacetic acid detected. While there were no clear trends of the concentrations of ions or
brominated acetic acids with season in the composite analysis, DBA levels increased in the
warmer months in the utility with the highest bromide levels. Some observed shifts in utilities
were also seen as the result of drought conditions and saltwater intrusion.
A.2.4.1 Seasonal Shifts in ICR Database
EPA/OW/OST/HECD IV-38
-------
Drinking Water Criteria Document for Brominated Acetic Acids
The seasonal mean concentrations of MBA, BCA, and DBA are presented in Tables IV-
11, IV-12, and IV-13, respectively. For simplicity of presentation, only the data required to
conduct a Student's t-test has been presented here. Additional data can be located in the ICR
database (U.S. EPA, 2000a).
EPA/OW/OST/HECD IV-39
-------
Drinking Water Criteria Document for Brominated Acetic Acids
TableIV-11. MBA by Sample Quarter
(Quarterly Distribution System Average)
Sample
Quarter
Summer '97
Fall '97
Winter '98
Spring '98
Summer '98
Fall '98
MBA
Surface Water
N
239
250
240
262
250
229
Mean
(^g/L)
0.29
0.23
0.25
0.17
0.28
0.32
STD
(^g/L)
0.73
0.55
1.74
0.53
0.83
1.03
Ground Water
N
92
86
101
105
104
93
Mean
(^g/L)
0.21
0.21
0.18
0.14
0.09
0.13
STD
(^g/L)
0.79
0.60
0.53
0.46
0.30
0.51
N:
Number of samples
STD: Standard deviation
Sample Quarter:
Summer '97: July, August, and September
Fall '97: October, November, and December
Winter '98: January, February, and March
Spring '98: April, May, and June
Summer '98: July, August, and September
Fall '98: October, November, and December
EPA/OW/OST/HECD
IV-40
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-12. BCA by Sample Quarter
(Quarterly Distribution System Average)
Sample
Quarter
Summer '97
Fall '97
Winter '98
Spring '98
Summer '98
Fall '98
BCA
Surface Water
N
236
252
243
260
251
232
Mean
(^g/L)
4.12
3.52
3.11
3.54
3.90
3.46
STD
(^g/L)
3.83
2.82
2.54
2.90
3.42
2.76
Ground Water
N
92
88
103
104
103
94
Mean
(^g/L)
1.60
1.40
1.48
1.33
1.55
1.46
STD
(^g/L)
2.54
2.09
2.02
1.99
2.11
2.17
N:
Number of samples
STD: Standard deviation
Sample Quarter:
Summer '97: July, August, and September
Fall '97: October, November, and December
Winter '98: January, February, and March
Spring '98: April, May, and June
Summer '98: July, August, and September
Fall '98: October, November, and December
EPA/OW/OST/HECD
IV-41
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table IV-13. DBA by Sample Quarter
(Quarterly Distribution System Average)
Sample
Quarter
Summer '97
Fall '97
Winter '98
Spring '98
Summer '98
Fall '98
DBA
Surface Water
N
241
252
244
262
253
232
Mean
(^g/L)
1.30
1.12
0.94
0.88
1.17
1.14
STD
(^g/L)
2.19
2.08
2.04
1.86
2.16
2.16
Ground Water
N
92
88
103
104
104
93
Mean
(^g/L)
0.89
0.88
0.81
0.67
0.80
0.88
STD
(^g/L)
1.92
1.39
1.33
1.30
1.53
1.41
N: Number of samples
STD: Standard deviation
Sample Quarter:
Summer '97: July, August, and September
Fall '97: October, November, and December
Winter '98: January, February, and March
Spring '98: April, May, and June
Summer '98: July, August, and September
Fall '98: October, November, and December
EPA/OW/OST/HECD
IV-42
-------
Drinking Water Criteria Document for Brominated Acetic Acids
An examination of the data using the Student's t-test showed that, based on only two
seasons of analysis, the mean concentrtions of MBA in surface water were significantly higher (p
= 0.05) in the summer than in the spring (Table IV-11). Also, based on only two seasons of
analysis, the mean concentrations of BCA in surface water were higher in summer than in winter.
Aside from these exceptions, there were no consistently significant differences (p = 0.05) in the
mean concentrations of BCA or DBA between one season and other in surface water (Table IV-
13). This is in apparent contrast to the findings of Krasner et al. (1989), who found that DBA
levels increased in the warmer months in the utility with the highest bromide levels. Seasonal
variations in brominated acetic acids may be dependent on seasonal fluctuations in bromide-ion
concentration, which were not evaluated in this analysis. No seasonal differences in mean MBA,
BCA, or DBA concentrations in groundwater could be discerned.
B. Exposure to Sources Other Than Drinking Water
MBA has been used in industry and in hospitals. Between 1981 to 1983, The National
Institute of Occupational Safety (NIOSH) conducted a survey of a sample of 4490 businesses
employing nearly 1,800,000 workers (NIOSH, 1990). Potential exposure estimates included
surveyor observations of the use of MBA and trade-name products known to contain MBA.
EPA/OW/OST/HECD IV-43
-------
Drinking Water Criteria Document for Brominated Acetic Acids
During the period from 1981 to 1983, 4874 workers were potentially exposed to MBA. The
largest numbers of exposures (1999) occurred in the commercial printing letterpress business.
Used-car dealers made up the next largest number of potential exposures (1723). The remainder
of potential exposures, in decreasing numbers, included workers in the production of plastics
(402), in hospitals (318), and in individuals working with electron tubes (192). Exposure levels
were not reported in this survey, and more-recent information on numbers of workers exposed
was not available.
During the period from 1981 to 1983, there were no reported survey observations of the
use of BCA, DBA, or trade-name products known to contain BCA or DBA in the workplace
(NIOSH, 1990).
No data were located on exposure to MBA, BCA, or DBA in food, air, or via dermal
exposure when showering or swimming.
C. Overall Exposure
EPA/OW/OST/HECD IV-44
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Only limited data on exposure to MBA, BCA, and DBA in sources other than drinking
water exposure was located. Exposure to drinking water is discussed in Section IV.A.
D. Body Burden
No data could be located on body burden. However, as discussed in Chapter 3, the
brominated acetic acids are metabolized rapidly and are not lipophilic at physiological pH, and so
would not be expected to bioaccumulate.
E. Summary
The ICR database (U.S. EPA, 2000a) contains extensive information on concentrations of
MBA, BCA, and DBA in drinking-water systems, and on how those concentrations vary with
input-water characteristics and treatment methods. The database contains information from six
EPA/OW/OST/HECD IV-45
-------
Drinking Water Criteria Document for Brominated Acetic Acids
quarterly samples from 7/97 to 12/98, from approximately 300 large systems covering
approximately 500 plants. The mean concentrations of BCA were 1.47 and 3.61 |-ig/L from
groundwater and surface water respectively. The mean concentrations of DBA were 0.82 and
1.09 |-ig/L in groundwater and surface water, respectively. Examination of the ICR data using the
Student's t-test indicates that the mean concentrations of MBA, BCA, and DBA in surface water
were significantly higher (p = 0.05) than the mean concentrations of these chemicals in
groundwater. In addition, the mean concentrations of BCA were significantly higher ( p = 0.05)
than the mean concentrations of DBA, which were significantly higher ( p = 0.05) than the mean
MBA concentrations in both surface water and groundwater.
Examination of the ICR data using the Student's t-test suggests that, although the
concentrations of MBA in surface water treated with chlorine are similar to those treated with
chlorine followed by chloramine. BCA and DBA concentrations were lower when free chlorine
was used both in the treatment plant and the distribution system. Although ozonation appeared to
significantly reduce the formation of BCA, there were no significant differences in MBA or DBA
concentrations (p = 0.05) with the use of ozone in treating surface water as compared to the
common (non-ozonation) chemical-disinfection processes. In addition there were no significant
differences (p = 0.05) between the two treatments using ozonation in treating surface water for
MBA, BCA and DBA.
EPA/OW/OST/HECD IV-46
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Consistent with the findings of other investigators, and the chemistry of the formation of
bromoacetic acids, a regression analysis of the ICR data indicates that, with the exception of MBA
in surface water, there was a significant correlation (at cc = 0.05) between influent bromide
concentration and the mean concentrations of BCA and DBA in surface water and groundwater.
In addition, for a given influent bromide concentration range, the mean concentrations of BCA
were generally higher (p = 0.05) that the mean concentrations of DBA and MBA in both surface
water and groundwater.
A regression analysis of the ICR data indicates that there was a significant correlation
(cc = 0.05) between influent TOC concentration and the mean concentrations of MBA, BCA, and
DBA in surface water. This is consistent with the formation of brominated acetic acids from the
reaction of humic acid and hypobromous acid, a compound formed by the reaction of bromide
ion with ozone and/or chlorine in the disinfection process. In addition, for a given influent TOC
concentration range in surface water, the mean concentrations of BCA were significantly higher (p
= 0.05) than the mean concentrations of DBA, which were significantly higher (p = 0.05) than the
MBA mean concentrations.
EPA/OW/OST/HECD IV-47
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Examination of the data using the Student's t-test showed that, based on only two seasons
of analysis, the mean concentrations of MBA in surface water were significantly higher ( p = 0.05)
in the summer than in the spring Also, based on only two seasons of analysis, mean
concentrations of BCA in surface water were higher in summer than in winter. Aside from these
exceptions, there were no consistently significant differences (p = 0.05) in the mean
concentrations of MBA, BCA or DBA between one season and another in either surface water or
groundwater. Seasonal variations in brominated acetic acids may be dependent on seasonal
fluctuations in bromide-ion concentration, which were not evaluated in this analysis.
The data on exposure to sources other than drinking water are limited, but MBA has been
used in industry and in hospitals. Between 1981 to 1983, 4874 workers were potentially exposed
to MBA. (NIOSH, 1990). No data were located on exposure to MBA, BCA, or DBA in food, air,
or via dermal exposure.
No data could be located on body burden levels of MBA, BCA, or DBA..
EPA/OW/OST/HECD IV-48
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Chapter V. Health Effects in Animals
The available database for the brominated acetic acids is limited and, therefore, many
toxicity endpoints have not been fully explored. In recognition of this paucity of data, there is a
large body of ongoing work, particularly for BCA and DBA. Preliminary results for many studies
have been reported in published abstracts and are included here to provide a sense of the spectrum
of effects induced by the brominated acetic acids. The full published studies would need to be
evaluated for a complete understanding of the chemicals' effects and determination of the
relevance of the data for quantitative risk-assessment purposes.
A. Short-Term Exposure
Monobromoacetic acid
Linder et al. (1994a) reported on the acute oral toxicity of MBA as part of a study on its
spermatogenic effects. Five male Sprague-Dawley rats per group were given single doses of 100
to 200 mg/kg MBA (specific dose levels not reported) by gavage in water. The LD50 was reported
as 177 mg/kg, with a 95% (fiducial) confidence-limit range of 156 to 226 mg/kg. Observed
clinical symptoms included excess drinking, hypomobility, labored breathing, and diarrhea. No
EPA/OW/OST/HECD V-l
-------
Drinking Water Criteria Document for Brominated Acetic Acids
histopathologic changes were observed in either the epididymal sperm or testes of surviving
animals.
MBA is irritating and corrosive to the human skin and mucous membranes (NIOSH,
2000). The ability of a variety of carboxylic acids to cause skin corrosion was investigated in
support of the development of a multivariate quantitative structure-activity relationship (QSAR)
analysis (Eriksson et al., 1994). Forty-five aliphatic carboxylic acids (including MBA, DBA, and
BCA) were evaluated in the QSAR analysis. Fifteen of the compounds, including MBA,
monochloroacetic acid, and dichloroacetic acid, were tested for cutaneous corrosion on adult
rabbits. In these studies, the test chemical was applied to bare, shaved skin under an occlusive
glass filter for one hour. The study description indicated that a 10 cm by 10 cm area on the trunk
of the rabbit was shaved, but did not specify that this entire area was exposed. The observed and
predicted lowest-observed-effect concentrations (LOECs) for inducing corrosion for MBA were
0.2 M and 0.1 M, respectively.
No short-term toxicity studies for exposure by the inhalation route were identified.
Bromochloroacetic acid
Systemic toxicity was evaluated as part of a reproductive and developmental toxicity-
screening assay for BCA (NTP, 1998). As part of a range-finding study, groups of male and
EPA/OW/OST/HECD V-2
-------
Drinking Water Criteria Document for Brominated Acetic Acids
female Sprague-Dawley rats (6/group) were exposed for 14 days to drinking water containing 0,
30, 100, 300, or 500 ppm BCA. The study authors reported that the estimated average doses
resulting from these treatments were 0, 3, 10, 28, and 41 mg/kg/day. No mortality was observed
and there were no treatment-related differences in body weight, body-weight gain, feed and water
consumption, or clinical observations as compared with controls. The NOAEL for general
toxicity was 41 mg/kg/day. A LOAEL could not be determined.
The results of the 14-day range-finding study were used to select doses for a reproductive
and developmental toxicity-screening assay (NTP, 1998). Sprague-Dawley rats were
administered BCA in their drinking water for various periods during a 35-day study period. Rats
were divided into two groups of males and three groups of females. Group A males (10/group)
were exposed on study days 6 - 35 to doses of 0, 60, 200, or 600 ppm. Group B males were
exposed on study days 6-31 (5/group at 0, 60, or 200 ppm and 8/group at 600 ppm ) and
administered bromodeoxyuridine (BrdU) via subcutaneously-implanted pumps for 3 days prior to
necropsy in order to measure cell proliferation. The female rats were grouped as follows: Group
A rats (peri-conception exposure group; 10/dose at 0, 60, 200, or 600 ppm) were given BCA on
study days 1-34 and were cohabitated with treated males on study days 13-18. Group B rats
(gestational-exposure group; 13/dose) were cohabitated with treated males on study days 1-5 and
exposed to BCA at doses of 0, 60, 200, or 600 ppm on gestation days (GD) 6 to parturition.
Group C (peri-conception exposure group; 5/dose group at 0, 60, 200 ppm and 8/ group at 600
EPA/OW/OST/HECD V-3
-------
Drinking Water Criteria Document for Brominated Acetic Acids
ppm) were exposed to BCA on study days 1 - 30, cohabitated with treated males on study days 13-
18, and administered BrdU via subcutaneously implanted pumps for 3 days prior to necropsy for
cell-proliferation assessment.
The average daily doses for male rats in both Group A and Group B were estimated by the
study authors as equivalent to 0, 5, 15, and 39 mg/kg/day. No mortality, clinical signs of toxicity,
or body-weight changes were observed. Water consumption was decreased by 21% to 34% at the
high dose, presumably due to taste aversion. Clinical chemistry was evaluated in Group A males,
and the following statistically significant changes were observed: a 16% increase in albumin to
globulin ratio in high-dose animals; decreased alanine aminotransferase (ALT) activity in both the
mid-dose (15% decrease) and high-dose (20% decrease) groups; and increased albumin (5%
increase in the low-dose group, no significant change in the mid-dose group, and 9% increase in
the high-dose group). Clinical-chemistry changes of these magnitudes are not generally
considered to be lexicologically significant and the changes in the high-dose group may have been
secondary to dehydration. Although absolute and relative liver weights increased with increasing
dose in both Group A and Group B, the relative liver weight was statistically different from
controls (10% increase) only in the high-dose group, and absolute liver weight was not
significantly elevated. Gross necropsy did not reveal any major changes. No dose-related
increases in individual hepatocyte necrosis (Group A) or labeling index (Group B only) were
observed. Histopathologic examination showed an increase in cytoplasmic vacuolization of
EPA/OW/OST/HECD V-4
-------
Drinking Water Criteria Document for Brominated Acetic Acids
hepatocytes of treated animals in Group A. This effect was observed in all dose groups, was more
prominent in the high-dose group, and was absent in controls. However, cytoplasmic
vacuolization was observed in both control and dosed animals of Group B, and was not increased
with BCA treatment. Although the study authors suggested that the biological significance of
these changes could not be determined without evaluation following longer-term exposure, the
highest dose was considered sufficient to induce general toxicity under the conditions of this
study. In light of the questions concerning the biological significance of the clinical chemistry,
liver weight and histopathology changes, 39 mg/kg/day (the highest dose tested) was considered
to be a marginal LOAEL in males, and the corresponding NOAEL was 15 mg/kg/day.
The estimated average daily doses for Group A and Group C females (peri-conception
exposure groups) were 0, 6, 19, and 50 mg/kg/day. No mortality, clinical signs of toxicity, or
body-weight changes were observed in the female rats of these groups, but water consumption
was decreased by 24-34% at the high dose; these findings were similar to those observed in
treated male rats. Decreased water consumption was the only effect observed in Group A
females. However, Group C females exhibited a dose-related increase in the incidence of renal
tubular dilatation/degeneration (0/5 in controls, 2/5 in the 19 mg/kg/day group, and 3/8 in the 50
mg/kg/day group). A statistical analysis of these data indicated that the incidences in treated
groups were not statistically different from those in the controls (p>0.05 using the Fisher exact
test); however, sample sizes were small and, thus, the power of the analysis was limited. No
EPA/OW/OST/HECD V-5
-------
Drinking Water Criteria Document for Brominated Acetic Acids
detailed pathology report was available to evaluate the biological significance of these
histopathological changes in the kidneys. The study authors concluded that only the high dose
resulted in renal toxicity. Cell proliferation analysis showed small differences between the
labeling index of the treated and control groups for the kidney in Group B males and for the liver
and urinary bladder in Group C females. However, these changes were not dose-dependent and
were not considered to be biologically significant.
Estimated average daily doses of BCA for Group B females (gestational exposure group)
were 0, 10, 25, and 61 mg/kg/day. Similar to other groups, the only consistent treatment-related
effect in Group B females was decreased water consumption at the high dose.
Overall, some histopathology findings in this study suggest that the liver (as evidenced by
increased relative weight and increased cytoplasmic vacuolization in males) and kidney (as
indicated by increased renal tubular dilatation/degeneration in females) may be target organs for
BCA toxicity. Although liver histopathology was only observed in high-dose Group A males,
these effects were considered to be treatment-related by the study authors, yielding a marginal
LOAEL of 39 mg/kg/day for equivocal liver effects. The study authors also concluded that the
histopathological changes observed in the kidneys of Group C females, although not statistically
different from controls, might be indicative of kidney toxicity. However, kidney histopathology
was not corroborated by the kidney-labeling index, suggesting that renal toxicity was not of
EPA/OW/OST/HECD V-6
-------
Drinking Water Criteria Document for Brominated Acetic Acids
sufficient severity to induce cellular proliferation and regeneration. Based on the study authors'
interpretation of the results, 39 mg/kg/day was selected as a LOAEL, and the corresponding
NOAEL was 19 mg/kg/day. Effects on reproductive and developmental endpoints and
determination of critical-effect levels for these systems are described in Section V.C.
Parrish et al. (1996) tested whether the ability of brominated acetic acids to induce
oxidative DNA damage was due to peroxisome proliferation. Male B6C3F1 mice (6/dose group)
were administered drinking water containing 0, 100, 500, or 2000 mg/L BCA for 3 weeks. The
authors did not provide information on average daily doses. Based on a default water-intake value
of 0.25 L/kg/day for male B6C3F1 mice (U.S. EPA, 1988), the corresponding doses were
estimated to be 0, 25, 125, and 500 mg/kg/day. General toxicity was assessed by measuring body
weight and liver weight; these data for BCA are summarized in Table V-l. The effects of BCA on
oxidative DNA damage and peroxisome proliferation were measured in the livers of male
B6C3F1 mice, using the following measures: (1) changes in the DNA adduct 8-hydroxy-2-
deoxyguanosine (8-OhdG) as an indicator of oxidative stress, and(2) changes in levels of cyanide
insensitive Acyl-CoA oxidase and 12-hydroxylation of lauric acid as indicators of peroxisome
proliferation. An additional dose group exposed to 3000 mg/L BCA (750 mg/kg/day) was
evaluated for the Acyl-CoA activity. Body weight was decreased by 8.5% at the highest dose
tested. Absolute and relative liver weights were increased at the high dose by 20% and 33%,
respectively. BCA had no effect on either measure of peroxisome proliferation after exposures up
EPA/OW/OST/HECD V-7
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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).
The absence of other measures of liver toxicity, such as histopathology or clinical
chemistry results, clouds the classification of these liver-weight changes as adverse. However, the
accompanying increase in oxidative DNA damage suggests potentially adverse liver effects.
Based on the data presented, the LOAEL for minimal liver effects was 500 mg/kg/day and the
NOAEL was 125 mg/kg/day.
EPA/OW/OST/HECD
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table V-l. Body and Liver Weight Changes Induced by BCA and DBA"
BCA
Drinking water (g/L)
Control
0.1(25mg/kg/day)b
0.5(125mg/kg/day)
2.0(500mg/kg/day)
Body weight (g)
27.2 ±0.5°
26.1 ±0.3
28.3 ±0.7
24.9±0.7**d
Liver weight (g)
1.5 ±0.01
1.5±0.1
1.8 ±0.1
1.8±0.1*
Relative liver weight
(% body weight)
5.4% ±0.1
5.8 % ±0.4
6.2 % ± 0.6
7.2% ± 0.4**
DBA
Control
0. 1 (25 mg/kg/day)
0.5(125mg/kg/day)
2.0 (500 mg/kg/day)
27.1 ±0.5
24.0 ±0.7**
25.6 ±0.8
26.1 ±0.4
1.5 ±0.01
1.4 ±0.1
2.1 ±0.2**
2.0±0.1**
5.4% ±0.1
5.8% ± 0.4
8.0% ± 0.5**
7.8% ± 0.6
Notes:
a. Adapted from Parrish et al., 1996
b. Estimated daily doses were calculated based on default drinking water values of 0.25
c. Mean ± standard error
* Statistical significance: p<0.05
** Statistical significance: p<0.01
The ability of a variety of carboxylic acids to cause skin corrosion was investigated using
multivariate quantitative structure-activity relationship QSAR (Eriksson et al., 1994). The
predicted lowest-observed-effect concentration (LOEC) for corrosion for BCA was 0.7 M.
EPA/OW/OST/HECD
V-9
-------
Drinking Water Criteria Document for Brominated Acetic Acids
No short-term toxicity studies for exposure to BCA by the inhalation route were identified.
Dibromoacetic acid
The acute toxicity of DBA was examined as part of a study on the spermatogenic effects of this
compound (Linder et al., 1994a). Male Sprague-Dawley rats (5/group) were given single doses of
1000 to 2000 mg/kg DBA (specific dose levels not reported) by oral gavage in water. Custom-
synthesized DBA of >99% purity was used because the commercial chemical is generally
contaminated with approximately 10% MBA. Surviving animals were killed 14-21 days after
dosing. The oral LD50 was 1737 mg/kg, with a 95% (fiducial) confidence-limit range of 1411 -
1952 mg/kg. Most of the animal deaths occurred within 48 hours of dosing. Observed symptoms
included excess drinking, hypomobility, labored breathing, diarrhea, and ataxia. Histopathologic
examination of the epididymal sperm in surviving animals showed the presence of mis-shapen
and degenerating sperm, as well as abnormal retention of Step 19 spermatids. Effects other than
spermatotoxicity were not examined.
In another single-dose spermatotoxicity study, Vetter et al. (1998) treated sexually-mature male
Crl:CD(SD)BR rats (4-5/dose group) with 0, 600, or 1200 mg/kg DBA in 10 mL/kg deionized
water. In the high-dose group, signs of overt toxicity included lethargy, irregular gait, decreased
feces, ocular discharge, and dyspnea. Abnormal respiratory sounds were observed in some
EPA/OW/OST/HECD V-10
-------
Drinking Water Criteria Document for Brominated Acetic Acids
animals (number of animals affected was not specified) and one animal in the high-dose group
died on study day 3. No overt toxicity was observed in the low-dose group. No changes in
measured sperm parameters (motility, morphology, and cell-membrane permeability) were
reported at either dose; however, mi testes histopathology (the presence of basophilic bodies) was
observed in both dose groups. Based on the clinical findings, 1200 mg/kg was considered to be
an acute frank effects level (PEL). The LOAEL was 600 mg/kg for testes histopathology. and a
NOAEL could not be determined.
Parrish et al. (1996) tested whether the ability of brominated acetic acids to induce oxidative
DNA damage was due to peroxisome proliferation. Male B6C3F1 mice (6/dose group) were given
drinking water containing 0, 100, 500, or 2000 mg/L DBA for 3 weeks. The authors did not
provide information on average daily doses. However, based on a default water-intake value of
0.25 L/kg/day for male B6C3F1 mice (U.S. EPA, 1988), the corresponding daily doses were
estimated to be 0, 25, 125, and 500 mg/kg/day, respectively. The effects of DBA on oxidative
DNA damage and peroxisome proliferation were measured in the livers of male B6C3F1 mice,
using the following measures: (1) changes in the DNA adduct 8-hydroxy-2-deoxyguanosine (8-
OhdG) as an indicator of oxidative stress, and (2) changes in levels of cyanide insensitive Acyl-
CoA oxidase and 12-hydroxylation of lauric acid as indicators of peroxisome proliferation. An
additional dose group exposed to 3000 mg/L DBA (750 mg/kg/day) was evaluated for the Acyl-
CoA activity . As part of this study, general toxicity was assessed by measuring body weight and
EPA/OW/OST/HECD V-11
-------
Drinking Water Criteria Document for Brominated Acetic Acids
liver weight, as summarized in Table V-l. No dose-related decrease in body weight was
observed, but absolute and relative liver weights were increased at the mid- and high-doses. At
the mid-dose (125 mg/kg/day), absolute and relative liver weights were increased by 40% and
48%, respectively. At the high dose (500 mg/kg/day), absolute and relative liver weights were
increased by 33% and 44%, respectively. DBA induced Acyl-CoA activity to a maximum of 3-
fold following exposures up to 3000 mg/L, but did not induce the 12-hydroxylation of lauric acid.
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 absence of a clear dose response, and the lack of other measures of liver
toxicity, such as histopathology or clinical chemistry, clouds the classification of the liver-weight
changes as adverse. However, the magnitude of the change and the accompanying increase in
oxidative DNA damage suggests potentially-adverse liver effects. Based on the data presented,
the LOAEL for minimal liver effects was 125 mg/kg/day. The NOAEL was 25 mg/kg/day.
As part of a male reproductive study, Linder et al. (1995) administered daily gavage doses of 0
or 250 mg/kg DBA to male Sprague-Dawley rats (10/dose group). Dosing was terminated after
42 days because severe toxic effects, including labored breathing, light tremor, difficulty moving
the hind limbs, and significant weight loss, developed. In a subsequent study, male rats (10/dose
group) were given 0, 2, 10, or 50 mg/kg/day DBA by oral gavage for 31 or 79 days. The only
observed effect was a slight decrease in body weight in the 50 mg/kg/day group on day 79 to
EPA/OW/OST/HECD V-l 2
-------
Drinking Water Criteria Document for Brominated Acetic Acids
approximately 95% of controls. Based on the findings of both studies, the PEL for general
toxicity was 250 mg/kg/day and the NOAEL was 50 mg/kg/day. Adverse-effect levels for
reproductive endpoints are described in Section V.C.
NTP (1999) evaluated the immunotoxicity of DBA in female B6C3F1 mice (8/dose group)
exposed to drinking water containing 0, 125, 250, 500, 1000, or 2000 mg/L DBA for 28 days.
Four separate studies were conducted and different general and immunologic endpoints were
examined in each study. Studies 1-3 investigated selected immunologic endpoints and body-
weight changes; body and organ weights, hematology, and gross pathology were examined in
Study 4. Key immunotoxicity responses are presented in Table V-2, and body and organ weights
are summarized in Table V-3. The study authors did not estimate DBA daily doses resulting from
drinking-water exposures; however, the average daily doses could be calculated based on water-
consumption and body-weight data provided in the study report. DBA dose ranges were similar
across the four studies and are presented below in conjunction with the experimental findings for
each study. No significant differences (p<0.05) in drinking-water consumption among dosed
groups and no clinical signs of overt toxicity were observed in any of the studies.
EPA/OW/OST/HECD V-13
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table V-2. Immunotoxicity of DBA in Female B6C3F1 micea
Study Number/Endpoint
Study #1
Spleen cell number x 107
Spleen Macrophages
(% of cells)
Natural Killer cell lytic activity
(LU/107 cells)'
Mixed leukocyte response
(CPM/105 spleen cells -
responders) e
Study #2
IgM antibody-forming colonies
(AFC/106 spleen cells)
Serum IgM liter (OD)g
Study #4
Macrophage Activating Factor
(% suppression without
stimulation)
(% suppression with stimulation)
Dose (mg/kg/day)b
0
15.97±0.50C
2.8± 0.2
12± 1
633 ± 69
0
1958 ± 145
103 ±7
0
29. 08 ±2.77
86. 80 ±4.26
19
16.69±
0.51
3.0 ± 0.2
14 ± 2
834 ± 160
20
1767 ± 62
100 ± 9
14
19.61 ±
5.80
92.02±
2.04
39
17.55 ± 1.02
2.9 ± 0.2
15± 1
863 ± 126
38
1590 ± 149
131 ± 19
33
16.45± 6.39
94.17±2.16
73
19.13±
0.58d**
3.5 ± 0.3
16± 1*
781 ±52
70
1333±89**
102 ± 12
68
10.22±
5.86*
92.01 ±4.01
150
18.49±
0.46*
4.2 ± 0.1**
20 ± 1**
693 ±50
143
1251
±120**
97 ±5
132
15.39±
2.50
93.62±
2.84
285
18.45±
0.39*
4.5 ± 0.2**
24 ± 2**
851 ±69
280
985 ±69**
85 ± 5
236
28. 58 ±3. 62
89.05 ±2.73
a. Adapted from NTP, 1999
b. Doses were estimated based on drinking-water concentrations and water-intake and body-weight data provided in
EPA/OW/OST/HECD
V-14
-------
Drinking Water Criteria Document for Brominated Acetic Acids
c. Lytic unit (LU) = the number of splenocytes required to kill 10% of the target cells.
d. CPM = counts per minute based on 3H-thymidine incorporation in responder cells.
e. IgM liter based on enzyme-linked immunosorbant assay.
* Statistical significance: p<0.05
** Statistical significance: p<0.01
Table V-3. General toxicity of DBA in Female B6C3F1 mice"
Body
weight (g)
Liver weight
(mg)
(% of body
weight)
Kidney
weight (mg)
(% of body
weight)
Spleen
weight (mg)
(% body
weight)
Thymus
weight (mg)
(% body
weight)
Reticulocyte
count (%)
Estimated Dose (mg/kg/day)b
0
23.7 ±0.4
1071 ±25
4.5± 0.1
287 ± 15
1.21 ±0.07
77 ± 3
0.328 ±0.015
67 ±4
0.280 ±0.014
4.01 ±0.11
14
24.0 ±0.4
1183 ± 32*
4.9±0.1*
298 ±6
1.24 ±0.01
84 ±3
0.349 ±0.013
60 ±4
0.251 ±0.017
4. 24 ±0.27
33
24.7 ±0.6
1293 ±41**
5.2 ± 0.1**
305 ± 5*
1.23 ±0.02
86 ±4
0.350 ±0.018
65 ± 3
0.265 ±0.010
4. 70 ±0.49
68
24.8 ±0.7
1386 ±62**
5.6 ± 0.1
312 ±9*
1.26 ±0.02
92 ±4*
0.370 ±0.015
63 ±4
0.252 ±0.011
4.61 ±0.16*
132
24.6 ±0.4
1479 ± 14**
6.0 ± 0.1**
317±6 *
1.29±0.01*
88 ± 3
0.359±0.011
58 ±2
0.236 ±0.007
4. 50 ±0.16*
236
23.0 ±0.5
1567 ±41**
6.8 ± 0.1**
335 ± 6*
1.46±0.01a*
96 ± 4**
0.418±0.014**
43 ± 3**
0.189±0.013**
5.21 ±0.25**
Notes:
a. Adapted from NTP, 1999
EPA/OW/OST/HECD
V-15
-------
Drinking Water Criteria Document for Brominated Acetic Acids
b. Doses were estimated based on drinking-water concentrations and water-intake and body-weight data
provided in the report.
c. Mean ± Standard Error.
* Statistical significance: p<0.05
** Statistical significance: p<0.01
EPA/OW/OST/HECD V-16
-------
Drinking Water Criteria Document for Brominated Acetic Acids
In Study 1, doses estimated using group-specific body weights and water-consumption
rates were 0, 19, 39, 73, 150, and 285 mg/kg/day. A statistically significant 8% decrease in
terminal body weight was observed in the high-dose group. Spleen-cell number was significantly
increased above controls (p<0.05) at 73 mg/kg/day and higher, but there was no dose-response.
At 73 mg/kg/day, spleen-cell number was elevated approximately 20% above controls, whereas at
both 150 and 285 mg/kg/day, the increase was approximately 12%. With the notable exception of
macrophages, the increase in the absolute number of most spleen-cell types paralleled the increase
in total number of spleen cells; thus the percentage of each cell type was generally unaltered. In
contrast, spleen macrophages increased in a dose-dependent manner to 50%, 77%, and 91% above
controls at 73, 150, and 285 mg/kg/day, respectively, (500, 1000, and 2000 mg/L, respectively)
An increase in natural killer (NK) cell lytic activity was also observed at the three highest DBA
dose groups (p<0.05) when expressed as specific activity; a significant increase was observed in
the four highest dose groups (i.e., 39 mg/kg/day and above) when expressed as total-spleen
activity. NK-cell activity was maximal at the highest dose, showing an increase of 100% when
measured as specific activity and of 143% when measured as total-spleen activity. However, the
significance of the treatment-related changes in NK-cell activity is unclear, because the positive
control used in the experiment produced a significant decrease in NK-cell activity. DBA
treatment had no effect on mixed-leukocyte response, which measures proliferative response of
splenic leukocytes from treated animals to allogenic lymphocytes (i.e., lymphocytic cells from a
genetically distinct strain of the same species) of DBA/2 mice. Overall, the results of Study 1
EPA/OW/OST/HECD V-17
-------
Drinking Water Criteria Document for Brominated Acetic Acids
demonstrated an increase in several measures of cellular immunity, with statistically significant
effects generally occurring at doses of 73 mg/kg/day and higher. The toxicological significance of
the findings of Study 1 are unclear and are discussed in more detail later in this section, using a
weight-of-evidence based on the results of all four studies.
In Study 2, estimated average daily doses were 0, 20, 38, 70, 143, and 280 mg/kg/day. No
significant effects of DBA treatment on body weight were observed. In contrast to Study 1, no
increase in spleen-cell number was observed. A statistically significant (p<0.05) dose-dependent
decrease in spleen IgM antibody-forming cell response to sheep erythrocytes was observed
beginning > 70 mg/kg/day. There was no change, however, in serum-IgM titer to sheep
erythrocytes. The study authors noted that the lack of concordance between these two measures of
humoral response was not uncommon. They suggested that discordance might arise from the fact
that the IgM antibody-forming cell response is a specific measure of immune response in the
spleen, whereas the IgM titer measures systemic humoral immunity and, thus, would reflect
changes in both bone-marrow and lymph-node antibody production, in addition to antibody
production in the spleen. Therefore, the IgM assay might not be a sensitive measure of toxicity
for substances whose target organ of toxicity is only the spleen.
Study 3 evaluated macrophage activation. The estimated average daily doses were 0, 16,
35, 69, 134, and 229 mg/kg/day. No significant effects on body weight were observed. To assess
EPA/OW/OST/HECD V-18
-------
Drinking Water Criteria Document for Brominated Acetic Acids
macrophage activation, peritoneal macrophages were stimulated by treatment with a combination
of gamma interferon and lipopolysaccharide, and their ability to kill or inhibit the growth of
B16F10 tumor cells was measured. No clear dose-dependent effects of DBA treatment on
B16F10 tumor-cell growth were observed with or without macrophage activation.
Body weight, organ weight, gross pathology, and hematology were evaluated in Study 4
(Table V-3). The estimated average daily doses were 0, 14, 33, 68, 132, and 236 mg/kg/day.
Based on pooled body-weight data for all studies, no significant change in terminal body weight
was observed. However, body-weight gain was decreased by 40% at the high dose. Statistically
significant changes in organ weight were also reported. Thymus weight was significantly
decreased only at the high dose. Absolute spleen weight was elevated at all the doses, but was
statistically different from controls (p<0.05) only at 68 mg/kg/day and 236 mg/kg/day (19% and
24% increase, respectively, as compared with controls). Although absolute spleen weight was
also elevated at 132 mg/kg/day (14% increase relative to controls), this difference was not
statistically significant, indicating an ambiguous dose-response. Relative spleen weight was
statistically significantly increased only at the high dose, although there was an increasing trend
with increasing dose. The absolute and relative liver-weight increases were dose-dependent and
were significantly elevated at all doses tested (> 14/mg/kg/day). Relative liver weight increased
with increasing dose to 9%, 16%, 24%, 33%, and 51% above control values for each of the dose
groups. Kidney weight was also statistically increased in a dose-dependent manner, beginning at
EPA/OW/OST/HECD V-19
-------
Drinking Water Criteria Document for Brominated Acetic Acids
33 mg/kg/day for absolute weight (6% increase) and 132 mg/kg/day for relative kidney weight
(7% increase). With the exception of a dose-related increase in reticulocytes that achieved
statistical significance at >68 mg/kg/day, no biologically significant hematology parameters were
affected by treatment. It should be noted, however, that increases in reticulocytes are generally
not considered adverse. No gross pathological lesions were identified. Although a
histopathologic examination was conducted for Study 4, no results were provided. Therefore,
treatment histopathology could not be assessed.
Overall, exposure to DBA in drinking water for 28 days resulted in body- and organ-
weight changes and alterations in several indicators of immunologic response. General toxicity
indicators included decreased body-weight gain in the highest-dose group tested in all four
studies, increased liver weights (both absolute and relative) at all doses tested > 14 mg/kg/day),
and increased absolute (33 mg/kg/day and above) and relative (132 mg/kg/day and above) kidney
weights. However, the absence of supporting clinical chemistry and/or histopathologic data
precludes determination of liver and kidney effects as adverse. Therefore, the observed increase
in liver weight was not selected as the critical effect. In Study 4, thymus weights were decreased
and spleen weights (absolute and relative) were increased at the highest dose tested (236
mg/kg/day). Spleen weight was also increased in a dose-dependent manner at lower doses;
however, the dose-response was not clear cut. Further, spleen weights were not increased in
Study 2 at similar doses. Therefore, the toxicologic significance of this finding remains unclear.
EPA/OW/OST/HECD V-20
-------
Drinking Water Criteria Document for Brominated Acetic Acids
A number of measures of cellular and humoral immunity were altered in a dose-dependent
manner by DBA treatment, beginning in the animals treated with 500 mg/L DBA in drinking
water (equivalent to 68-73 mg/kg/day, depending on the study). Spleen-cell number was
increased in Study 1 but was not elevated in Study 2, limiting interpretation of these findings. As
previously discussed, NK-cell lytic activity (expressed as either specific activity or total spleen
activity) was increased in DBA-treated animals in Study 2, but was decreased by exposure to a
positive control in the same experiment. Therefore, these findings are inconclusive. The number
of spleen macrophages, however, increased statistically in a dose-dependent manner a >500 mg/L
(73 mg/kg/day) DBA, indicating an immunotoxic response in this target organ. In Study 3,
exposure to 500 mg/L (70 mg/kg/day) DBA and above also decreased spleen IgM antibody-
forming cell response, which represents a depression in humoral immunity. The decrease in
spleen IgM antibody-forming cell response and the increase in spleen macrophages are the
clearest indicators of an immunotoxic effect. Based on these findings, the LOAEL for
immunotoxicity under the conditions of this study was 70 mg/kg/day, and the NOAEL was 38
mg/kg/day.
No short-term toxicity studies for exposure to DBA by the inhalation or dermal route were
identified.
B. Long-Term Exposure
EPA/OW/OST/HECD V-21
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Monobromoacetic acid
No long-term toxicity studies for any exposure route were identified.
Bromochloroacetic acid
In a published abstract, Stauber et al. (1995) reported on preliminary data suggesting that
BCA induces liver tumors in B6C3F1 mice. Noncancer liver effects included glycogen
accumulation and hepatocyte vacuolization. No long-term toxicity studies for any exposure route
were identified in the peer-reviewed literature. However, BCA is currently undergoing 90-day
subchronic and 2-year chronic bioassays (NTP, 2000a).
Dibromoacetic acid
In published abstracts, So and Bull (1995) reported that DBA increased the formation of
aberrant crypt foci in the colon of treated rats, and Stauber et al. (1995) reported on preliminary
data suggesting that DBA induces liver tumors in B6C3F1 mice. The effects of longer-term DBA
exposure on noncancer endpoints were not described.
EPA/OW/OST/HECD V-22
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Moser et al. (2004) examined the neurotoxicity of DBA in adolescent (28-day-old) male
and female F344 rats (12/sex/dose) given DBA in drinking water at concentrations of 0, 200, 600,
or 1500 mg/L (mean doses calculated by the authors as 0, 20, 72, and 161 mg/kg/day) for 6
months. In both sexes, body weight was significantly depressed in the high-dose group but
overall health status was unaltered. A neurobehavioral test battery was administered to all
animals at 1,2, 4, and 6 months. Dose-dependent neuromuscular toxicity, characterized by mild
gait abnormalities, hypotonia, and decreased forelimb and hindlimb grip strength, was observed in
both sexes. Sensorimotor responsiveness, as measured by responses to a tail pinch and click, was
reduced at all doses, but did not progress with continued exposure to DBA. Decreased motor
activity was noted in both sexes in the high-dose group, whereas a chest clasping response was
only observed in high-dose females. Neuropathologic examination revealed significant myelin
sheath degeneration, axonal swelling, and axonal degeneration in the lateral and ventral areas of
the spinal cord white matter in the high-dose group. In the mid- and high-dose groups, small
numbers of swollen, eosinophilic or faintly basophilic, and occasionally vacuolated neurons were
observed in the spinal cord gray matter, and appeared to represent axonal degeneration. No
treatment-related neuropathology was noted in the eyes, peripheral nerves, peripheral ganglia, or
brain. Based on neurobehavioral abnormalities, the LOAEL was 20 mg/kg/day, the lowest dose
tested, and a NOAEL could not be determined. Additionally, the doses used in this studies were
not well quantified because the dosages decreased greatly over the exposure period. However, the
EPA/OW/OST/HECD V-23
-------
Drinking Water Criteria Document for Brominated Acetic Acids
results of this study suggest that neurotoxicity should be considered in the overall hazard
evaluation of haloacetic acids
No long-term systemic toxicity studies for any exposure route were identified in the peer-
reviewed literature. However, DBA is currently undergoing 90-day subchronic and 2-year chronic
bioassays (NTP, 2000b).
C. Reproductive and Developmental Effects
Much of the emphasis on reproductive toxicity of brominated acetic acids has focused on
the potential spermatotoxicity of these compounds. Therefore, to enhance the reader's evaluation
of the following study descriptions, a short summary of spermatogenesis relevant to assessing
male reproductive toxicity is provided here. For additional information the reader is referred to
Zenick et al. (1994), from which the following summary text was largely developed.
The development of mature sperm (spermatogenesis) is a multiple-step process that begins
within the seminiferous tubules in the testes and is completed with the movement of spermatids
through the caput, corpus, and cauda epididymis for further functional development and transport
to the vas deferens. The seminiferous tubule is comprised of spermatogenic cells and support
cells such as Sertoli cells. The spermatogenic cells undergo a well-defined step-wise maturation
EPA/OW/OST/HECD V-24
-------
Drinking Water Criteria Document for Brominated Acetic Acids
process. As the cells mature, they move from the basal membrane of the seminiferous tubule until
eventual release from the supporting Sertoli cells into the tubule lumen. The release of the
spermatids from the Sertoli cells is termed spermiation. The resulting sperm cells are transported
from the seminiferous tubule lumen to the epididymis where they undergo further functional
development, including the acquisition of motility and reproductive viability.
The maturation of the spermatogenic cells in the seminiferous tubules occurs through a
series of phases and the increasingly-mature spermatogenic cells sequentially develop into
spermatogonia, spermatocytes, and spermatids. Each of these three major developmental phases
includes a series of smaller developmental steps. For example, for the rat there are 19
developmental phases or "steps" for spermatids. Within the seminiferous tubule, spermatogenic
cells in various steps of development are found in distinct and repeatable associations. Each
common set of associations is called a Stage of the seminiferous epithelium. For example, a
cross-section of a rat seminiferous tubule at Stage VIE would typically contain PI and P
spermatocytes, and Step 8 and 19 spermatids. Thus, perturbations in normal cell associations can
serve as an indication of spermatotoxicity. The time period between the appearance of the same
Stage at a given point in the epithelium is called the cycle length of the seminiferous epithelium.
The stages and cycle length vary across species, but are nearly constant in the same species. The
consistent length of time for spermatogenesis can be useful for identifying targets for
spermatotoxicity, particularly for single-dose studies. For example, information about potential
EPA/OW/OST/HECD V-25
-------
Drinking Water Criteria Document for Brominated Acetic Acids
targets of toxicity can often be gained by determining the amount of time from the time of
exposure to a toxicant to the appearance of adverse effects by tracing back to the phase of
development that the affected sperm was in at the time of exposure.
Spermatogenesis can also be perturbed through toxicity directed at cell populations that
aid in the maturation of the spermatogenic cells. Sertoli cells provide support functions and
developmental regulation of sperm cells, and thus can be an important target for toxicity. Sertoli
cells play important roles in endocrine regulation of spermatogenesis, provide a protective semi-
permeable barrier for the seminiferous tubules, and provide direct support for development of
spermatogenic cells through phagocytosis. For example, the Sertoli cells phagocytize a portion of
the cytoplasm and overlying membrane of the spermatid to form a residual body at spermiation.
Monobromoacetic acid
Linder et al. (1994a) reported the results of acute-toxicity and acute-spermatotoxicity
studies of MBA. In the spermatotoxicity study, male Sprague-Dawley rats (8/group) were given a
single dose of either 0 or 100 mg/kg MBA in a volume of 5 mL/kg in water, and were sacrificed 2
or 14 days after dosing. The selected single dose of 100 mg/kg was an approximate LD01, and was
chosen to provide a relatively-high dose with a minimal likelihood of mortality. Measures of
male reproductive toxicity included reproductive-organ weights, sperm counts, sperm
EPA/OW/OST/HECD V-26
-------
Drinking Water Criteria Document for Brominated Acetic Acids
morphology, sperm motility, and histopathological examination of the seminiferous tubules. No
adverse effects were observed in the single-dose study; therefore, a repeated-dosing protocol
experiment was also conducted. Groups of eight rats were given daily doses of 0 or 25 mg/kg/day
MBA in water for 14 days, and were sacrificed 24 hours after the last dose. MBA also failed to
induce any spermatotoxicity in this repeated-do sing study.
In a published abstract, Randall et al. (1991) reported on the reproductive and
developmental toxicity of MBA. Pregnant Long-Evans rats were given oral gavage doses of 0,
25, 50, or 100 mg/kg/day MBA in distilled water on gestation days 6-15. In the high-dose group,
maternal weight-gain was reduced and one dam died. No effects on reproduction were observed.
Several developmental effects were noted in the high-dose group, including decreased size of live
fetuses (the affected measure of size was not provided in the study summary) and increased
incidence of soft-tissue malformations, most of which were cardiovascular and craniofacial.
Based on the limited data provided in the abstract, the LOAEL for both maternal and
developmental effects is 100 mg/kg/day and the NOAEL is 50 mg/kg/day.
No reproductive or developmental toxicity for MBA was identified following dosing by
the inhalation or dermal routes.
Bromochloroacetic acid
EPA/OW/OST/HECD V-27
-------
Drinking Water Criteria Document for Brominated Acetic Acids
NTP (1998) reported the results of a short-term reproductive and developmental toxicity-
screening protocol for BCA. Details of the protocol for this study are provided in Section V. A.
Briefly, male and female Sprague-Dawley rats were administered 0, 60, 200, or 600 ppm BCA in
their drinking water for various periods during a 35-day study period. The rats were divided into
two groups of males and three groups of females. Group A males (10/group) were exposed on
study days 6-35. Group B males were exposed on study days 6-31 to 0, 60, or 200 ppm (5/group)
or to 600 ppm (8/group), and were subsequently treated with BrdU for 3 days prior to necropsy to
evaluate cell proliferation. The study authors reported that the estimated average doses for males
were 0, 5, 15, and 39 mg/kg/day. Male rats were evaluated for clinical pathology, organ weights,
sperm analysis (group A only), and histopathology. No consistent treatment-related effects on
epididymal sperm measures, spermatid head counts, sperm morphology, or sperm motility were
observed at necropsy.
Among females, Group A (10/dose group) rats were treated with BCA on study days 1-34
and cohabitated with treated males on study days 13-18. Group B (13/dose group) females were
cohabitated with treated males on study days 1-5 and exposed on GD 6 through parturition.
Group C females (5/group at 0, 60, 200 ppm, and 8 animals at 600 ppm) were exposed to a dosing
regime similar to that of Group A, but were removed from treatment on study day 30 and
subsequently administered BrdU to assess target-tissue cell proliferation. Thus, the treatment
protocol for Group A resulted in exposure for 12 days prior to mating and from GD 1-16 or 1-21,
EPA/OW/OST/HECD V-28
-------
Drinking Water Criteria Document for Brominated Acetic Acids
depending on the number of days of cohabitation required for mating. The treatment protocol for
Group C females resulted in 12 days of premating exposure and exposure beginning on GDI and
continuing through GD 12-16, depending on the number of days required for mating. Both groups
were evaluated for indices of mating and fertility and number of corpora lutea, live and dead
fetuses, and implantation sites. The study authors reported that the estimated average daily doses
resulting for both groups were 0, 6, 19, and 50 mg/kg/day. No effects were observed on the
mating index (number of females with vaginal sperm / number of cohabitating pairs), pregnancy
index (number of fertile pairs / number of cohabitating pairs) or fertility index (number of fertile
pairs / number of females with vaginal sperm). Due to the limited number of pregnancies
evaluated and the similar dosing protocols, reproductive-outcome data were pooled for Groups A
and C females. Analysis of the combined results revealed statistically significant decreases of up
to 70% in the number of live fetuses per litter and up to 75% in total implants per litter, as
compared with controls. Pre-implantation losses increased up to 249% of controls in the
combined high-dose group, but this result was not statistically significant. A summary of
selected endpoints for the combined Group A and C female data is provided in Table V-4.
Statistically significant treatment-related effects by individual groups included a 16%
decrease in total implants per litter in the 600 ppm Group A females and a 50% decrease in
number of live fetuses per litter in 600 ppm Group C females. A number of other outcomes for
either Group A or C were reported to be adversely altered by BCA treatment but did not differ
EPA/OW/OST/HECD V-29
-------
Drinking Water Criteria Document for Brominated Acetic Acids
statistically from controls: (1) post-implantation losses were increased in the 600 ppm Group C
females; (2) pre-implantation losses were increased in the 600 ppm Group A females and all
dosed groups in Group C; (3) an increase in total resorptions was observed in the 600 ppm Group
C females; and (4) decreased total implants per litter occurred in all dosed groups in Group C.
The study author noted that the reason that many of these adverse outcomes lacked statistical
significance may have been due to the small number of pregnancies (N = 2 to 5) per treatment
group evaluated in this screening protocol.
EPA/OW/OST/HECD V-30
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table V-4. Reproductive and Developmental Toxicity of BCA following Peri-conception
Exposure (Combined Data for Female Groups A and C)a
Parameter
Live fetuses per litter
(% of controls)
Total implants per litter0
(% of controls)
% Pre-implantation lossd
(% of controls)
%Post-implantation loss6
(% of controls)
Total resorptionsf
(% of controls)
Dead fetuses per litter
(Number of pregnant females)
Estimated Dose (mg/kg/day)
0
14.9 ± 1.05b
(100%)
16.4 ± 1.18
(100%)
12. 92 ±6.24
(100%)
8.67± 1.62
(100%)
1.5 ±0.27
(100%)
0.0 ±0.00
(10)
6
12. 2 ± 1.36
(82%)
13.7 ± 1.41
(84%)
18. 52 ±6.81
(143%)
11.50± 3.10
(133%)
1.5 ±0.41
(100%)
0.0 ±0.00 (11)
19
13. 2 ±0.63
(89%)
14. 6 ±0.73
(89%)
15.27±4.18
(118%)
8.65 ±3.16
(100%)
1.4 ±0.54
(93%)
0.0 ±0.00 (11)
50
10.5 ± 1.14*
(70%)
12.3 ± 1.29*
(75%)
32. 17 ±7.48
(249%)
12. 68 ±3.88
(146%)
1.8 ±0.62
(120%)
0.0 ±0.00 (12)
Notes:
a. Adapted from NTP, 1998.
b. Mean ± standard error.
c. Total implants = number of viable fetuses + early resorptions + late resorptions + dead fetuses
d. % Pre-implantation loss = [(corpora lutea -total implants) / corpora lutea] x 100
e. % Post-implantation loss = [(resorptions + dead fetuses) /total implants] x 100
f. Total resorptions = early resorptions + late resorptions
* Statistical significance: p<0.05
Group B females (cohabitation with males on study days 1 to 5 and exposure on GD 6 to
parturition) were assessed for maternal body weight; feed and water consumption; number of
EPA/OW/OST/HECD
V-31
-------
Drinking Water Criteria Document for Brominated Acetic Acids
uterine implantations; number, weight, and anogenital distance of pups; and evaluation of fetal
heart and brain for soft-tissue malformations. The study authors reported that the estimated
average daily doses of BCA were 0, 10, 25, or 61 mg/kg/day for this group. The only observed
effect was an increase in post-implantation losses in all groups, which decreased with increasing
dose, although all losses were elevated relative to controls (303%, 190%, and 184% of the control
value in the 60, 200, and 600 ppm groups, respectively). In addition, total resorptions were
increased to 200%, 137%, and 137% of controls in the 60, 200, and 600 ppm groups, respectively.
None of the effects were statistically different from controls, and the negative dose-response
makes it difficult to assess the biological significance of the findings. No treatment-related effects
were observed in soft-tissue examination (heart and brain) of the fetuses.
Evaluation of the total data set of both significant and non-significant effects suggested to
the authors that BCA adversely affected the ability of females to conceive and carry a full litter to
term. The effects of BCA appear to be particularly relevant for early 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.
Determination of a LOAEL and NOAEL for this study is undermined by the small sample sizes
used in the screening protocol and the low number of pregnancies per dose group. Nonetheless, a
number of reproductive and development effects of significant severity were reported in all dose
groups. Based on biologically-relevant changes that were statistically different from control
EPA/OW/OST/HECD V-32
-------
Drinking Water Criteria Document for Brominated Acetic Acids
values, the LOAEL for reproductive and developmental effects (reduced implants per litter and
live fetuses per litter) was 50 mg/kg/day (high-dose group) and the NOAEL was 19 mg/kg/day
(mid-dose group). It should be noted, however, that the LOAEL and NOAEL might have been
significantly lower if the statistical power of these experiments had been increased by the use of a
larger sample size. As discussed in Section V.A., the high dose was a marginal LOAEL, and 19
mg/kg/day was a NOAEL for maternal toxicity.
The effects of BCA on reproduction in male mice have also been evaluated following oral
gavage dosing. Luft et al. (2000) reported in an abstract on a study in which male C57BL/6 mice
(12 mice/group) were administered daily gavage doses of 0, 8, 24, 72, or 216 mg/kg BCA for 14
days. After 14 days, 5 mice/group were necropsied for histopathological examination of the
testes, epididymis, and seminal vesicles. The remaining 7 males were used in a 40-day breeding
assay to evaluate the effects of BCA treatment on fertility. Coital plug-positive females
(presumably untreated) were replaced daily and uteri were dissected 14 days later; the numbers of
implantations, resorptions, and fetuses were determined. No effects on body weight or
reproductive-organ weights were observed for any of the dose groups. The results of
histopathologic examination of the male reproductive tissues were not reported in the abstract.
However, BCA treatment with 72 or 216 mg/kg/day resulted in adverse reproductive performance,
but only during the first 10 days following treatment (data not shown). Adverse measures of
reproductive outcome included statistically significant decreases in both of the dose groups for (1)
EPA/OW/OST/HECD V-33
-------
Drinking Water Criteria Document for Brominated Acetic Acids
mean number of litters per male (1.1 for both dose groups compared to 3 for controls); (2)
percentage of litters per mated female as measured by the percent of plug-positive females that
became pregnant (36% and 30% in the 72 and 216 mg/kg/day groups, respectively, as compared
to 68% for controls); and (3) total number of fetuses per male (10 and 9 in the 72 and 215
mg/kg/day groups as compared with 27 for controls). There was no difference in the number of
coital plugs, suggesting that treatment did not result in adverse behavioral effects on mating. The
number of fetuses per litter, number of resorptions, and number of terata were also unaltered,
indicating that, under the conditions of this study, adverse reproductive effects in male mice did
not induce developmental toxicity. This study appears to have identified a LOAEL for decreased
male fertility of 72 mg/kg/day and a NOAEL of 24 mg/kg/day, but a definitive conclusion would
require a review of the full study.
Klinefelter et al. (2002a) administered BCA in a dose range finding study (dissolved in
deionized water and pH-adjusted to 6.5) by gavage to adult male Sprague-Dawley rats (12/dose) at
doses of 0, 24, 72, or 216 mg/kg/day for 14 days. The doses were selected to represent the BCA
molar equivalents of 0, 30, 90 and 270 mg/kg/day DBA, previously tested in the same laboratory
(Linder et al., 1994). Endpoints assessed included body weight; testes, epididymes, and seminal
vesicle weights; ex vivo assessment of testosterone production; and serum levels of testosterone,
luteinizing hormone (LH), follicle-stimulating hormone (FSH), and prolactin. Sperm motility,
sperm morphology (cauda and caput), and sperm counts (testicular sperm head count and
EPA/OW/OST/HECD V-34
-------
Drinking Water Criteria Document for Brominated Acetic Acids
epididymal sperm counts) were also evaluated. Testis sections were examined by light
microscopy for delayed spermiation, formation of atypical residual bodies, and germ cell
depletion.
Body weight was significantly decreased in the highest dose group. Testis, epididymis,
and seminal vesicle weights were unaffected by BCA treatment. While spermatid numbers were
not altered by BCA exposure, a significant dose-related decline in epididymal sperm reserves was
observed at 72 and 216 mg/kg/day, with the effect on cauda epididymal sperm being more severe
than on caput epididymal sperm. Dose-related decreases in serum LH, FSH, and prolactin were
noted in all dosed groups, with statistical significance occurring in the two highest dose groups.
No effects on testis sperm production or serum testosterone were observed.
The percentage of motile and progressively motile cauda sperm decreased in a dose-related
fashion, achieving significance in the two highest dose groups. Sperm motion parameters (i.e.,
velocity and linearity) were similarly affected. A dose-dependent reduction in the percentage of
morphologically normal cauda and caput epididymal sperm was also observed. For cauda sperm,
the percent normal sperm decreased to 33% in the 216 mg/kg/day group, as compared with 98.3%
in controls. A similar decrease occurred in caput sperm, with the percent normal sperm being
31.2% in the 216 mg/kg/day dose group, as compared with 94.8% in controls. Caput epididymal
sperm abnormalities were characterized by an increased number of sperm with misshapen heads
EPA/OW/OST/HECD V-35
-------
Drinking Water Criteria Document for Brominated Acetic Acids
or tail defects, whereas cauda sperm abnormalities consisted mainly of an increased number of
isolated heads. Histological evaluation of the testis showed a dose-related increase (statistically
significant in the two highest dose groups) in the number of Step 19 spermatids retained in Stage
X and XI of the spermatogenic cycle. Other findings included a dose-related increase in the
number and size of atypical residual bodies in Stages X and XI (not quantified) and a shift in
localization of these bodies, from basal migration to luminal release, with increasing BCA dose.
According to the study authors, the LOAEL for altered spermiation in this study was 24
mg/kg/day, the lowest dose tested, and a NOAEL could not be determined.
In a subsequent definitive study by the same authors (Klinefelter et al., 2002b), adult male
Sprague-Dawley rats (10/dose) were administered 14 daily gavage doses of BCA (dissolved in
deionized water and pH-adjusted) of 0, 8, 24, or 72 mg/kg/day. End points evaluated were the
same as those assessed in the previous study. Additionally, sperm protein was extracted and
analyzed, and a fertility assessment was conducted via in utero insemination of untreated females
with sperm from treated males.
For the fertility assessment, the estrus cyclicity of a cohort of females was synchronized
by administering a subcutaneous injection of an luteinizing hormone releasing hormone (LHRH)
agonist at 115 hours prior to insemination. At the beginning of the dark cycle following proestrus,
each female was paired with a sexually experienced vasectomized male for 30 minutes. Receptive
EPA/OW/OST/HECD V-36
-------
Drinking Water Criteria Document for Brominated Acetic Acids
females (as indicated by the presence of a copulatory plug) were subsequently anaesthetized, and
epididymal sperm from treated males were injected into each uterine horn at an amount (5 x 106)
that results in approximately 75% fertility in control animals. The sperm from a single male was
used to inseminate a single female. Inseminated females were sacrificed 9 days following
treatment, and implanted embryos and corpora lutea of pregnancy were counted. Male fertility
was expressed as a percentage equivalent to the number of implants/corpora lutea x 100.
No treatment-related changes in body weight, testes weight, and the weight of the seminal
vesicles were observed. However, in contrast with the previous study conducted by the same
authors, epididymal weights were reduced at 72 mg/kg/day, and there were no differences
between treated and control groups in any of the hormonal measurements. Sperm motion
parameters were consistently altered by BCA exposure. Although the percentage of motile sperm
was only decreased in the high-dose group (72 mg/kg/day), progressive sperm motility was
decreased at all doses tested. Altered sperm morphology was only observed at 72 mg/kg/day;
abnormalities in both cauda and caput sperm were similar to those observed in the earlier study,
with the cauda sperm showing increased incidences of sperm with tail defects and the caput sperm
showing increased incidences of sperm with isolated heads. In utero insemination of untreated
females with the cauda epididymal sperm from treated males showed a significant reduction in
fertility at all doses, but no dose-response. Fertility rates in the 8, 24, and 72 mg/kg/day groups
EPA/OW/OST/HECD V-37
-------
Drinking Water Criteria Document for Brominated Acetic Acids
were 33%, 44%, and 37%, respectively, as compared with 75% in control animals. The LOAEL
for this study was 8 mg/kg/day, the lowest dose tested, and a NOAEL could not be determined.
In the sperm protein extraction phase of the study, two-dimensional evaluation of 120
proteins showed significant reductions in two proteins, SP22 and SP9. The shape of the dose-
response curve for SP22 paralleled the reduction in fertility, whereas that for SP9 did not. The
Pearson correlation coefficient was 0.53 (p < 0.001) for SP22 and fertility, and 0.23 for SP9 and
fertility. When the data were fitted to a non-linear threshold response model, the resulting
correlation coefficient (r2) for SP22 and fertility was 0.843. 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.
The authors of the study concluded that the LOAEL for these studies is 72 mg/kg, based on
perturbation of spermatogenesis. However, no NOAEL was determined for this study since it was
concluded that the NOAEL for BCA is less than 8 mg/kg (the lowest dose tested) because the
fertility of sperm from the cauda epididymis was reduced significantly for males exposed to all
doses (8, 24, or 72 mg/kg) although the average fertility of cauda sperm from animals in 8 mg/kg
treatment group was somewhat less than the means for animals in the 24 and 72 mg/kg treatment
groups.
EPA/OW/OST/HECD V-38
-------
Drinking Water Criteria Document for Brominated Acetic Acids
No reproductive- or developmental-toxicity studies for BCA were identified following
dosing by the inhalation or dermal routes.
Dibromoacetic acid
There has been considerable interest in the male reproductive effects of DBA, in part
because its chlorinated analog, dichloroacetic acid (DCA), is known to be a male reproductive
toxicant. Linder et al. (1994a) reported the results of acute-toxicity and acute-spermatotoxicity
studies of DBA. A single-dose protocol was used to identify stages of spermatogenesis that might
be impacted by DBA. In the spermatotoxicity study, male Sprague-Dawley rats (8/group) were
administered a single gavage dose of 0 or 1250 mg/kg DBA and sacrificed 2, 14, or 28 days after
dosing. The approximate LD01 dose of 1250 mg/kg was selected to provide a relatively-high dose
with minimal likelihood of mortality. The study duration was extended to 28 days because of
evidence from the acute-toxicity study that effects on epididymal sperm could peak more than 14
days after dosing. Reproductive-organ weights and sperm-quality parameters were measured, and
a histopathologic examination was performed. Only marginal reproductive-organ-weight changes
were induced by DBA. Epididymis weights on Days 2 and 28 were decreased to 93% and 83% of
control values (p<0.05), respectively, but were not different from controls on Day 14. Testes
weights were decreased to 93% of controls on Day 28 (p<0.05). Prostate weights were
significantly increased to 109% of controls on Day 2 (p<0.05). DBA treatment did not affect
EPA/OW/OST/HECD V-39
-------
Drinking Water Criteria Document for Brominated Acetic Acids
body weights, suggesting that reproductive-organ-weight changes were not secondary to general
toxicity. Serum-testosterone levels fell to 17% of control values 2 days after a single dose of 1250
mg/kg, but returned to control levels by Day 14.
Several measures of spermatotoxicity were reported in this study. Caput-sperm count was
significantly reduced on Day 2 to 85% of controls, but was not affected on Days 14 or 28. Cauda-
sperm count was decreased to 54% and 44% of controls on study Days 14 and 28, respectively.
Testicular-sperm head count was not affected, suggesting that DBA was not inhibiting overall
sperm production. Sperm morphology was also seriously affected by exposure to DBA. The
percent of sperm having fiagellar defects and atypical heads was significantly increased in caput
sperm on Day 28, with 16% showing abnormal morphology. In the control group, about 5% of
sperm were estimated to be abnormal, based on direct inspection of the data presented in a figure
in the paper. Cauda sperm showed a dramatic increase in fiagellar defects on Day 14 (p<0.05),
but not on Day 28. Significant increases in sperm with atypical heads and with both atypical head
and fiagellar defects were increased on Day 28 (64% of sperm displayed abnormal morphology).
According to the study authors, the appearance of different morphological changes (fiagellar
versus acrosomal) on Days 14 and 28 indicated that the epididymal sperm underwent two
sequential morphological changes as a result of DBA exposure. Several measures of sperm
motility were significantly reduced at Day 14, including percent motile (38% of controls), percent
progressive motility (32% of controls), straight-line velocity (73% of controls), and curvilinear
EPA/OW/OST/HECD V-40
-------
Drinking Water Criteria Document for Brominated Acetic Acids
velocity (82% of controls). Only the first two measures were significantly reduced at Day 28
(51% and 41% of controls, respectively). The study authors suggested that these decreases were
related to the flagellar defects. Thus, treatment with a single dose of 1250 mg/kg of DBA resulted
in significantly adverse effects on sperm count, morphology, and motility.
Histopathology examination revealed altered spermiation at all three time-points examined
(Days 2, 14, and 28). On all three days, Step 19 spermatids were retained beyond their normal
release in Stage VIII of the seminiferous epithelium cycle. Other abnormal histological signs
included the presence of remnants of residual bodies in Stages X and XI, and the presence of
anucleate cytoplasmic debris in the lumen of the epididymal duct and in the caput epididymis on
Day 2. On Day 14, debris from the testes was evident in the epididymis, much of which
resembled residual bodies. On Days 14 and 28, abnormal late spermatids were observed in Stages
I-VIII. Similar histological changes were observed on both days, although the changes were
characterized by the authors as less severe on Day 28. Varying amounts of cytoplasmic debris
were also observed in the epididymis on Day 28.
These results show that a single high dose of DBA is spermatotoxic in the rat. The target
cells for adverse effects of DBA were not conclusively identified, although the authors described
several aspects of sperm maturation that might be impacted, based on consideration of the normal
transit times (assuming that the kinetics of sperm development were not affected), and on
EPA/OW/OST/HECD V-41
-------
Drinking Water Criteria Document for Brominated Acetic Acids
consideration of the timing of the observed effects. The study authors noted that flagellar
degeneration was observed in the cauda epididymis on Day 14, but not in the caput epididymis.
The flagellar changes might be due to a DBA effect during transit through the epididymis.
Alternatively, as the normal transit time of sperm through the caput epididymis is 3 days in the rat,
and accounting for the remaining 11 days since dosing, the authors suggested that late spermatids
before or during spermiation might have been affected. On day 28, both altered sperm heads and
flagellar degeneration were observed in both the caput and cauda sperm. As spermatids that were
in Step 11 to 15 on Day 2 of exposure would have comprised the majority of caput sperm on Day
14 (when only a minimal effect on head development was seen), the authors suggested that the
abnormal head development may have resulted from an effect of DBA on Step 10 or earlier
spermatids. Alternatively, they noted that the same effects would have been seen if the effect of
DBA was on later steps but the action of DBA was delayed for several days following dosing. As
the retention of Step 19 spermatids is an effect observed following treatments that alter hormone
status, the observation that DBA reduced circulating-testosterone levels is consistent with the
effects on Step 19 spermatids noted in the study. According to the authors, another potential
target for DBA might be Sertoli cells, since the presence of testicular debris might suggest
disruption of the endocytic activity of these cells. While DBA treatment adversely affected sperm
quality, it did not appear to inhibit sperm production, based on histological analysis of the testes
and the absence of an effect on testicular sperm-head counts.
EPA/OW/OST/HECD V-42
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Linder et al. (1994b) studied the spermatotoxicity of DBA following 14 daily exposures.
The effects on selected endpoints are presented in Table V-5. Male Sprague-Dawley rats (8/dose
group), approximately four months old, were given daily gavage DBA doses of 0, 10, 30, 90, or
270 mg/kg/day, and sacrificed immediately after the last dose. The dose vehicle was distilled
water and the dose volume was 5 mL/kg, adjusted weekly for body weight. No effects on body
weight or serum-testosterone levels were noted at any dose level. Several parameters were
affected (p<0.05), primarily at the highest dose level of 270 mg/kg. These included mildly
reduced testis (93% of controls) and epididymis weights (86% of controls). Absolute and relative
(to testis weight) testicular sperm-head counts were also repressed to 81% and 88% of control
values, respectively. Dose-dependent effects on various measures of sperm motility were also
observed, with statistically significant decreases observed at the highest dose. Caput-sperm
counts were reduced significantly in a dose-dependent fashion beginning at the low dose of 10
mg/kg/day. The percent of morphologically-normal sperm was statistically decreased (79% of
controls) only in the high-dose group, with atypical heads observed more frequently than
degenerating fiagella and a notable increase detected in fused sperm. Cauda-sperm count was
significantly reduced to 76% and 30% of control values in the 90 and 270 mg/kg/day dose groups,
respectively. The percent of morphologically-normal sperm was decreased to 86% and 32% of
controls in the same groups. Morphological changes in cauda sperm were mainly related to
degenerative changes of the fiagella. Percent motile sperm and percent progressive motility were
reduced to less than 10% of control values at 270 mg/kg/day. Straight-line velocity and linearity
EPA/OW/OST/HECD V-43
-------
Drinking Water Criteria Document for Brominated Acetic Acids
were also reduced at the highest dose. Curvilinear velocity was significantly affected only at 90
mg/kg/day.
Histopathological evidence of altered spermiation was noted beginning at 10 mg/kg/day.
Histopathological findings included retention of Step 19 spermatids in Stages IX to XII and
atypical acrosomal development of Step 15 spermatids at 10 mg/kg/day. The severity of these
effects increased with increasing dose. The presence of atypical structures resembling residual
bodies in the testis and caput epididymis was observed at the two highest doses.
EPA/OW/OST/HECD V-44
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table V-5. Sperm Quality Parameters in Rats Given 14 Daily Doses of DBA"
Parameter
Caput sperm 106
Cauda sperm 106
Caput sperm
(% morphologically normal)
Cauda sperm
(% morphologically normal)
Percent motile
Progressive motility (%)
Straight-line velocity
(fim/sec)
Curvilinear velocity (fim/sec)
Linearity
Dose (mg/kg/day)
0
131 ± 14b
264 ± 21
95 ±4
98 ± 1
78 ±7
67 ±4
75 ± 12
154 ±7
52 ±7
10
112± 13*
250 ±70
90 ± 11
90 ± 15
74 ± 14
61 ± 16
68 ± 17
145 ± 10
51 ± 9
30
118± 10*
243 ± 34
96 ± 1
96 ±2
83 ± 6
68 ± 12
66± 16
136± 16
49 ± 8
90
108± 11*
200 ± 59*
93 ± 3
84 ± 10*
66 ± 17
56± 16
63 ± 15
128 ±25*
50 ±5
270
100 ± 10*
78± 15*
75 ± 12*
31 ± 17*
6 ±7*
4± 5*
44 ±22*
141 ±42
30 ± 10**
Notes:
a. Adapted from Linder et al., 1994b
b. Mean ± standard deviation.
* Statistical significance: p<0.05
** Statistical significance: p<0.01
In summary, a variety of male reproductive-tract toxicity parameters were affected by
DBA in this study. Adverse spermatogenic effects were noted beginning at the lowest dose of 10
mg/kg/day and generally increased in severity with increasing dose. Mildly decreased caput-
EPA/OW/OST/HECD
V-45
-------
Drinking Water Criteria Document for Brominated Acetic Acids
sperm count was also observed at this dose, but the effect was not clearly dose-dependent, with
similar decrements (ranging from 76% to 90% of control values) observed at all doses. Decreased
cauda-sperm count, frequency of abnormal cauda-sperm morphology, and changes in sperm
motility were statistically significant only at the two highest doses, 90 and 270 mg/kg/day.
Noticeable histopathological changes (delayed release of Step 19 spermatids and atypical Step 15
spermatid acrosomal development) began at the low dose. The LOAEL for this study is the
lowest dose tested, 10 mg/kg/day, based on histopathological changes in the male reproductive
tract, and a NOAEL could not be determined.
Linder et al. (1995) studied the longer-term (up to 79 days) effects of DBA in male rats on
both reproductive competence (summarized in Table V-6 and Table V-7) and on sperm quality
(summarized in Table V-8 and Table V-9). The highest dose of 250 mg/kg/day was selected
based on the expectation that it would produce substantial spermatotoxicity and permit the
investigation of the time course of DBA effects on fertility and reproductive competence. Lower
doses of 2, 10, or 50 mg/kg/day were selected to obtain dose-response data. Selected doses were
based on the results of previous short-term studies (Linder et al., 1994a; Linder et al., 1994b),
which suggested that both no-efect and significant-effect dose levels would fall within this dose
range. Daily doses of custom-synthesized, high-purity DBA in a distilled-water vehicle were
given by gavage to 105-day-old male Sprague-Dawley rats whose reproductive competence had
EPA/OW/OST/HECD V-46
-------
Drinking Water Criteria Document for Brominated Acetic Acids
been proven. There were essentially two experimental protocols employed, each of which will be
reviewed separately here.
EPA/OW/OST/HECD V-47
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table V-6. Reproductive Outcomes in Rats Following Oral Dosing with DBA"
Mated
Days
8-14
15-21
30-37
65-71
199-213
Dose
mg/kg
0
250
0
250
0
2
10
50
250
0
2
10
50
250
0
250
Male
10
10
10
10
10
10
10
10
10
10
10
10
10
9
10
9
Female
10
10
10
10
10
10
10
10
10
20
20
20
20
9
20
18
Cop
Pairs"
9
3*e
10
5*
9
9
7
7
7
16
17
10*
13
6
15
15
Copulatory
Plugs0
3.3 ± 1.3d
1.3 ± 1.5
3.6 ± 1.1
1.4 ±0.9*
3.6± 2.1
3.1 ± 1.3
2.6 ± 1.4
4.3± 2.2
1.4± 1.1*
3.4 ± 1.3
3.3± 1.3
2.6 ± 1.8
2.4 ± 1.8
1.3 ±0.8*
3.5± 2.0
2.9± 1.4
Sperm
positive
females
9
3*
10
4*
9
9
7
7
1*
14
16
10
13
2*
15
14
Fertile
males
9
2*
10
0*
9
8
7
7
0*
9
10
7
9
0*
10
3*
No.
litters
9
2*
10
0*
9
8
7
7
0*
15 (6)e
14(4)
9(2)
10(1)*
0*
15 (5)
5* (2)
Implants
14.1 ± 1.6
5.5± 6.4*
13.7 ±2.3
—
15.7 ±2.5
15.3 ±3.1
16.3 ± 1.7
13.0 ±5.2
—
14.8 ± 1.4
15.9 ± 1.3
15.8 ±2.9
14.9 ±2.2
—
14.9 ± 1.7
15.3 ± 1.8
Fetuses
13.0 ± 1.3
5.0 ±5.7*
12.6 ±2.9
13.6 ±3.2
14.3 ±3.5
14.9 ±2.0
11.9 ±5.3
—
14.0 ± 1.4
15.1 ± 1.2
14.6 ±2.4
14.1 ±2.4
—
14.4 ± 1.9
14.5 ±2.2
a. Adapted from Linder et al., 1995.
b. Copulatory pairs as evidenced by the presence of copulatory plug or birth of a litter.
c. Per copulating pair
d. Mean± SD.
e. Numbers in parentheses are the number of males siring two litters.
* Statistical significance: p<0.05
EPA/OW/OST/HECD
V-48
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table V-7. Outcome of Artificial Insemination of Sperm from Rats
Dosed with DBA"
Day
9
16
31
79
Dose
mg/kg
0
250
0
250
0
2
10
50
250
0
2
10
50
Number of
Inseminations
6
6
6
5(l)c
6
6
6
6
1 (5)
10
9(1)
10
10
Number of
litters
5
5
6
1*
5
6
3
5
0
7
6
8
5
Implants
5.40±3.36b
6. 00 ±3. 32
7. 83 ±4.79
4.00
7.40 ± 4.77
9. 50 ±5.24
5.33 ±2.08
5. 80 ±3. 11
—
7. 86 ±2. 91
7. 17 ±4.26
8. 75 ±2.43
9. 00 ±2.83
Fetuses
5. 20 ±3.63
6. 00 ±3. 32
7. 67 ±4. 63
4.00
7. 40 ±4.77
9. 17 ±4. 96
5.00± 1.73
5. 80 ±3. 11
—
7.71 ±2.69
6. 83 ±4.45
8.63 ±2.67
9. 00 ±2. 83
Notes:
a. Adapted from Linder et al., 1995.
b. Litter means ± SD.
c. Number in parentheses is the number of males with insufficient sperm for insemination.
* Statistical significance, p<0.05
EPA/OW/OST/HECD
V-49
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table V-8. Reproductive Organ Weights and Sperm Counts
in Rats Given Daily Doses of DBAa
Slumber
of
Doses
2
5
9
16
31
79
Oe
42f
mg/
kg
0
250
0
250
0
250
0
250
0
2
10
50
250
0
2
10
50
0
250
IT
6
6
6
6
6
6
5
6
6
5
6
6
6
10
10
10
10
10
9
Body
weight (g)
402 ± 17d
401 ± 14
400 ± 16
405 ± 14
411 ±21
404 ±28
407 ±23
375 ± 58
432 ± 14
417± 16
424 ±21
425 ±22
368 ± 38*
458 ± 24
455 ± 16
446 ±20
434 ± 24*
522 ±22
483 ± 19*
Testis
weight (g)
1.97±0.13
1.92 ±0.09
1.89 ± 0.14
1.85 ±0.29
1.94 ±0.08
1.87±0.14
1.96 ±0.21
1.74±0.18
1.93 ±0.12
1.93 ±0.08
1.94±0.15
1.90 ± 0.14
1.81 ±0.05
2.01 ±0.15
1.96±0.12
2. 02 ±0.14
1.97±0.11
2. 02 ±0.10
0.99 ±0.48*
Epididymis
weight (g)
0.64 ±0.02
0.63 ±0.02
0.62 ±0.05
0.58 ±0.08
0.62 ±0.03
0.60 ±0.05
0.64 ±0.04
0.51 ±0.07*
0.64 ±0.03
0.65 ±0.02
0.64 ±0.06
0.60 ±0.02
0.51 ±0.03
0.68 ±0.03
0.67 ±0.03
0.69 ±0.04
0.64 ±0.05
0.68 ±0.06
0.49 ±0.09*
TSHCC
(million)
271 ± 17
278 ± 15
261 ±26
273 ±56
295 ± 15
276 ±34
288 ±26
263 ±27
265 ±24
280 ±26
264 ±35
282 ±28
283 ±25
298 ± 32
270 ±25
288 ±21
289 ±24
260 ±23
50 ± 106*
TSHC per
gram
testis
(million)
149 ± 10
157 ± 8
151 ± 12
158± 10
162 ±7
156± 8
162± 19
162± 11
147 ± 8
155± 10
145± 16
160± 11
167± 13*
159± 11
148± 11*
155 ±6
159± 11
139± 6
34 ±56*
Caput
sperm
(million)
124 ± 8
115± 9
115± 12
97 ±24
124 ±7
119± 14
122 ±27
92 ±21
128 ± 9
125 ± 13
121 ± 9
111 ± 10*
47 ± 9*
126±7
126± 10
122 ± 10
112± 11*
118± 9
20 ± 40*
Cauda
sperm
(million)
254 ±28
278 ±41
239 ±29
230 ±72
250 ±23
225 ±42
255 ±23
75 ±32*
240 ± 14
225 ±44
225 ± 34
169 ± 34*
33 ± 5*
240 ±40
247 ±28
242 ±34
196 ±47*
249 ±45
37 ±74*
Serum
testosterone
(ng/mL)
8.0 ± 4.1
4.1 ± 2.1
10.7 ±4.9
6.9± 6.1
3. 4 ±1.7
3.4 ± 3.2
4.1 ± 3.5
2.9± 1.5
4.0 ± 3.3
11.8 ± 6.4
5.6± 5.7
3.9± 2.5
2.6 ± 1.6
10.5 ±8.4
7.1 ± 3.8
6.2 ± 4.9
3.8± 1.6
2.7 ± 1.3
2.7 ± 1.5
Notes:
a. Adapted from Linder et al., 1995.
b. N= number of rats.
c. TSHC = testicular sperm head count
d. Group mean ± SD.
e. Nondosed controls.
EPA/OW/OST/HECD
V-50
-------
Drinking Water Criteria Document for Brominated Acetic Acids
f. Given 42 doses then allowed to recover for 186 days.
* Statistical significance: p<0.05
EPA/OW/OST/HECD V-51
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table V-9. Sperm Quality Parameters in Rats Given Daily Doses of DBA3
Number
of
Doses
2
5
9
16
31
79
mg/kg
0
250
0
250
0
250
0
250
0
2
10
50
250e
0
2
10
50
Nb
6
6
6
6
6
6
5
6
6
5
6
6
6
10
10
10
10
Motile
sperm (%)
84 ±5d
87 ±7
84 ±5
83 ±8
84 ±8
84 ±7
86 ±5
13 ± 18*
78 ± 3
81 ±4
72 ± 20
47 ±32*
3 ±6
75 ± 8
77 ±8
76 ± 11
66 ± 13
Progressive
motility
(%)
75 ±7
75 ± 5
72 ±7
70 ±7
72 ± 8
73 ± 8
70 ± 6
10 ± 15*
61 ±7
65 ± 6
53 ± 16
36 ±28
2 ± 4*
61 ±8
63 ±8
62 ±9
53 ± 12
VSLC
(fim/sec)
80 ±7
79 ±7
80 ± 18
77 ±9
87 ± 17
79 ± 15
63 ±24
31 ± 13*
67 ± 15
74 ± 10
68 ± 16
55 ±28
—
70 ± 9
72 ± 11
76 ± 10
64 ± 9
VCLC
(fim/sec)
105 ± 9
107 ±7
108 ± 17
107 ± 11
120 ±21
107 ±20
92 ±36
51 ±24*
107 ±22
113 ± 16
116 ±31
91 ±36
—
107 ±11
106 ± 13
115 ± 17
96 ± 14
Linearity
48 ±4
46 ±5
46 ±7
44 ±3
47 ±4
49 ±3
43 ± 3
32 ± 10*
39 ±5
42 ±4
41 ±5
35 ±8
—
42 ± 3
42 ±4
42 ±2
41 ±4
Caput sperm
% normal
96 ± 1 (0)d
96 ± 2 (0)
97 ± 2 (0)
90±4*(3)
96 ±2 (0)
55 ± 17* (8)
98 ± 1 (0)
61 ±37* (2)
96 ± 2 (0)
97 ± 1 (0)
96 ± 6 (0)
88 ±7* (.3)
1 ±2* (5)
97 ± 1 (0)
94 ± 6 (0.02)
96 ± 2 (0)
68 ±20* (1)
Cauda sperm
% normal
96 ± 2 (0)e
97 ± 1 (0)
96 ± 2 (0)
95 ± 4 (0)
96 ± 1 (0)
95 ± 2 (0)
97 ± 1 (0)
33 ±20* (3)
96 ± 1 (0)
96 ± 2 (0)
96 ± 4 (0)
88 ± 10* (0)
2 ±2* (3)
96 ±2 (0)
93 ± 8 (0)
94 ± 4 (0)
75 ± 17* (0.06)
Notes:
a. Adapted from Linder et al., 1995.
b. N = number of rats.
b. VSL is straight line velocity and VCL is curvilinear velocity.
c. Group means ± SD.
d. Number in parentheses is the percent of fused sperm.
e. Motile sperm (15% and 1% motile) were present in only two rats.
* Statistical significance: p<0.05
EPA/OW/OST/HECD
V-52
-------
Drinking Water Criteria Document for Brominated Acetic Acids
EPA/OW/OST/HECD V-53
-------
Drinking Water Criteria Document for Brominated Acetic Acids
In the first protocol, groups of 10 male rats were given daily gavage doses of either 0 or 250
mg/kg/day DBA. During study days 8-14, 15-21, and 30-37, the males were paired with females
and allowed to mate by natural insemination. Dosing was terminated after 42 days because of the
onset of overt toxicity, including labored breathing, light tremor, difficulty moving the hind limbs,
and severe weight loss. The animals, however, were allowed to mate during recovery, on days 49-
56, 65-71, and 199-213. During the mating period on study days 8-14, only 3/10 males copulated;
only two males were fertile and only two litters were produced. The numbers of implants and
fetuses in these litters were reduced by more than 50% as compared to controls. During the
mating periods on days 15-21 and 30-37, there were no fertile males and no litters were produced,
even though 5/10 and 7/10 males, respectively, copulated during these periods. To distinguish
fertilization failure from pre-implantation loss, females from the mating period on days 49-56 were
sacrificed on GD 1 and examined for the presence of fertilized eggs. There were no fertilized eggs.
During the mating period on days 65-71, no males were fertile and no litters were produced.
During the mating period on days 199-213 (after 5 months of recovery), only 3/9 males were fertile
and 5 litters were produced, even though all males copulated. Thus, although the reproductive
performance of animals in the 250 mg/kg/day improved significantly during recovery, they never
fully recovered.
Artificial insemination of luteinizing hormone releasing hormone (LHRH)-synchronized
females, with sperm from treated males from an additional group of 6 animals, sacrificed on Days
EPA/OW/OST/HECD V-54
-------
Drinking Water Criteria Document for Brominated Acetic Acids
9, 16, and 31 (and also used for interim necropsy), was conducted to distinguish behavioral- mating
effects from effects due to physiologic reproductive competence. Five litters were sired with Day 9
sperm; no significant adverse effects were observed on the number of implants or fetuses. The
absence of a significant effect on fertility by artificial insemination with Day 9 sperm suggested to
the study authors that the reproductive effects observed from the Day 8-14 mating may have been
due to a transient effect on libido. In contrast, only one litter was produced as a result of artificial
insemination with Day 16 sperm, and no litters resulted from insemination with Day 31 sperm,
indicating that reproductive incompetence was due to spermatotoxic effects. Consistent with these
reproductive outcomes, measures of sperm motility were not affected until Day 16, but were
severely affected thereafter. The cauda-sperm count was normal until Day 16, at which time it was
reduced to 29% of controls. Caput-sperm counts decreased progressively to as low as 37% of
normal on Day 31. The percent of sperm with normal morphology was significantly decreased
(p<0.05) beginning on Day 5 for caput sperm and on Day 16 for cauda sperm. Only minimal
developmental-toxicity data were provided in the paper. No effects on fetal weight were observed;
minimal changes in the incidence of malformations were inconsistently observed and were not
considered to be treatment-related by the study authors.
In the second protocol, groups of 10 male rats were given daily gavage doses of 0, 2, 10, or
50 mg/kg/day DBA for up to 79 days. The only systemic effect was a slight decrease in body
weight (to 95% of controls) that was apparent by Day 53 in the 50 mg/kg/day group. The rats were
EPA/OW/OST/HECD V-55
-------
Drinking Water Criteria Document for Brominated Acetic Acids
mated during study days 30-37 and 49-56 with one female per male, and during study days 65-71
with two females per male. No effects on the number of fertile males, litter size, fetal body weight,
or number of implants per litter were observed. There was a dose-dependent, but not always
statistically significant, reduction in copulating pairs and copulatory plugs relative to controls
during the 65-71 day mating period. During this final mating period, there was also a dose-
dependent decrease in the number of males siring two litters, which was statistically significant
only at the highest dose tested, 50 mg/kg/day. The effects on mating behavior were similar at 10
mg/kg/day and 50 mg/kg/day, but there was no clear dose response. For example, there were fewer
copulatory plugs and multiple litters at 50 mg/kg/day than at 10 mg/kg/day, but the higher dose had
more copulatory pairs and/or inseminations (depending on the mating period). The mating-
behavior effects in the 10 mg/kg/day dose group included fewer copulating pairs, fewer
inseminations, fewer copulatory plugs, and fewer multiple litters, but the only statistically
significant (p<0.05) effect at this dose was fewer copulating pairs in the Day 65-71 group.
Artificial insemination of LHRH-synchronized females was performed with sperm from an
ancillary group exposed to 0, 2, 10, or 50 mg/kg/day and sacrificed on Day 31 (6 males/group) or
sacrificed on Day 79 (10 males/group). No significant effects on reproductive outcomes were
observed. Necropsy results revealed that caput and cauda sperm counts were significantly reduced
(p<0.05) at the high dose to 87% and 70% of control values, respectively, on Day 31, and to 89%
and 82% of control values, respectively, on Day 79. The percent motile sperm was affected on
EPA/OW/OST/HECD V-56
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Day 31, but not on Day 79 at 50 mg/kg/day. Exposure to 50 mg/kg resulted in moderate changes in
sperm morphology, including head and tail defects and fused sperm, when examined at either Day
31 or Day 79. No gross effects on sperm morphology were seen at 2 or 10 mg/kg/day. A slight
increase in dead fetuses of 5% (compared to 0% in the controls) was reported for the Day 79
artificial insemination group dosed with 50 mg/kg DBA. No other statistically significant,
developmentally-toxic effects were reported.
Histopathologic results from the Linder et al. (1995) study were reported separately (Linder
et al., 1997a). Necropsies of rats dosed with 0 or 250 mg/kg/day DBA were performed 24 hours
after the last of 2, 5, 9, 16, or 31 daily doses. A necropsy was also done at Day 228, which
included 42 days of exposure and a 6-month recovery period. At Days 2 and 5, there was moderate
to extensive retention of Step 19 spermatids (normally released in Stage VIII) in Stage IX of the
cycle of the seminiferous epithelium. This retention was also seen in Stages IX and X at Days 9
and 16, and in virtually all Stage IX and X tubules at Day 31. Also at Day 31, there was retention
of Step 19 spermatids and degenerating Step 19 spermatids in Stage XI to XIV tubules. Basally-
located remnants of Step 19 nuclei were seen in Stages X to XII at Day 5, Stages X to XII at Day 9,
Stages K-XII at Day 16, and in Stages XI-XIV at Day 31. Fused Step 19 spermatid flagella in
Stage IX were seen from Day 5 through Day 31. Atypical residual bodies were seen from Day 5
through Day 16 at numerous Stages, and appeared in the epididymis on Days 9, 16, and 31. Debris
from these atypical residual bodies, along with other cytoplasmic debris from the testes, was
EPA/OW/OST/HECD V-57
-------
Drinking Water Criteria Document for Brominated Acetic Acids
observed throughout the epididymis on Day 31. Acrosomes and/or heads of Step 12 and later
spermatids were affected at Days 16 and 31, and maloriented late spermatids were also observed at
these times. Vacuolization of the Sertoli-cell cytoplasm was seen at Day 31. After a fairly long (6
month) recovery period, male rats dosed with 250 mg/kg/day still displayed atrophic seminiferous
tubules, disorganization of sperm-producing tubules, and degeneration of mature and round
spermatids, all likely resulting from effects on the structure and/or function of Sertoli cells and
indicative of permanent damage to the reproductive system.
In the same study, necropsies of rats given 0, 2, 10, or 50 mg/kg/day were performed 24
hours after the last of 31 or 79 daily doses. Retention of Step 19 spermatids near the tubule lumen
in Stage IX was observed at Day 31 at doses of 10 mg/kg/day and higher. At Day 79 in the 10
mg/kg/day dose group, there was also retention of Step 19 spermatids in Stages IX-XI at both the
lumenal and basal surfaces. One animal at this dose also had disorganized and atrophied tubules,
but this effect was not considered to be treatment-related because tubular disorganization was not
seen at 50 mg/kg/day. The 50 mg/kg/day dose group (after 31 doses) had increased retention of
Step 19 spermatids, with moderate numbers observed at the lumenal surface at Stage IX and at the
basement membrane at Stages X-Xn. Similar effects were seen when this dose group was treated
for 79 days, with the retention of Step 19 spermatids also occurring at Stages IX and X. Atypical
residual bodies were present in Stage IX at Day 31 and occasionally at Stage IX and other stages at
Day 79. Cytoplasmic debris was observed throughout the epididymis at Day 79 in the 50
EPA/OW/OST/HECD V-58
-------
Drinking Water Criteria Document for Brominated Acetic Acids
mg/kg/day dose group. No histopathological changes were detected at 2 mg/kg/day on either Day
31 or Day 79.
Thus, the two papers for this study (Linder et al, 1995; Linder et al, 1997a) describe
increasingly severe effects with increasing dose and increasing exposure duration. No adverse
effects were seen at the lowest dose tested, 2 mg/kg/day. Dosing with 250 mg/kg/day for as few as
8-14 days caused decreased mating and decreased fertility, and adverse reproductive effects were
only partially reversible after 42 days of exposure and a 6-month recovery period. Overall, this
study identified an equivocal LOAEL of 10 mg/kg/day and a corresponding NOAEL of 2
mg/kg/day for male reproductive effects, based on histological evidence for changes in
seminiferous tubule staging of altered spermatid development.
Collectively, the studies by Linder and colleagues have used a number of different
experimental protocols to investigate the effects of DBA on spermatogenesis and the resulting
effects on male fertility (Linder et al., 1994a; Linder et al., 1994b; Linder et al., 1995; Linder et al.
1997a). Based on the results of these studies, DBA was clearly spermatotoxic in rats following
high-dose single exposures or repeated exposures for longer periods of time (up to 79 days).
Effects on spermatogenesis were the most sensitive endpoint because they were observed in the
absence of other toxicity indicators. Significant changes in sperm count, morphology, and motility
were generally observed at doses higher than those associated with early histopathologic changes in
EPA/OW/OST/HECD V-59
-------
Drinking Water Criteria Document for Brominated Acetic Acids
spermatogenesis. However, the susceptibility of humans to DBA-induced reproductive toxicity is
not known and it is possible that rats are more sensitive than humans. In the absence of valid and
reliable human data on the relationship between sperm quality and human fertility, and on the
relative sensitivity of humans versus rats to DBA-associated reproductive effects, adverse
histopathologic or sperm-quality changes in rodents are considered to be the more appropriate
choice for the critical effect in the studies by Linder and his colleagues than fertility changes.
In contrast to the results of Linder et al. (1994a), Vetter et al. (1998) did not observe a
significant spermatotoxic effects of acute treatment with DBA. Vetter et al. (1998) evaluated
spermatotoxic effects of DBA as a positive control to validate a computer-assisted semen analysis
and a flow cytometric assay for cell-membrane integrity as alternatives to sperm-motility assays for
assessing male-reproductive toxicity. Sexually-mature male Crl:CD(SD)BR rats (4-5/group) were
given single oral doses of 0, 600, or 1200 mg/kg DBA in 10 mL/kg deionized water. The high
dose, but not the low dose, resulted in overt toxicity. The rats were sacrificed after 13 days, at
which time vas-deferens sections were harvested for sperm analysis, and sections of the testes and
epididymides were taken for histopathologic analysis. The average percent motile sperm was
74.4%, 74.8%, and 65.7% for the control, low-, and high-dose groups, respectively. The average
percent of viable sperm was 90.9%, 91.4%, and 88.7%, for the control, low- and high-dose groups,
respectively. Neither sperm motility nor membrane permeability following DBA treatment were
statistically different from controls. The absence of an effect was not likely due to general failure
EPA/OW/OST/HECD V-60
-------
Drinking Water Criteria Document for Brominated Acetic Acids
of the assays since a second positive control, cc-chlorohydrin, did decrease sperm motility.
Following DBA treatment, no morphologic changes were observed in the sperm; however, large
basophilic bodies were observed in the testes of low-dose rats (3/5 males) and high-dose rats (4/4
males), and in the epididymides of high-dose rats (4/4 males). The study authors noted that
differences in experiments, such as the source of the sperm from the vas deferens versus the cauda
epididymis, the strain of rat, and the time of sacrifice (13 versus 14 days post-treatment) are
unlikely explanations for the absence of spermatotoxicity in this study as compared with the results
reported by Linder et al. (1994a). In the absence of further data, the reasons for the differing
results in the Linder et al. (1994a) and Vetter et al. (1998) experiments remain unresolved. The
Vetter et al. (1998) study identified a LOAEL of 600 mg/kg/day, based on histopathologic changes
in the testes. A NOAEL could not be determined.
Although the effects of DBA on male reproductive-tract toxicity have been well studied,
fewer studies have evaluated the potential reproductive effects of DBA in females. Cummings and
Hedge (1998) studied the effects of DBA exposure during early pregnancy in rats. Female
Holtzman rats (8/dose group) were administered gavage doses of 0, 62.5, 125, or 250 mg/kg/day
DBA dissolved in water on GD 1-8. Administration of a higher dose, 500 mg/kg/day, induced
moribund behavior and lethality; therefore, dosing at this level was discontinued and these animals
were not further evaluated for reproductive endpoints. Treated animals from the other dose groups
were sacrificed on GD 9, and body and reproductive-organ weights, serum levels of progesterone,
EPA/OW/OST/HECD V-61
-------
Drinking Water Criteria Document for Brominated Acetic Acids
17p-estradiol, and luteinizing hormone, the number of implantation sites, the number of
resorptions, the number of corpora lutea, and pre-implantation losses were assessed. The only
affected response was a 170 % increase in serum 17p-estradiol at 250 mg/kg/day. A second group
of females was dosed similarly to the first group, sacrificed on GD 20, and evaluated for body
weight, preimplantation losses, number of resorptions, number of pups per litter, pup weights, and
placental weights. No differences in any of these measures were observed between treated and
control animals. The authors concluded that DBA had little effect on female reproduction for the
measures assessed in the study, but noted that effects on ovarian function and future fertility were
not tested; such tests would be warranted by the observed increase in serum 17p-estradiol. Based
on the increase in serum 17p-estradiol in this study, the LOAEL is the highest dose tested, 250
mg/kg/day, and the NOAEL is 125 mg/kg/day. An acute PEL of 500 mg/kg/day was also
identified, based on moribund behavior and lethality in the pregnant dams. This study, however, is
limited by the small sample size of each of the groups.
Christian et al. (2001) evaluated the reproductive and developmental toxicity of DBA in
Sprague-Dawley rats. Male and female rats (10/sex/group) were given DBA in deionized drinking
water at concentrations of 0, 125, 250, 500 or 1000 ppm, beginning 14 days prior to cohabitation
and continuing through gestation and lactation (63-70 days of treatment). The average daily doses
(based on measured water consumption and body weights) varied, depending on the phase of
reproduction. For males throughout the study (SD 1-70), mean daily doses were 10.2, 20.4, 35.7,
EPA/OW/OST/HECD V-62
-------
Drinking Water Criteria Document for Brominated Acetic Acids
and 66.1 mg/kg/day, respectively. For females on SD 1-15 (pre-mating), mean daily doses were
13.3, 26.2, 41.8 and 60.2 mg/kg/day, respectively; and 14.8, 30.3, 48.5 and 81.6, respectively, on
gestation day (GD) 0-21. During lactation (LD 1-29), the estimated doses were 0, 43.5, 86.6, 150.7
and 211.7 for the 0, 125, 250, 500, and 1000 ppm groups, respectively; however, these doses
included consumption of water by the pups and thus overestimated the mean daily intake for
lactating females. Among the pups, two male and two female weanlings from each litter were
selected for one additional week of observation (postweanling days 1-8, commencing on LD 29);
daily food intake, drinking water consumption and body weights were recorded, and necropsy was
conducted at sacrifice. The mean daily doses for the weanling pups were 0, 31.8, 58.5, 122.9 and
254.7 mg/kg/day for males, and 0, 33.3, 61.5, 123.8, and 241.2 for females in the 0, 125, 250, 500
and 1000 ppm groups, respectively.
Apparent taste aversion was associated with an exposure-dependent reduction in water
consumption, which was paralleled by a reduction in food intake at all concentrations. Decreased
body weight gain was observed in parental animals and postweanling pups at the two highest
exposure levels. Estrous cycling was unaffected in the female rats. The only observed adverse
reproductive effect was a possible reduction in mating performance in the 1000 ppm group, as
evidenced by a slight but nonsignificant increase in the number of days of cohabitation and a
decrease in the number of mated pairs (6/10 in the 1000 ppm group versus 9-10/10 in all other
groups). There were no effects on pre- and postimplantation losses, live litter sizes, and gross
EPA/OW/OST/HECD V-63
-------
Drinking Water Criteria Document for Brominated Acetic Acids
external morphology or sex ratios in the pups. Although an exposure-related decrease in pup body
weights was noted, these findings were attributed to decreased water and food consumption
resulting from the poor palatability of DBA-treated drinking water. Based on a lack of statistically
significant, treatment-related findings, the parental and reproductive/developmental NOAEL for
this study is the highest dose tested, and a LOAEL could not be determined. For males, the
paternal NOAEL is 66 mg/kg/day; for females, the corresponding NOAEL is not less than 60
mg/kg/day, and is likely to be higher, as water consumption and corresponding mean DBA daily
doses were increased during gestation (to 82 mg/kg/day) and lactation (mean daily doses could not
be determined due to the confounding effects of water consumption by the pups). Similarly, the
NOAEL for developmental effects is at least 82 mg/kg/day (maternal dose during gestation).
The Chemistry Council (CCC, 2001; Christian et al, 2002) recently completed a two-
generation drinking water study of DBA in rats, conducted according to Good Laboratory Practice
(GLP) standards and U.S. EPA test guidelines. The report has recently been published and has also
been independently reviewed and accepted by an EPA scientific advisory group. Because this
study addresses a key data gap, a fairly detailed summary of the reported findings is presented here.
Male and female Crl:CD Sprague-Dawley rats (30/sex/exposure group) were administered DBA in
drinking water at concentrations of 0, 50, 250, or 650 ppm continuously from initiation of exposure
of the parental (P) generation male and female rats through weaning of the F2 offspring. The
concentrations were chosen based on a range-finding study that found that 650 ppm was the
EPA/OW/OST/HECD V-64
-------
Drinking Water Criteria Document for Brominated Acetic Acids
highest concentration expected to allow survival of the Fl offspring. For the P generation, DBA
exposure was initiated at 43 days of age and continued from premating until study day (SD) 92 for
males; and from premating through gestation and a 29-day period of lactation (LD 1-29) (for
approximately 120 days of exposure) for females. Parental generation offspring (Fl males and
females) were exposed in utero during gestation, and during lactation (LD 1-29); selected Fl males
and females (30/sex/exposure group) were further exposed during a postweaning period of at least
71 days, which continued through mating, gestation, and lactation. All other Fl pups were
sacrificed on LD 29. All Fl adult females and their offspring (F2 generation) were sacrificed on
LD 22. All females in the P and Fl generations were evaluated once daily for estrous cycling
(from 21 days before cohabitation through GD 0). All females were also assessed for duration of
gestation, fertility index, gestation index, number and sex of offspring per litter, number of
implantation sites, litter size and viability, viability index, lactation index, percent pup survival and
litter sex ratio, general condition of the dam and litter during the postpartum period, and maternal
behavior during lactation. Litters were examined to identify external abnormalities, physical signs
of toxicity, pup weights, and litter viability. Necropsy of all P and Fl adults included gross
evaluation of the cranial, thoracic, abdominal, and pelvic viscera. Specialized measurements
evaluating sperm parameters (concentration, percent motility, morphology, number of sperm, and
testicular spermatid count). Fl generation pups were also evaluated for age at sexual maturation
(as determined by vaginal patency in females, preputial separation in males, and anogenital
distance in both sexes) were performed. Individual organ weights were recorded for major organs,
EPA/OW/OST/HECD V-65
-------
Drinking Water Criteria Document for Brominated Acetic Acids
including testes and ovaries, as well as uterus with oviducts and cervix, epididymides, prostate
gland, and seminal vesicles with coagulating glands. All gross lesions were examined
histologically. Histopathology was also conducted on adrenal and pituitary glands; testis,
epididymides, prostate, seminal vesicles, coagulating glands in males; ovaries, oviducts, uterus,
cervix and vagina in females; and selected additional organs based on observed organ weight
changes. Additionally, testicular histopathology examining the caput, corpus, and cauda of the
epidiymis was conducted in males, and the reproductive organs of all rats suspected of reduced
fertility were subjected to histological examination.
The average daily doses (based on measured water consumption and body weights) are
presented in Table V-10. Daily doses varied both between exposure groups and among
reproductive stages (premating, gestation, lactation). Significant increases in pup mortality in Fl
litters were considered to be unrelated to DBA exposure because the incidences were within the
historical control of the testing facility. Other unscheduled deaths in the study were also unrelated
to exposure to DBA. Clinical signs of toxicity were observed in various groups exposed to 250
and 650 ppm, and included soft or liquid feces, dehydration, and ungroomed coats. Water
consumption were statistically significantly decreased in the P and Fl generation at all exposure
levels, presumably due to taste aversion, and food intake was significantly reduced at the highest
dose group in the P generation and the two high exposure groups in the Fl generation. Body
weights and body weight gains for high-dose P males and females were significantly reduced
EPA/OW/OST/HECD V-66
-------
Drinking Water Criteria Document for Brominated Acetic Acids
during the premating period and were significantly decreased for high-dose P females during
gestation and lactation. Fl male and female pups had significantly reduced body weights at all
exposure levels during the lactation period, sufficient for the study authors to delay weaning until
LD 29 to ensure pup survival.
Table V-10. Average Consumed Daily Doses (mg/kg/day) for Male and Female Sprague-
Dawley Rats in the Two-Generation Reproductive/Developmental Toxicity Study"
DBA Exposure Groups
0 ppm
50 ppm
250 ppm
650 ppm
P Generation - Male Rats
Premating to Termination (SD 1-92)
0.0
4.4
22.4
52.4
P Generation - Female Rats
Premating to Cohabitation (SD 1-70)
Gestation (GD 0-21)
Lactation (LD 1-15)
0.0
0.0
0.0
6.0
6.4
11.6
28.1
30.1
55.6
69.1
76.1
132.0
Fl Generation - Male Rats
Premating (PD 1-71)
Weaning to Termination (PD 1-134)
0.0
0.0
5.7
4.5
29.7
22.0
74.6
54.7
Fl Generation - Female Rats
Weaning to Cohabitation (PD 1-71)
Gestation (GD 0-21)
Lactation (LD 1-15)
0.0
0.0
0.0
6.6
6.2
10.0
32.1
28.5
49.6
83.4
67.1
114.7
Chlorine Chemistry Council (2001), unpublished report; Christian et al., (2002)
bSD = study day; GD = gestation day; LD = lactation day; PD = postweaning day
EPA/OW/OST/HECD
V-67
-------
Drinking Water Criteria Document for Brominated Acetic Acids
By LD 29, the body weights of pups in the 50 ppm group were similar to control pup
weights. Throughout the postweaning/premating period, Fl males and females in the 250 and 650
ppm groups weighed significantly less than controls, and the females continued to exhibit
significant reductions in body weight (compared to controls) during gestation and lactation. The
body weights of F2 pups in the two highest dose groups were also reduced by the end of lactation;
however, these reductions were not reported to be statistically significant, relative to control group
values. No treatment-related effects were reported in either generation for estrous cycling, number
of days in cohabitation, duration of gestation, mating indices, fertility indices, number and sex of
offspring per litter, number of implantation sites, litter size, lactation index, percent pup survival,
pup sex ratio, and gross malformations. Total litter loss observed in the P generation for two dams
in the 250 ppm exposure group and one dam in the 650 ppm group was not considered to be
treatment-related.
For the Fl generation 650 ppm exposure group, preputial separation was significantly
delayed in the male rats (50.5 days versus 48.1 days in controls), and vaginal patency in female rats
was also significantly retarded (36.3 days versus 33.4 days in controls); no significant difference
was seen when the data were analyzed using body weight as a covariant. These effects were
considered to be due to a general retardation of growth associated with the significant reduction in
body weight in this exposure group at weaning. In F2 male and female pups, anogenital distance
did not differ from controls on LD 1 but was significantly reduced in male pups in the 250 and 650
EPA/OW/OST/HECD V-68
-------
Drinking Water Criteria Document for Brominated Acetic Acids
ppm by LD 22; these findings were also considered to be associated with a general retardation of
growth rather than being treatment-related.
An increased incidence of malformations of the male reproductive tract, including small
testes and small or absent epididymides, was observed in four males in the Fl group exposed to
650 ppm and was considered to be treatment-related. Histomorphologic examination of these
organs in these males revealed a minimal increase in abnormal residual bodies, retained Step 19
spermatids, hypospermia, atrophied epididymis and/or atrophied testis. Histopathologic
examination of reproductive organs of P and Fl male rats in the 250 and 650 ppm groups (N =
30/group/generation) showed a consistent and significant exposure-related increase in retained Step
19 spermatids in Stage IX and X tubules and in increased and abnormal residual bodies in affected
seminiferous tubules (Table V-l 1). Diffuse testicular atrophy and phagocytized Step 19 nuclei in
the basilar area of affected seminiferous tubules were also observed, although at a lower incidence.
Other testicular abnormalities in 250 and 650 ppm male rats of both generations included increased
amounts of exfoliated spermatogenic cells/residiual bodies in epididymal tubules, atrophy, and
hypospermia. Percent motile sperm, sperm count, sperm density, and number and percent of
morphologically abnormal sperm for exposed groups were within historical control values for the
test laboratory and were unaffected by treatment. No effects were observed in the prostate gland,
seminal vesicles, or coagulating glands of any of the male rats of either generation. All gross
EPA/OW/OST/HECD V-69
-------
Drinking Water Criteria Document for Brominated Acetic Acids
lesions in other organs in P and Fl parental rats and in Fl ands F2 pups were considered to be
unrelated to DBA treatment.
Table V-ll. Incidences of Exposure-Related Histopathologic Findings in
the Testes of Rats Consuming DBA in Drinking Water "
DBA Concentrations (ppm)
Number of Testes Examined
Retention of Step 19 Spermatids
Abnormal/Increased Residual Bodies
P
0
30
4
3
50
30
3
5
250
30
13
15
650
30
23
25
Fl
0
30
0
1
50
30
1
2
250
30
12
10
650
30
20
14
1 Chorine Chemistry Council (2001), unpublished report
Histologic examination of the ovaries often P and Fl female rats in the control, 250 and
650 ppm exposure groups did not reveal any functional abnormalities; corpora lutea and growing
and antral follicles were present and apparently normal. There were no significant differences in
the number of postlactational ovarian primordial follicles among any of these groups.
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 Fl males and females. There was no dose-
response 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
EPA/OW/OST/HECD
V-70
-------
Drinking Water Criteria Document for Brominated Acetic Acids
in 10 rats/sex/group, and the study authors did not consider the organ weight changes to be
lexicologically 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. Fl 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 Fl 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 Fl males, the reproductive/developmental toxicity LOAEL and NOAEL
are 250 and 50 ppm, respectively. For the P generation, these drinking water concentrations
correspond to a LOAEL and NOAEL of 22 and 4 mg/kg/day, respectively. For the Fl 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 Fl 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.
EPA/OW/OST/HECD V-71
-------
Drinking Water Criteria Document for Brominated Acetic Acids
The developmental toxicity of DBA has been reported in two related abstracts. Narotsky et
al. (1996) studied the developmental toxicity of DBA in CD-I 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.
In a second published abstract, DBA was administered to CD-I 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
EPA/OW/OST/HECD V-72
-------
Drinking Water Criteria Document for Brominated Acetic Acids
delayed parturition at 24 mg/kg/day and above, the second abstract showed no developmentally-
adverse 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
other half was used for harvesting of proximal cauda-epididymis sperm for artificial insemination
of LHRH-synchronized females. Decreased body weight throughout the reproductive-development
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
EPA/OW/OST/HECD V-73
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 GDI5 throughout life. The average daily
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
EPA/OW/OST/HECD V-74
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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
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
EPA/OW/OST/HECD V-75
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 Fl 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-I 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-I 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
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;
EPA/OW/OST/HECD V-76
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Christian et al., 2002); as previously noted, the significant decreases in the body weight gain in
pups of both the P and Fl 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 (LC50), embyronic malformations (EC50),
minimum concentration to inhibit growth (MCIG), and a teratogenicity index (TI - the ratio of the
LC50 to the EC50). 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 LC50 nor the EC50 could be estimated. In the third test, the LC50 and EC50 without metabolic
activation were 7,354 and 11,723 mg/L, respectively. With metabolic activation, the LC50's for
three tests were 6,244, 69, and 3,787 mg/L; the reasons for the low LC50 in the second test were
EPA/OW/OST/HECD V-77
-------
Drinking Water Criteria Document for Brominated Acetic Acids
unclear and a pooled estimate was not calculated. The EC50, 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 typhimurium
strain TA100 in the Ames-fluctuation test. This assay is a modification of the Ames test in which
EPA/OW/OST/HECD V-78
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 |_ig/mL without S9 activation, and at 0.3 to 300 |_ig/mL in the presence
of S9 activation. No mutagenic effect was detected in the absence of S9 activation. The study
authors indicated that 10 |_ig/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 i-ig/mL. The decrease in positive mutagenic responses at the high doses (with S9
metabolic activation) was consistent with the onset of cytotoxicity at 100 |_ig/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 |_imole) 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.
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
EPA/OW/OST/HECD V-79
-------
Drinking Water Criteria Document for Brominated Acetic Acids
induction of DNA repair. In this assay, Escherichia coli strain PQ37 was exposed to
concentrations ranging from 1 to 1000 |_ig/mL MBA without metabolic activation, and from 3 to
3000 i-ig/mL with metabolic activation by S9 mix. Toxic concentrations were 300 |_ig/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 i-ig/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 |_im (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 DNA-
strand breaks was increased compared to controls immediately after the 1-hour treatment, and
EPA/OW/OST/HECD V-80
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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.
Table V-12. Genotoxicity Studies of MBA
Endpoint
Assay system
Results
(wo/w
activation)
Comments
Reference
EPA/OW/OST/HECD
V-81
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Gene mutation-
bacteria
Clastogenicity
DNA damage
Salmonella typhimurium
Salmonella typhimurium
TA100
Salmonella typhimurium
TA98, TA100
Micronuclei in
Pleurodeles waltl larvae
(Newt) erythrocytes
Escherichia coli strain
PQ37 SOS chromotest
Single strand breaks in L-
1210 mouse leukemia
cells
+
-/+
-/+
-/NT
-/-
+/NT
Detailed results were not
available
Tested to cytotoxic doses
in Ames fluctuation
protocol
Positive in TA98 and
TA100 in suspension
assay. Data provided in a
published abstract
None
Tested to cytotoxic doses
MBA was shown to be
cytotoxic to L-1210 cells
from 50 uM.
NTP, 2000a
Giller etal., 1997
Kohan et al.,
1998
Giller etal., 1997
Giller etal., 1997
Stratton et al.,
1981
NT = Not tested
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
EPA/OW/OST/HECD
V-82
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 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
EPA/OW/OST/HECD V-83
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 water-
intake value of 0.25 L/kg/day (U.S. EPA, 1988), are 25, 125, and 500 mg/kg/day. No dose-related
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-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
EPA/OW/OST/HECD V-84
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 |_ig/mL
without S9 activation, and from 10 to 10,000 |_ig/mL in the presence of S9 fraction. Toxic
concentrations were 1000 |_ig/mL without, and 10,000 |_ig/mL with, metabolic activation.
Genotoxicity of DBA was detected at concentrations of 10 to 750 |_ig/mL without activation, and at
30 to 3000 i-ig/mL with activation.
Similar results were reported for the microsuspension Ames assay of DBA reported in a
published abstract by Kohan et al. (1998). DBA (2.0 |_imole) induced a positive mutagenic
response in strains TA98 and TA100 +S9 at subtoxic concentrations (personal communication).
EPA/OW/OST/HECD V-85
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 |_ig/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
1-ig/mL DBA without metabolic activation, and to 3 to 10,000 |_ig/mL with S9 metabolic activation.
Toxic concentrations were reported as 1000 |_ig/mL without and 10,000 |_ig/mL with S9 activation.
DBA induced a positive response regardless of metabolic activation. The concentrations that
induced DNA repair were 200 to 750 |_ig/mL without activation, and 100 to 3000 |_ig/mL with
activation. Mayer et al. (1996) presented a scheme for the concentration and analysis of water
samples for trace analytes, and coupled it with the umu Microtest, which measures induction of
EPA/OW/OST/HECD V-86
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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.
EPA/OW/OST/HECD V-87
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 |_ig/mL, in the absence of S9)
EPA/OW/OST/HECD V-88
-------
Drinking Water Criteria Document for Brominated Acetic Acids
significantly increased the number of erythrocytes with micronuclei. The co-clastogenic effects of
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 |-ig/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
EPA/OW/OST/HECD V-89
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 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.
EPA/OW/OST/HECD V-90
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table V-13. Genotoxicity Studies of DBA
Endpoint
Assay system
Results (wo/w
activation)
Comments
Reference
In vitro assays
Gene mutation-
bacteria
Clastogenicity
DNA damage
Salmonella typhimurium
Salmonella typhimurium
TA100
Salmonella typhimurium
TA98, TA100
Salmonella typhimurium
TA98 and TA100
Micronuclei in
Pleurodeles waltl larvae
(newt) erythrocytes
Not specified.
Escherichia coli strain
PQ37 SOS chromotest
umu Microtest
+
+/+
-/+
+/+
-/NT
+
+/+
+/+
Detailed results were not
availab le
Tested to cytotoxic doses in
Ames fluctuation protocol
Positive in TA98 and TA100 in
suspension assay. Data
provided in a published abstract
Positive in TA 100. Reference
in Japanese; only study
summary in English was
reviewed.
None
Co-clastogenic with mitomycin
C. Reference in Japanese; only
study summary in English was
reviewed.
Tested to cytotoxic doses
Reference in German; only
study summary in English was
reviewed.
NTP, 2000c
Giller et al,
1997
Kohan et al.,
1998
Saito et al.,
1995
Giller et al.,
1997
Sasaki and
Kinae, 1995
Giller et al.,
1997
Mayer et al.,
1996
In vivo assays
DNA damage
Oxidative DNA damage
mouse liver in vivo
Oxidative DNA damage
mouse liver in vivo
+
+
2- to 3-fold induction in SOHdG
levels
2- to 3-fold induction in SOHdG
levels
Austin et al.,
1996
Parrish et al.,
1996
EPA/OW/OST/HECD
V-91
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table V-14. Summary of Genotoxicity Data for Brominated Acetic Acids
Assay
Mutagenicity
SOS DNA repair induction
DNA damage
Chromosome damage
MBA
- S9
-
-
+ S9
+
-
±
-
BCA
- S9
+ S9
+
NTb
NT
+
NT
DBA
- S9
±a
+
+ S9
+
+
+
±
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).
EPA/OW/OST/HECD
V-92
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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
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 LD50 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
EPA/OW/OST/HECD V-93
-------
Drinking Water Criteria Document for Brominated Acetic Acids
decreased maternal weight gain, decreased live-fetus size, and an increased incidence of soft-tissue
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
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
EPA/OW/OST/HECD V-94
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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, 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 lexicologically 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
EPA/OW/OST/HECD V-95
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 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 LD50 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). Moser et al. (2004) 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
EPA/OW/OST/HECD V-96
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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, and a NOAEL could not be determined.
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 reproductive-
tract 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
EPA/OW/OST/HECD V-97
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 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 two-generation reproductive toxicity study (Chlorine
Chemistry Council, 2001; Christian et al., 2002), impaired spermatogenesis was observed in male
rats of the P and Fl 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 Fl
generation); abnormal pathology of the testes and epididymes was noted in some males of the Fl
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
EPA/OW/OST/HECD V-98
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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-P
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 post-
natal 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
EPA/OW/OST/HECD V-99
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 Fl or F2 pups exposed
during gestation and lactation to 650 ppm DBA. These findings were 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
EPA/OW/OST/HECD V-100
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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).
EPA/OW/OST/HECD V-101
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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
EPA/OW/OST/HECD VI-1 Draft, do not cite or quote
-------
Drinking Water Criteria Document for Brominated Acetic Acids
isolated neural-tube defects. Based on this more stringent case definition, a prevalence odds ratio
(FOR) 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 (FOR 1.2,
95% confidence interval 0.5-2.6) was found for cases when analyzed based on total haloacetic-acid
tertiles. The specific haloacetic acids that were measured as part of the total haloacetic acid-
exposure 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.
EPA/OW/OST/HECD VI-2 Draft, do not cite or quote
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Chapter 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
EPA/OW/OST/HECD VII-1
-------
Drinking Water Criteria Document for Brominated Acetic Acids
through alkylation of sulfhydryl groups. MBA reactions with amino groups in proteins have also
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 toxicologjc 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
EPA/OW/OST/HECD VII-2
-------
Drinking Water Criteria Document for Brominated Acetic Acids
the potential for MBA to adversely impact cellular macromolecules, but whether this ability is
responsible for MBA toxicity has not been clearly shown.
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-
EPA/OW/OST/HECD VII-3
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 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-I mouse embryos (early somite-stage conceptuses) were cultured for up to 24
hours with 300 |_im DBA, 300 |_im BCA, Bis I (a specific PKC inhibitor with previously defined
EPA/OW/OST/HECD VII-4
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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
revealed the accumulation of sub-Gl events (indicative of apoptosis) with BCA and staurosporine,
but not DBA, Bis I or Bis V. For BCA, sub-Gl events were particularly pronounced in the head
region but not in the heart. Although sub-Gl 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
EPA/OW/OST/HECD VII-5
-------
Drinking Water Criteria Document for Brominated Acetic Acids
(NTDs) at 1000 |_im, 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,
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).
EPA/OW/OST/HECD VII-6
-------
Drinking Water Criteria Document for Brominated Acetic Acids
The potential for developmental toxicity among haloacetic compounds, including the
mono-, di-, and tri-substituted fiuoro-, 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
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 >
trifiuoro > difiuoro. 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 >
fiuoro), 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 > difiuoro), although the model predicted more similar potencies of the
difiuoro- and the dichloro-compounds than were seen experimentally. However, the model was
EPA/OW/OST/HECD VII-7
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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
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.
EPA/OW/OST/HECD VII-8
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 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 DCA-
induced 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
EPA/OW/OST/HECD VII-9
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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.
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
EPA/OW/OST/HECD VII-10
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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
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
EPA/OW/OST/HECD VII-11
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 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
EPA/OW/OST/HECD VII-12
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 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
EPA/OW/OST/HECD VII-13
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 sec) to
pregnenolone, a precursor of progesterone. 3p-hydroxysteroid dehydrogenase (3P-HSD) catalyzes
the conversion of pregnenolone to progesterone, which is then converted through a series of
catalyzed steps to testosterone.
Balchak et al. (2000) conducted an in vitro study on the effect of DBA on female reproductive
activity of Sprague-Dawley rats. In this study, the rats were given drinking water doses of DBA at
concentrations equivalent to 0, 10, 30, 90, or 270 mg/kg/day for 14 days and estrous cyclicity was
monitored during treatment and for an additional 2-week post-treatment interval. A dose-related
alteration in cyclicity was observed at 90 and 270 mg/kg/day, which persisted through the post-
treatment monitoring in the high dose group. An in vitro exposure of preovulatory follicles to DBA
was then used to assess the influence of DBA on steroid release. To select a concentration for use,
a single oral exposure to 270 mg/kg was administered, and the mean blood levels of DBA were
EPA/OW/OST/HECD VII-14
-------
Drinking Water Criteria Document for Brominated Acetic Acids
determined over a 5 hour interval. For the in vitro part of the study, pairs of preovulatory follicles
from PMSG-primed immature rats were exposed to 0 or 50 mg/mL DBA over a 24 hour period and
evaluated for estradiol and progesterone release under baseline and human chorionic gonadotropin
(hCG)-stimulated conditions. The influence of tumor necrosis factor (TNFa) exposures under
these conditions was also determined. In the non-stimulated condition, DBA was found to increase
the release of estradiol, but had no detectable effect on estradiol release in the animals primed with
hCG. Progesterone, however, showed marked suppression under hCG stimulation following
exposure to DBA, while non-stimulated secretion was unaffected. TNFa by itself also suppressed
stimulated progesterone release, but had no additional effect in combination with DBA. The data
suggest that one factor in the disruption in estrous cyclicity could be an alteration in steroid
production, which was characterized by separate effects on both estradiol and progesterone
secretion.
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 |_ig/mL DBA, but not with 2 or 10
1-ig/mL, under both baseline and hCG-stimulated conditions. The suppression in progesterone
release at 50 i-ig/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
EPA/OW/OST/HECD VII-15
-------
Drinking Water Criteria Document for Brominated Acetic Acids
were observed. In a second experiment, follicular cultures were supplemented with pregnenolone
to evaluate any effects of DBA (50 |_ig/mL) on 3p-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, 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
EPA/OW/OST/HECD VII-16
-------
Drinking Water Criteria Document for Brominated Acetic Acids
exposure to DBA may have an effect on the StAR-mediated transport of cholesterol within the
mitochondrial membrane and an effect on receptor or post-receptor events triggered by hCG, but
only at high doses.
Subsequently, Bodensteiner et al. (2004), exposed four groups of female Dutch-belted rabbits to
daily doses of 0, 1, 5, or 50 mg DBA/kg body weight in drinking water beginning in utero from
gestation day 15 throughout life to determine if DBA affects ovarian folliculogenesis. Functionality
of the endocrine axis was assessed by measuring serum concentrations of gonadotropins following
an intramuscular injection of 10 |_ig gonadotropin releasing hormone (GnRH) at 12 (prepubertal;
n=6/dose group) and 24 (postpubertal; n=10/dose group) weeks of age. A day after GnRH
challenge, the number of ovulation sites and ovarian weights were determined at necropsy. Left
ovaries were processed for histopathology, serially sectioned at 6 |_im, and every twelfth section
stained with hematoxylin and eosin was evaluated. All healthy follicles were categorized as
primordial, primary, small preantral (cavity), large preantral, or small antral follicles. The area of
each section evaluated was measured and the number of follicles in each category expressed per
mm2 unit area. In prepubertal animals, DBA caused a reduction in number of primordial follicles
(p< 0.05) and total healthy follicles (p< 0.05) at 50 mg/kg dose level. In adult animals, there were
fewer primordial follicles in both the 5 (p< 0.01) and 50 (p= 0.1) mg/kg dose groups. No profound
changes in gonadotropin profiles were observed. Although chronic exposure to DBA did not
appear to have an effect on late follicular development or ovulation, DBA reduced the population
EPA/OW/OST/HECD VII-17
-------
Drinking Water Criteria Document for Brominated Acetic Acids
of primordial follicles. The authors concluded that the long-term health consequences of
diminished primordial follicles are unknown, but it is very likely that reproductive senescence
would occur earlier. The 5 mg/kg/day seemed to be an effect levels in this study but it is difficult
to determine if the observed effect can be classified as adverse.
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.
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
EPA/OW/OST/HECD VII-18
-------
Drinking Water Criteria Document for Brominated Acetic Acids
(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
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.
EPA/OW/OST/HECD VII-19
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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-I 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 GDI5 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
(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 Fl generation was attributed to a general
EPA/OW/OST/HECD VII-20
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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.
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. GST-
EPA/OW/OST/HECD VII-21
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Zeta 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 involved
in glycogen degradation) represent another group that may be more susceptible to DBA or BCA
toxicity.
EPA/OW/OST/HECD VII-22
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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, anti-
oxidant 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.
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 whole-
embryo culture system. As described above, the results of Andrews et al. (1999b) in this in vitro
EPA/OW/OST/HECD VII-23
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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
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
EPA/OW/OST/HECD VII-24
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 development at doses
below those causing overt toxicity, and spermatotoxicity at doses causing overt systemic toxicity
EPA/OW/OST/HECD VII-25
-------
Drinking Water Criteria Document for Brominated Acetic Acids
(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 haloacetic acids acting on SP22 and other
sperm proteins may indirectly compromise androgen-dependent maintenance of spermatogenesis
(Klinefelter et al., 2002b).
EPA/OW/OST/HECD VII-26
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 whole-embryo culture system. It should
be noted, however, that although the findings from whole-embryo 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.
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
EPA/OW/OST/HECD VII-27
-------
Drinking Water Criteria Document for Brominated Acetic Acids
no multi-generation studies have been conducted for MBA or BCA. In addition, no data on age-
dependent 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 maybe 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
EPA/OW/OST/HECD VII-28
-------
Drinking Water Criteria Document for Brominated Acetic Acids
source of increased susceptibility. All of these possibilities remain speculative, and none has been
tested directly in in vivo studies.
EPA/OW/OST/HECD VII-29
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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.I. Quantification of Noncarcinogenic Effects
A.I.I. 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-
EPA/OW/OST/HECD VIII-1
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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
UFxMF
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 quantitative rating.
EPA/OW/OST/HECD VIII-2
-------
Drinking Water Criteria Document for Brominated Acetic Acids
The total uncertainty factor is the product of the component uncertainties. The individual
components of the uncertainty are as follows:
UFH 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.
UFA 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).
UFS An additional factor of 1, 3, or 10 used when extrapolating from less-than-
chronic 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.
UFL 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.
UFD 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 1A Iog10 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 UFD.
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
EPA/OW/OST/HECD VIII-3
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 ofnoncarcinogenic 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:
EPA/OW/OST/HECD VIII-4
-------
Drinking Water Criteria Document for Brominated Acetic Acids
DWEL = (RfD) x 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, noncarcinogenie endpoint of toxicity. The HAs are derived as follows:
HA = NOAEL or LOAEL x BW
UFxWI
where:
NOAEL or LOAEL = No- or lowest-observed-adverse-effect-level in mg/kg bw/day
EPA/OW/OST/HECD VIII-5
-------
Drinking Water Criteria Document for Brominated Acetic Acids
BW = Assumed body weight of a child (10 kg) or an adult (70 kg)
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 x RSC
where:
DWEL = Drinking water equivalent level
EPA/OW/OST/HECD VIII-6
-------
Drinking Water Criteria Document for Brominated Acetic Acids
RSC = Relative source contribution. The fraction of the total exposure
allocated to drinking water.
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-of-
evidence narrative. Additionally, more than one conclusion maybe reached for an agent (e.g., an
agent is "likely to carcinogenic" by inhalation exposure and "not likely to be carcinogenic" by oral
exposure.
EPA/OW/OST/HECD VIII-7
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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.
EPA/OW/OST/HECD VIII-8
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 biologically-based
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 maybe 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
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
-------
Drinking Water Criteria Document for Brominated Acetic Acids
nature of the target organ system(s) examined and the actual rate of exposure of the internal 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.
B. Noncarcinogenic Effects
B.I 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
Linder et al.,
1994a
Species
Sprague
Dawley
Rat
(male)
Route
Oral
Gavage
in water
Exposure
Duration;
Doses
Acute
Single dose;
100 to 200
mg/kg
Endpoints
Lethality, clinical
observation
NOAEL
LOAEL
(mg/kg/day)
-
LD50177
mg/kg
Comments
Doses not
specified.
EPA/OW/OST/HECD
VIII-10
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Reference
Linder et al.,
1994a
Randall et al.,
1991
Linder et al.,
1994a
Species
Sprague
Dawley
Rat
(male)
Long-
Evans Rat
(female)
Sprague
Dawley
Rat
(male)
Route
Oral
Gavage
in water
Oral
Gavage
in water
Oral
Gavage
in water
Exposure
Duration;
Doses
Acute
Single dose;
0, 100
mg/kg
Gestation
day 6-15;
0, 25,50,
100
mg/kg/day
14 day;
0, 25
mg/kg/day
Endpoints
Sperm analysis,
reproductive-tract
histopathology
Decreased maternal
weight gain,
decreased live-fetus
size, increased
incidence of soft-
tissue malformations
Sperm analysis,
reproductive-tract
histopathology
NOAEL
LOAEL
(mg/kg/day)
100
50
25
-
100
-
Comments
None
Published
abstract does
not provide
adequate
details for
definitive
review.
Only listed
endpoints
were
evaluated
B.I.I One-Day Health Advisory for MBA
The oral toxicity data for MBA are very limited. Linder et al. (1994a) reported an LD50 of
177 mg/kg in Sprague-Dawley rats. Clinical signs included excess drinking, hypomobility, labored
breathing, and diarrhea. However, LD50 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
EPA/OW/OST/HECD
VIII-11
-------
Drinking Water Criteria Document for Brominated Acetic Acids
approximate LD01 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 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
EPA/OW/OST/HECD VIII-12
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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-
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.I.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
EPA/OW/OST/HECD VIII-13
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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.
Table VIII-2. Summary of Oral Studies of BCA Toxicity
Reference
NTP, 1998
Luft et al.,
2000
Species
Sprague
Dawley
Rat
(male
and
female)
C57BL/6
Mouse
(male)
Route
Drinking
Water
Oral
Gavage
in water
Exposure
Duration^
Doses
14 day;
0, 3, 10, 28,
41 mg/kg/day
14 day;
0, 8,24, 72,
216 mg/kg/day
Endpoints
Clinical
observation,
body weight,
body weight
gain
Decrease in
mean number
of litters per
male, decreased
percent
of litters per
bred female
NOAEL
LOAEL
(mg/kg/ day)
41
24
-
72
Comments
None
Published
abstract does not
provide adequate
details for
definitive review.
Results of
histopathology
analysis were not
reported.
EPA/OW/OST/HECD
VIII-14
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table VIII-2. Summary of Oral Studies of BCA Toxicity
Reference
Klinefelter
et al.,
(2002a),
manuscript
Klinefelter
et al.,
(2002a),
manuscript
Parrish et
al., 1996
Species
Sprague-
Dawley
Rat
(male)
Sprague-
Dawley
Rat
(male)
B6C3F1
Mouse
(male)
Route
Oral,
Gavage
in water
Oral,
Gavage
in water
Drinking
water
Exposure
Duration^
Doses
14 day,
0, 24,72, 216,
mg/kg/day
(range-finding
study)
14 day,
0, 8,24, 72
mg/kg/day
(dose-response
study)
21 days;
0, 25, 125,
500 mg/kg/day
Endpoints
Decreased
serum hormone
levels, sperm
abnormalities,
altered
spermiation
Decreased
progressive
sperm motility,
significantly
reduced
fertility,
decrease in
sperm protein
SP22.
Increased liver
weight,
increased
oxidative DNA
damage
NOAEL
LOAEL
(mg/kg/ day)
-
-
125
24
8
500
Comments
The LOAEL was
the lowest dose
tested.
Decreased serum
hormone levels
not observed.
Decrease in sperm
protein SP22
paralleled the
reduction in
fertility.
The LOAEL was
the lowest dose
tested.
500 mg/kg/day is
considered a
marginal LOAEL.
Doses were
calculated from
default water-
intake estimates.
EPA/OW/OST/HECD
VIII-15
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table VIII-2. Summary of Oral Studies of BCA Toxicity
Reference
NTP,
1998
Species
Sprague
Dawley
Rat
(male)
Route
Drinkin
g water
Exposure
Doses
30 or 26
days; 0, 5,
15,39
mg/kg/day
Endpoints
Liver weight
and
histopatholog
y, sperm
quality
NOAEL
LOAEL
(mg/kg/ day)
15
39
Comments
Marginal
LOAEL; males
had marginal
liver weight and
histopathology
changes
EPA/OW/OST/HECD
VIII-16
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table VIII-2. Summary of Oral Studies of BCA Toxicity
Reference
NTP, 1998
Species
Sprague
Dawley
Rat
(female)
Route
Drinking
water
Exposure
Duration;
Doses
Group A:
34 days peri-
conception (12
days pre-
mating up to 5
days
cohabitation,
and up to 2 1
days
gestation); 0,
6, 19,50
mg/kg/day
Group C:
30 days peri-
conception (12
days pre-
mating 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
Endpoints
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 toxic ity
NOAEL
LOAEL
(mg/kg/ day)
19
19
61
50
50
-
Comments
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 post-
implantation loss
was not
statistically
significant and
lacked a dose-
EPA/OW/OST/HECD
VIII-17
-------
Drinking Water Criteria Document for Brominated Acetic Acids
EPA/OW/OST/HECD VIII-18
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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 non-statistically 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 endpoints were not
EPA/OW/OST/HECD VIII-19
-------
Drinking Water Criteria Document for Brominated Acetic Acids
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.
EPA/OW/OST/HECD VIII-20
-------
Drinking Water Criteria Document for Brominated Acetic Acids
B.3 Dibromoacetic Acid
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
mg/kg/day
Comments
General Toxicity Studies (by duration of treatment)
Linder et
al., 1994a
Parrish et
al., 1996
NTP, 1999
Sprague-
Dawley rat
(male)
B6C3F1
mouse
(male)
B6C3F1
mouse
(female)
Oral
Gavage
in water
Drinking
water
Drinking
water
Acute
single dose;
1000 to 2000
mg/kg
21 days; 0,25,
125, 500
mg/kg/day
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
Lethality, clinical
observation
Increased liver
weight, oxidative
DNA damage
Decreased
antibody-forming
cell response
--
25
38
LD501737
mg/kg
125
70
Doses not specified.
Doses were
calculated from
default water intake
estimates
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.
EPA/OW/OST/HECD
VIII-21
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table VIII-3. Summary of Oral Studies of DBA Toxicity
Reference
Moser et
al. (2004)
Species
F344 rat
(male and
female
adolescents)
Route
Drinking
water
Exposure
Duration; Doses
6 months,
0, 20,72, 161
mg/kg/day
Endpoints
Neuromuscular
and
neurobehavioral
abnormalities in
all dosed groups,
spinal cord
neuro pathology in
mid- and high-
dose groups
NOAEL
LOAEL
mg/kg/day
-
20
Comments
Published abstract.
Neurotoxicity and
neuro pathology only
end points examined.
Reproductive Toxicity Studies (by duration of treatment)
Linder et
al., 1994a
Vetter et
al., 1998
Sprague-
Dawley
rat (male)
Crl:CD
(SD)Br rat
(male)
Oral
Gavage
in water
Oral
Gavage
in water
Acute
single dose, up to
28-day recovery;
0 or 1250 mg/kg
Acute
single dose;
0, 600, 1200 mg/kg
Reproductive-
organ weight
changes,
decreased serum
testosterone,
sperm -quality
changes,
reproductive-
tract
histopathology
Testes
histopathology
-
-
1250
600
A single dose level
was used.
Sperm analysis was
limited to motility
and membrane
permeability, with no
adverse effects
reported.
EPA/OW/OST/HECD
VIII-22
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table VIII-3. Summary of Oral Studies of DBA Toxicity
Reference
Cummings
and Hedge,
1998
Linder et
al., 1994b
Linder et
al., 1995
Linder et
al., 1995
and Linder
et al.,
1997a
Species
Holtzman
rat (female)
Sprague-
Dawley rat
(male)
Sprague-
Dawley rat
(male)
Sprague-
Dawley rat
(male)
Route
Oral
Gavage
in water
Oral
Gavage
in water
Oral
Gavage
in water
Oral
Gavage
in water
Exposure
Duration; Doses
Gestation days 1-8;
0, 62.5, 125,250,
500 mg/kg/day
14 daily doses; 0,
10, 30,90, or 270
mg/kg
2, 5,9, 16,31, or
42 days;
0, 250 mg/kg/day
Up to 79 days;
0, 2, 10, or 50
mg/kg/day
42 days; 250
mg/kg/day
Endpoints
Clinical
observation
Reproductive-
tract
histopathology
Decreased
reproductive
performance on
Day 8-14 mating
and day 15-21
mating .
Decreased fertility
of male sperm day
16and31.
Altered sperm
parameters
beginning on day
9.
Reproductive-
tract
histopathology
NOAEL
LOAEL
mg/kg/day
250
-
-
2
500 (PEL)
10
250
10
(equivocal)
Comments
Reproductive
parameters were not
affected at 250
mg/kg/day or less
and were not
measured in the high
dose group due to
overt toxicity.
None
None
Histopathology
analysis presented in
Linder etal., 1997
Effects at LOAEL
became significant at
31 days.
EPA/OW/OST/HECD
VIII-23
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table VIII-3. Summary of Oral Studies of DBA Toxicity
Reference
Veerama-
chaneni et
al., 2000
Christian et
al., 1999
Species
Dutch-
belted
rabbits
(male)
Sprague-
Dawley rat
(male and
female)
Route
Drinking
water
Drinking
water
Exposure
Duration; Doses
Dams gestation
days 15 through
life, male offspring
through 24 weeks;
0,0.97, 5.05, 54.2
mg/kg/day
Sires from study
day (SD) 1-70: 10,
20, 36,66
mg/kg/day;
Dams from SD 1-
15: 15,30, 49,82
mg/kg/day; from
gestation day (GD)
0-21: 15, 39,49,
82 mg/kg/day;
from lactation day
(LD) 1-29: 44,87,
151, 212
mg/kg/day
Endpoints
At 24 weeks,
fertility of sperm
tested in
artificially
inseminated does
Slight
nonsignificant
decrease in
mating
performance and
number of mated
pairs at highest
dose tested;
reduced body wt
gain, body wt,
water
consumption,
food intake
attributed to taste
aversion of DBA-
treated water
NOAEL
LOAEL
mg/kg/day
nd
66
(sires)
> 60
(dams)
>82
(develop-
mental)
0.97
-
Comments
Published abstract
does not provide
adequate details for
definitive review.
Mean daily water
intake and
corresponding mean
DBA daily doses
significantly
increased in pregnant
and lactating females
EPA/OW/OST/HECD
VIII-24
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table VIII-3. Summary of Oral Studies of DBA Toxicity
Reference
Chlorine
Chemistry
Council,
2001;
Christian et
al., (2002)
Species
Sprague-
Dawley rat
(male and
female)
Route
Drinking
water
Exposure
Duration; Doses
Two -generation
reprod. toxicity
study, 0, 50,250,
650 ppm in water;
estimated daily
doses are:
P males - SD 1-92:
0, 4.4, 22,52
mg/kg/day;
P females - SD 1-
70: 0,6, 28,69
mg/kg/day; GD 0-
21: 0,6, 30,76
mg/kg/day; LD 1-
15: 0, 12, 56, 132;
Fl males: 0, 5-6,
22-30; 55-75
mg/kg/day;
Fl fern ales:
Premating: 0,7,
32, 83 mg/kg/day;
GDO-21: 0,6, 29,
67 mg/kg/day; LD
1-15:0, 10,50,
115 mg/kg/day
Endpoints
Impaired
spermatogenesis,
testicular
histopathology in
P and Fl males;
no treatment-
related effects in
females
Reduced body wt
gain, body wt,
water
consumption,
food intake in P,
Fl, F2 animals
and changes in
organ weights in
P and Fl
attributed to taste
aversion effects of
DBA -treated
water
NOAEL
LOAEL
mg/kg/day
4
(P males)
> 5
(Fl males)
22
(P males)
>22
(Fl males)
Comments
Study report was
reviewed by an
independent
scientific advisory
panel
Recently published in
International Journal
of Toxicology
21:237-276,2002.
Developmental Toxicity Studies
EPA/OW/OST/HECD
VIII-25
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Table VIII-3. Summary of Oral Studies of DBA Toxicity
Reference
Narotsky et
al., 1996
Narotsky
etal., 1997
Species
CD-I
mouse
(female)
CD-I
mouse
(female)
Route
Oral
Gavage
in water
Oral
Gavage
in water
Exposure
Duration; Doses
Gestation days 6-
15;
0, 24,50, 100,200,
392, 610, or 806
mg/kg/day
Gestation days 6-
15;
0, 50, 100, or 400
mg/kg/day
Endpoints
Increased
postnatal
mortality;
decreased pup
weight, tail
defects
Fetal
malformations;
hydronephrosis
NOAEL
LOAEL
mg/kg/day
392
50
610
100
Comments
Published abstract
does not provide
adequate details for
definitive review.
Maternal toxicity was
limited to decreased
maternal motor
activity at the high
dose.
Published abstract
does not provide
adequate details for
definitive review.
Hydronephrosis and
renal agenesis at 400
mg/kg/day.
B.3.1 One-Day Health Advisory for DBA
The acute oral toxicity data for DBA are very limited. Linder et al. (1994a) reported an
LD50 of 1737 mg/kg in Sprague-Dawley rats. However, LD50 studies are not suitable for the
development of One-day health advisories. In the same paper, Linder et al. (1994a) assessed the
effects of a single oral dose of 0 or 1250 mg/kg/day on the male reproductive system. After
EPA/OW/OST/HECD
VIII-26
-------
Drinking Water Criteria Document for Brominated Acetic Acids
dosing, the animals were followed for up to 28 days. The single dose tested was spermatotoxic,
and induced severe changes in sperm-quality parameters and reproductive-tract histopathology.
The absence of dose-response data limits the utility of this study for health-advisory derivation. In
a similar acute-dosing study validating a new test method, Vetter et al. (1998) also evaluated the
spermatotoxicity of DBA in male Crl:CD(SD)Br rats given single oral doses of 0, 600, or 1200
mg/kg DBA. The high dose, but not the low dose, resulted in clinical observations of toxicity and
testes histopathology, but no effects on sperm motility, morphology, or cell-membrane
permeability; analysis was limited to evaluation of these measures. Due to the limited dose-
response, and the testing for a limited number of endpoints, this study is not suitable for deriving a
One-day health advisory.
B.3.2 Ten-Day Health Advisory for DBA
A number of toxicity studies have been reported for DBA that are of suitable duration for
derivation of a Ten-day health advisory. These studies have evaluated reproductive- and
developmental-toxicity endpoints, as well as some indices of systemic toxicity.
Two studies by Narotsky and colleagues (Narotsky et al., 1996; Narotsky et al., 1997)
observed adverse developmental effects, including skeletal and soft-tissue malformations, in the
EPA/OW/OST/HECD VIII-27
-------
Drinking Water Criteria Document for Brominated Acetic Acids
offspring of pregnant mice administered DBA by gavage on GD 6-15. Full study reports have not
been published, so these studies are not appropriate for the derivation of a health advisory.
Male fertility and sperm parameters were also evaluated in groups of rats administered 0 or 250
mg/kg/day DBA by gavage for 2-42 days (Linder et al., 1995), but only a single high dose was
tested. Linder et al. (1994b) identified an equivocal LOAEL of 10 mg/kg/day based on
histopathological changes of the seminiferous tubules in adult male Sprague-Dawley rats
administered 14 daily gavage doses of DBA; aNOAEL could not be determined. In a study of
female reproductive function and fetal development, gavage doses of up to 125 mg/kg/day on GD
1-8 had no effects on reproductive parameters or clinical observations of toxicity in rats
(Cummings and Hedge, 1998). The highest dose tested in this study (500 mg/kg/day) was lethal
and not evaluated for reproductive outcome. None of these reproductive studies are appropriate for
the derivation of a Ten-day health advisory for a 10-kg child because the findings are only relevant
to sexually- mature animals and not to children.
Two other studies were of an appropriate duration for derivation of a Ten-day health
advisory, but these studies did not include evaluation of a complete array of systemic endpoints.
Parrish et al. (1996) evaluated the ability of DBA to induce oxidative DNA damage in the livers of
mice treated with DBA in drinking water for 21 days. Increased liver weight and levels of 8-
OHdG, a measure of oxidative stress, were observed at 125 mg/kg/day, but the absence of
histopathology or clinical chemistry data makes it unclear whether the observed increase in liver
EPA/OW/OST/HECD VIII-28
-------
Drinking Water Criteria Document for Brominated Acetic Acids
weight was adverse. Other organ systems and end points were not evaluated. In an
immunotoxicity-screening assay comprised of 4 short-term studies (NTP,1999), female mice
treated with DBA in the drinking water for 28 days exhibited an array of immunotoxic effects;
however, many of these effects were inconsistent and/or did not exhibit a dose-response. Based on
decreased spleen IgM antibody-forming response, the NOAEL was 38 mg/kg/day and the LOAEL
was 70 mg/kg/day. Thymus and spleen weights were also evaluated, but other organs and end
points were not assessed. Therefore, neither of these studies (Parish et al, 1996; NTP, 1999) are
considered to be suitable for derivation of a Ten-day health advisory in the absence of other
toxicity studies that have adequately evaluated systemic toxicity.
B.3.3 Longer-Term Health Advisory for DBA
A number of studies that examined the reproductive or developmental toxicity of DBA
were evaluated for the potential derivation of a Longer-term HA. As described previously,
published abstracts are available for two developmental-toxicity studies in mice (Narotsky et al.,
1996, 1997), which demonstrated adverse developmental effects including increased postnatal
mortality, and skeletal (tail defects) and soft-tissue (kidney defects) malformations. A published
abstract for a neurotoxicity study in rats has shown that DBA produces neurobehavioral toxicity,
including neuromuscular abnormalities, decreased sensorimotor responsiveness, and increased
motor activity, as well as spinal cord neuropathology indicative of axonal degeneration (Moser et
EPA/OW/OST/HECD VIII-29
-------
Drinking Water Criteria Document for Brominated Acetic Acids
al., 2004). However, none of these studies have been published and, thus, these results are not
suitable for the derivation of human-health advisories.
Linder et al. (1995, 1997a) evaluated the spermatotoxicity and fertility in male Sprague-
Dawley rats administered daily gavage doses of DBA of 0, 2, 10, or 50 mg/kg/day for up to 79
days. In a companion study, male rats were gavaged with either 0 or 250 mg/kg/day daily for 42
days, at which time dosing was terminated due to severe overt toxicity. Fertility in the dosed males
was assessed through day 213 by mating treated males with untreated females at different time
periods. Based on the results of these studies, DBA is clearly spermatotoxic and effects on sperm
histopathology appear to be the most sensitive endpoint, because these effects are observed in the
absence of other reproductive toxicity endpoints. Changes in retention of Step 19 spermatids was
the only effect that occurred at the lowest dose. This effect was equivocally noted following
repeated dosing with 10 mg/kg/day, but not 2 mg/kg/day, for 31 or 79 days. However, the
biological significance of this finding for a Longer-term human-health advisory is unclear because
changes in sperm count, morphology, and motility were observed at higher doses (50 mg/kg/day)
than those associated with these early and mild histopathological changes, and male fertility was
significantly affected only at 250 mg/kg/day (Linder et al., 1995). At doses of 50 mg/kg/day and
lower, there were no significant effects on reproductive outcome as indicated by a number of
different measures and indices of successful mating behavior were not significantly altered. These
results are described in detail in Section V. Although it has been proposed that the fertility of
EPA/OW/OST/HECD VIII-30
-------
Drinking Water Criteria Document for Brominated Acetic Acids
rodents may be less sensitive to changes in sperm count than fertility in humans (U.S. EPA, 1996a;
Zenick et al., 1994), human data are highly variable and generally inconsistent across studies.
Further, there may be significant differences in the susceptibility of different species and rodent
strains to DBA-induced reproductive toxicity. In an acute spermatotoxicity study by Vetter et al.
(1998), changes in sperm motility and morphology were not observed in male Crl:CD(SD)BR rats
at acute gavage doses of up to 1200 mg/kg, whereas Linder et al. (1994a) observed significant
alterations in both of these parameters in male Sprague-Dawley rats following an acute gavage
dose of 1250 mg/kg. A published abstract by Veeramachaneni et al. (2000) reported that male
Dutch-belted rabbits exposed in utero from GD 15 through lactation and post-weaning for 24
weeks exhibited decreased fertility, as evidenced by reduced conception in females artificially
inseminated with sperm from treated animals, at drinking-water doses as low as 0.97 mg/kg/day.
However, a full report of this study has not been published, and thus these results cannot be
comprehensively evaluated. It is also not known whether humans would be less, or more, sensitive
to DBA-induced male reproductive-tract toxicity than rats or rabbits. No reproductive
epidemiologic data on DBA are available, and comparative in vitro studies have not been
conducted.
Although the studies by Linder et al. (1994a, 1994b, 1995, 1997) adequately characterize
the male reproductive hazards in rats repeatedly administered DBA by oral gavage, these studies
are not considered to be suitable for quantitative dose-response assessment in the absence of
EPA/OW/OST/HECD VIII-31
-------
Drinking Water Criteria Document for Brominated Acetic Acids
sub chronic and chronic toxicity studies that have adequately evaluated DBA systemic toxicity. In
the recent two-generation reproductive/developmental toxicity study conducted by the (Chlorine
Chemistry Council, 2001; Christian et al., 2002), impaired spermatogenesis was also observed in
male rats of the P and Fl 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 not less than 22 mg/kg/day for
the Fl generation); abnormal pathology of the testes and epididymes was noted in some males of
the Fl generation at 650 ppm (equivalent to a LOAEL of not less than 75 mg/kg/day). The
corresponding NOAEL was 4 mg/kg/day in the P generation and at least 5 mg/kg/day in the Fl
generation. 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) (Chlorine Chemistry Council, 2001; Christian
et al., 2002). Alternatively, these studies in combination may define a NOAEL/LOAEL boundary
for functional effects of DBA on reproduction. No treatment-related adverse developmental effects
other than impaired spermatogenesis were noted in either males or females of the Fl and F2
generations. To date, other developmental toxicity studies have only been published in abstract
form and thus cannot be comprehensively evaluated until full study reports are available for
review. Subchronic and chronic toxicity testing of DBA is planned or in progress (NTP, 2000c).
EPA/OW/OST/HECD VIII-32
-------
Drinking Water Criteria Document for Brominated Acetic Acids
A number of additional studies are currently ongoing. Therefore, it is not appropriate to develop a
Longer-term health advisory at this time.
B.3.4. Reference Dose and Drinking Water Equivalent Level for DBA
As discussed in the previous section on the Longer-term health advisory, the toxicity
database for DBA is currently limited, and there are no suitable studies of appropriate design and
duration to derive an oral RfD at this time.
C. Carcinogenic Effects
No epidemiology or animal studies were identified to develop a quantitative cancer-risk
assessment for MBA, BCA, or DBA. No studies were identified that directly evaluated the human
carcinogenicity of MBA, BCA, or DBA. 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
cancer (Boorman et al., 1999), but brominated acetic acids have not been specifically implicated.
C.I. Monobromoacetic acid
EPA/OW/OST/HECD VIII-33
-------
Drinking Water Criteria Document for Brominated Acetic Acids
No epidemiology or animal studies were identified that evaluated the carcinogenicity of
MBA. The data are inadequate for determining whether MBA is genotoxic. The mutagenicity of
MBA might be metabolism-dependent, based on different results for these compound in the
presence or absence of microsomal activation in S. typhimurium strain TA100. MBA was
reported to be mutagenic in three independent studies in bacteria (Giller et al, 1997; Kohan et al.,
1998; NTP 2000a), but MBA did not induce DNA repair as measured by the SOS chromotest
(Giller et al., 1997). Due to the limited database, there is insufficient evidence to determine the
genotoxicity of MBA.
Following the EPA's 1986 (U.S. EPA, 1986) Guidelines for Cancer Risk Assessment,
MBA is best classified as Group D, "not classifiable as to human carcinogenicity ". This
classification is appropriate because no data are available on human or animal carcinogenicity.
Under the 1999 Draft Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999b), the data are
"inadequate for an assessment of human carcinogenic potential" of MBA.
C.2. Bromochloroacetic acid
No epidemiology studies have evaluated the carcinogenicity of BCA. The carcinogenicity
of BCA has not been tested in a full cancer bioassay. However, BCA is currently slated for testing
(NTP, 2000b). The only carcinogenicity-testing data in animals for BCA that was identified was
EPA/OW/OST/HECD VIII-34
-------
Drinking Water Criteria Document for Brominated Acetic Acids
reported in a published abstract (Stauber et al., 1995). The abstract reported preliminary data
suggesting that BCA induces hepatic tumors in B6C3F1 mice. However, no experimental details
were provided in the brief study summary and the full study report has not been published. BCA
was reported as positive in the single standard assay identified, a Salmonella reverse-mutation
assay (NTP, 2000b). The reports of Austin et al. (1996) and Parrish et al. (1996) demonstrated that
BCA treatment could induce oxidative DNA damage in the livers of treated mice. While these
data are suggestive of genotoxic potential, BCA has not been sufficiently tested to make a
determination as to its genotoxicity.
Following the EPA's 1986 Guidelines for Carcinogen Risk Assessment, BCA is best
classified as Group D, "not classifiable as to human carcinogenicity" (U.S. EPA, 1986). This
classification is appropriate because no data are available on human carcinogenicity and there are
only preliminary animal carcinogenicity data. Under the 1999 Draft Guidelines for Carcinogen
Risk Assessment (U.S. EPA, 1999b), the data are "inadequate for an assessment of human
carcinogenic potential" of BCA.
C.3. Dibromoacetic acid
No epidemiology studies have evaluated the carcinogenicity of DBA. The carcinogenicity
of DBA has not been tested in a full cancer bioassay. However, DBA is currently undergoing
EPA/OW/OST/HECD VIII-35
-------
Drinking Water Criteria Document for Brominated Acetic Acids
testing (NTP, 2000c). 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. Experimental details are not available for either of these studies because neither
has been published in peer-reviewed journals, but the findings of So and Bull (1995) might be of
particular significance since colon cancer has been implicated as a potential cancer site in humans
exposed to drinking-water disinfectant by-products, including haloacetic acids (Boorman et al.,
1999).
Much of the concern for the potential carcinogenicity of DBA arises from the demonstrated
high-dose rodent-liver tumorigenicity of its chlorinated analog, dichloroacetic acid (DCA). The
ability of both compounds to induce a similar spectrum of noncarcinogenic toxic effects suggests
the possibility that this might also be the case for carcinogenic effects. For example, both
compounds are potent spermatotoxicants and induce a similar spectrum of effects in the male
reproductive tract (Linder et al., 1997b). Both compounds also affect the liver. Treatment with
DBA or DCA resulted in increased liver weight, although only DBA increased the formation of
oxidative DNA damage in this study (Parrish et al., 1996). DBA and DCA also appear to have
similar kinetics (Schultz et al., 1999), but insufficient data are available on DBA metabolism, renal
elimination, and tissue distribution to fully compare the kinetics of DBA and DCA. These
similarities between DBA and DCA suggest that, like DCA, DBA might also be tumorigenic at
high drinking-water doses administered for a lifetime. However, the weight-of-evidence for DCA
EPA/OW/OST/HECD VIII-36
-------
Drinking Water Criteria Document for Brominated Acetic Acids
genotoxicity indicates that DCA is nongenotoxic except possibly at high doses that also induce
cytotoxicity. In contrast, although the DBA database for genotoxicity/ mutagenicity is more
limited than that for DCA, the weight-of-evidence to date indicates that DBA is genotoxic Thus,
although both DBA and DCA exhibit similar toxicokinetics and similar systemic and reproductive
toxicity, they may well differ in their carcinogenic potential and mode(s) of carcinogenic action.
Thus, insufficient data are available to assess DBA carcinogenic hazard, and DBA is
classified as Group D, "not classifiable as to human carcinogenicity" under the 1986 Carcinogen
Risk Assessment Guidelines. Under the 1999 Draft Guidelines for Carcinogen Risk Assessment
(U.S. EPA, 1999b), the data are "inadequate for an assessment of human carcinogenic potential"
of DBA.
D. Summary
In the absence of a comprehensive toxicity database, no adequate studies of suitable design
and/or duration were identified to serve as the basis for any health advisories for MBA, BCA, or
DBA.
MBA, BCA, and DBA are all classified as "not classifiable as to human carcinogenicity "
under the 1986 Carcinogen Risk Assessment Guidelines , and "inadequate for an assessment of
EPA/OW/OST/HECD VIII-37
-------
Drinking Water Criteria Document for Brominated Acetic Acids
human carcinogenic potential" under the 1999 Draft Guidelines for Carcinogen Risk Assessment
(U.S. EPA, 1999b).
EPA/OW/OST/HECD VIII-38
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Chapter IX. References
Anderson, W.B., P.O. Board, B. Gargano and M.W. Anders. 1999. Inactivation of glutathione
transferase zetaby dichloroacetic acid and other fluorine-lacking alpha-haloalkanoic acids. Chem.
Res. Toxicol. 12(12): 1144-1149.
Andrews, J.E., J. Schmidt, H. Nichols, E.S. Hunter and G. Klinefelter. 1999a. Developmental
toxicity of structurally related disubstituted haloacetic acids in embryo culture. Toxicol. Sci. 48(1
suppl): 16.
Andrews, J.E., J. Schmid, H. Nichols, E.S. Hunter and G. Klinefelter. 1999b. Development
toxicity of mixtures: the water disinfection byproducts dichloro-, dibromo-, and bromochloro-
acetic acid in embryo culture. Teratology. 59(6): 384.
Arora, H., M.W. LeChavallier, and K.L. Dixon. 1997. DBF Occurrence Survey. J. AWWA
89(6): 60-68.
Austin, E.W., J.M. Parrish, D.H. Kinder and R.J. Bull. 1996. Lipid peroxidation and formation of
8-hydroxydeoxyguanosine from acute doses of halogenated acetic acids. Fundamental and Applied
Toxicology. 31:77-82.
EPA/OW/OST/HECD IX-1
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Balchak, S.K., Hedge, J.M., Murr, A.S., Mole, M.L., Goldman JM. 2000. Influence of the drinking
water disinfection by-product dibromoacetic acid on rat estrous cyclicity and ovarian follicular
steriod release in vitro. Reprod Toxicol. Nov-Dec: 14(6):533-9.
Blackburn, A.C., H.F. Tzeng, M.W. Anders and P.O. Board. 2000. Discovery of a functional
polymorphism in human glutathione transferase zeta by expressed sequence tag database analysis.
Pharmacogenetics. 10(1): 49-57.
Bodensteiner K.J., Sawyer H.R., Moeller C.L., Kane C.M.,. Pau K-Y.F., Klinefelter G.R,
Veeramachaneni D.N.R. 2004. Chronic Exposure to DBA, a Water Disinfection By-product,
Diminishes Primordial Follicle Populations in the Rabbit. Toxicological Sciences 80, 83-91.
Boorman, G.A., V. Dellarco, J.K. Dunnick, R.E. Chapin, S. Hunter, F. Hauchman, H. Gardner, M.
Cox and R.C. Sills. 1999. Drinking water disinfection byproducts: Review and approach to
toxicity evaluation. Environ. Health Perspect. 107(Suppl. 1): 207-217.
Bull, R. J. 2000. Mode of action of liver introduction by trichloroethylene and its metabolites,
trichloroacetate and dichloroacetate. Environ. Health Perspect. 108(Suppl. 2): 241-259.
EPA/OW/OST/HECD IX-2
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Chlorine Chemistry Council (unpublished). 2001. Oral (Drinking Water) Two-Generation
Reproductive Study of Dibromoacetic Acid (DBA) in Rats. R.G. York, Study Director, Argus
Research Laboratory, Horsham, PA 19044. Study Protocol No. 2403-005, 1546 pages.
Christian, M.S., York, R.G., Hoberman, A.M., Diener, R.M., Fischer, L.C., Gates, G.A. 2001.
Biodisposition of Dibromoacetic Acid (DBA) and Bromodichloromethane (BDCM) Administered
to Rats and Rabbits in Drinking Water during Range-Finding Reproduction and Developmental
Toxicity Studies. Int. J. Toxicol. 20:239-253.
Cicmanec, J.L., L.W. Condie, G.R. Olson and S.R. Wang. 1991. 90-day toxicity study of
dichloroacetate in dogs. Fundamental and Applied Toxicology. 17(2): 376-389.
Christian, M.S. York, R.G. Hoberman, A.M., Frazee J., Fisher, L.C., Brown,W.R. Creasy, D.M.
2002 Oral (drinking water) two-generation reproductive toxicity study of dibromoacetic acid
(DBA) in rats. Int. J. Toxicol. 21(4):237-76
Cornett, R., M. James, G. Henderson, J. Cheung, A. Shroads and P. Stacpoole. 1999. Inhibition of
glutathione S-transferase and tyrosine metabolism by dichloroacetate: A potential unifying
mechanism for its altered biotransformation and toxicity. Biochemical and Biophysical Research
Communications. 262: 752-756.
EPA/OW/OST/HECD IX-3
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Cummings, A.M. and J.M. Hedge. 1998. Dibromoacetic acid does not adversely affect early
pregnancy in rats. Reproductive Toxicology. 12(4): 445-448.
Dawson, D.A., Bantle, J.A. 1987. Development of a reconstituted water medium and preliminary
validation of the frog embryo teratogenesis assay-Xenopus (FETAX). J. Appl. Toxicol. 1987 Aug;
7(4)237-44.
DeAngelo, A. B., F. B. Daniel, L. McMillan, P. Wernsing, and R. E. Savage. 1989. Species and
strain sensitivity to the induction of peroxisome proliferation by chloroacetic acids. Toxicol. Appl.
Pharmacol. 101:285-298.
Dourson, M.L. 1994. Methods for establishing oral reference doses (RfDs). In: Risk Assessment
of Essential Elements, W. Mertz, C.O. Abernathy and S.S. Olin, Ed. ILSI Press, Washington, DC.
pp. 51-61.
Dourson, M.L., L. Knauf and J. Swartout. 1992. On reference dose (RfD) and its underlying
toxicity database. Tox. Ind. Health. 8(3): 171-189.
EPA/OW/OST/HECD IX-4
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Eriksson, L., R. Berglind and M. Sjostrom. 1994. A multivariate quantitative structure-activity
relationship for corrosive carboxylic acids. Chemometrics and Intelligent Laboratory Systems. 23:
235-245.
Gardner, H.S. and M.W. Toussant. USACEHR drinking water disinfection byproduct testing with
FETAX: Bromodichloromethane, dibromoacetic acid, and chlorinated surface water. U.S. Army
Center for Environmental Health Research. October 15, 1999.
Giller, S., F. Le Curieux, F. Erb and D. Marzin. 1997. Comparative genotoxicity of halogenated
acetic acids found in drinking water. Mutagenesis. 12(5): 321-328.
Goldman, J.M. and Murr, A.S. 2002. Alterations in ovarian follicular progesterone secretion by
elevated exposures to the drinking water disinfection byproduct dibromoacetic acid: examination
of the potential site(s) of impact along the steriodogenic pathway. Toxicology 171:83-93.
Goldman, J.M. and Murr, A.S. 2003. Dibromoacetic acid-induced elevations in circulating
estradiol: effects in both cycling and ovariectomized/steroid-primed female rats. Reproductive
Toxicology (RTX) 5542: 1-8. In press.
EPA/OW/OST/HECD IX-5
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Gonzalez,-Leon, A., J.L. Merdink, R.J. Bull and I.R. Schultz. 1999. Inhibition of metabolism by
chlorinated and brominated di-haloacetates and differential recovery in B6C3F1 mice and F344
rats. Toxicol. Sci. 48(Suppl. 1): 206.
Gorlatov, S. N. and T. C. Stadtman. 1998. Human thioredoxin reductase from HeLa cells:
selective alkylation of selenocysteine in the protein inhibits enzyme activity and reduction with
NADPH influences affinity to heparin. Proc. Natl. Acad. Sci. U.S.A. 95(15): 8520-8525.
Hansch, C., Leo, A., D. Hoekman. 1995. Exploring QSAR - Hydrophobic, Electronic, and Steric
Constants. Washington, DC: American Chemical Society. 3.
Hunter, E.S., E.H. Rogers, J.E. Schmid and A. Richard. 1996. Comparative effects of haloacetic
acids in whole embryo culture. Teratology. 54: 57-64.
Hunter, E.S. and E.H. Rogers. 1999. Dysmorphogenic effects of three metabolites of haloacetic
acids in mouse embryo culture. Teratology. 59(6): 402.
Ito, K., D. Akiyama and N. Minamiura. 1994. Evidence for an essential histidine residue on active
site of human urinary DNase I: carboxymethylation and carbethoxylation. Arch. Biochem.
Biophys. 313(1): 126-130.
EPA/OW/OST/HECD IX-6
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Jacangelo, J.G., N.L. Patania, K.M. Reagan, E.M. Aieta, S.W. Krasner and M.J. McGuire. 1989.
Ozonation: assessing its role in the formation and control of disinfection by-products. J Am Water
Works Assn. 81:74-84.
James, M.O., Zimeng, Y., Cornett, R. et al. 1998. Pharmacokinetics and Metabolism of
[14C]Dichloroacetate in Male Sprague-Dawley Rats. Drug, Metabolism and Disposition.
26(11):1143-1143.
Jeffrey, C.J. 1999. Monlighting proteins. Trends in Biochem. Sci. 24:8-11.
Kato-Weinstein, J., M.K. Lingohr, G.A. Orner, B.D. Thrall and R.J. Bull. 1998. Effects of
dichloroacetate on glycogen metabolism in B6C3F1 mice. Toxicol. 130: 141-154.
Kennedy, C.K., K.B. Cohen, W.E. Bechtold, I.Y. Chang, A.F. Eidson, A.R. Dahl and R.F.
Henderson. 1993. Effect of dose on the metabolism of 1,1,2,2-tetrabromoethane in F344/N rats
after gavage administration. Toxicol. Appl. Pharmacol. 119: 23-33.
Klinefelter, G., L. Strader, J. Suarez, N. Roberts, M. Holmes andL. Mole. 2000. Dibromoacetic
acid, a drinking water disinfection by-product, alters male reproductive development and fertility.
EPA/OW/OST/HECD IX-7
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Paper presented at the annual meeting for the Society for the Study of Reproduction, Madison,
Wisconsin.
Klinefelter G.R,. L.F Strader, J.D. Suarez, and N.L. Roberts. 2002a. Bromochloroacetic acid
exerts qualitative effects on rat sperm: implications for a novel biomarker. Toxicological Sciences
68: 164-173.
Klinefelter, G.R, Welch, J.E., Perrault, S.D., Moore, H.D., Zucker, R.M., Suarez, J.D., Roberts,
N.L., Bobseine, S. Jeffay. (2002b). Localization of the sperm protein SP22 and inhibition of
fertility in vitro and in vivo. J. Androl., in press.
Klotz, J.B. and L. A. Pyrch. 1999. Neural tube defects and drinking water disinfection by-
products. Epidemiology. 10: 383-390.
Kohan, M.J., G. Huggins-Clark and S.E. George. 1998. Mutagenicity of chlorinated and
brominated acetic acids. 29th Annual Meeting of the Environmental Mutagen Society, Anaheim,
California, March 21-26, 1998. Environ. Mol. Mutagen. 31: 36.
Krasner, S.W., M.J. McGuire, J.G. Jacangelo, N.L. Patania, K.M. Reagan and E.M. Aieta. 1989.
The occurrence of disinfection by-products in U.S. drinking water. J. Am. Water Works Assn.
81:41-53.
EPA/OW/OST/HECD IX-8
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Lide, D.R., eds. 1999. CRC Handbook of Chemistry and Physics, 73rd Edition. Boca Raton, FL:
CRC Press.
Lin, E. L., J. K. Mattox, and F. B. Daniel. 1993. Tissue distribution , excretion, and urinary
metabolites of dichloroacetic acid in the male Fischer 344 rat. J. Toxicol. Environ. Health.
38(l):19-32.
Linder, R.E., G.R. Klinefelter, L.F. Strader, D.N.R. Veeramachaneni, N.L. Roberts and J.D.
Suarez. 1997a. Histopathologic changes in the testes of rats exposed to dibromoacetic acid.
Reproductive Toxicology. 11(1): 47-56.
Linder, R.E., G.R. Klinefelter, L.F. Strader, M.G. Narotsky, J.D. Suarez and N.L. Roberts. 1997b.
Spermatotoxicity of dichloroacetic acid. Reproductive Toxicology. 11(5): 681-688.
Linder, R.E., G.R. Klinefelter, L.F. Strader, M.G. Narotsky, J.D. Suarez, N.L. Roberts and S.D.
Perreault. 1995. Dibromoacetic acid affects reproductive competence and sperm quality in the
male rat. Fundamental and Applied Toxicology. 28:9-17.
Linder, R.E., G.R. Klinefelter, L.F. Strader, J.D. Suarez and C.J. Dyer. 1994a. Acute
spermatogenic effects of bromoacetic acids. Fundamental and Applied Toxicology. 22: 422-430.
EPA/OW/OST/HECD IX-9
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Linder, R.E., G.R. Klinefelter, L.F. Strader, J.D. Suarez, N.L. Roberts and C.J. Dyer. 1994b.
Spermatotoxicity of dibromoacetic acid in rats after 14 daily exposures. Reproductive Toxicology.
8(3): 251-259.
Lipscomb, J. C., D. A. Mahle, W. T. Brashear, and H. A. Barton. 1995. Dichloroacetic acid:
metabolism in cytosol. Drug Metab. Dispo. 23: 1202-1205.
Luft, J.C., J.B. Garges, J.R. Rockett, and D.J. Dix. 2000. Male reproductive toxicity of
bromochloroacetic acid in mice. Paper presented at the annual meeting for the Society for the
Study of Reproduction, Madison, Wisconsin.
Lykins, Jr., B.W., W.E. Koffskey, and K.S. Patterson. 1994. Alternative disinfectants for drinking
water treatment. J. Environ. Eng. 120(4):745-758.
Marhaba, T.F. and D. Van. 2000. The variation of mass and disinfection by-product formation
potential of dissolved organic matter fractions along a conventional surface water treatment plant.
J. Hazardous Materials A74: 133-147.
Mayer, P., R. Hirsch, G. Reifferscheid and K. Haberer. 1996. Development of a joint sample
pretreatment for the successive determination of organic compounds and genotoxicity in
chlorinated water. Vom Wasser. 86: 305-320.
EPA/OW/OST/HECD IX-10
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Michal, G. (Ed.). 1999. Biochemical Pathways. John Wiley and Sons, Inc. New York.
Miltner, R.J., E.W. Rice and A. A. Stevens. 1990. Pilot-scale investigation of the formation and
control of disinfection byproducts. In: 1990 Annual Conference Proceedings. AWWA Annual
Conference, Cincinnati, OH. 2:1787-1802.
Narotsky, M.G., B.T. Hamby and D.S. Best. 1997. Developmental effects of dibromoacetic acid
(DBA) in a Segment II study in mice. Teratology. 55(1):67.
Narotsky, M.G., B.T. Hamby, D.S. Best and E.S. Hunter. 1996. In vivo developmental effects of
dibromoacetic acid (DBA) and dichloroacetic acid (DCA) in mice (Abstract). Teratology 53(2):
96-97.
NAS. 1980. National Academy of Sciences. Drinking water and health, Vol.3. Washington, DC:
National Academy Press.
NAS. 1977. National Academy of Sciences. Drinking water and health. Washington, DC:
National Academy Press.
Nieuwenhuijsen, M.J., M.B. Toledano, N.E. Eaton, J. Fawell and P. Elliott. 1999. Occup.
Environ. Med. 57: 73-85.
EPA/OW/OST/HECD IX-11
-------
Drinking Water Criteria Document for Brominated Acetic Acids
NIOSH 2000. National Institute for Occupational Safety and Health. The Registry of Toxic Effects
of Chemical Substances (RTECS). Updated June, 2000 and downloaded October, 2000.
NIOSH. 1990. Unpublished provisional data as of 7/1/90, National Occupational Exposure Survey
(1981-83). Cincinnati, OH: National Institute for Occupational Safety and Health.
NTP. 2000a. Water disinfection byproducts (bromoacetic acid). Available on-line at http://ntp-
server.niehs.nih.gov/htdocs/Results_Status/Resstatb/M920034.html.
NTP. 2000b. Water disinfection byproducts (bromochloroacetic acid). Available on-line at
http://ntp-server.niehs.nih.gov/htdocs/Results_Status/Resstatw/M980085.html.
NTP. 2000c. Water disinfection byproducts (dibromoacetic acid). Available on-line at http://ntp-
server.niehs.nih.gov/htdocs/Results_Status/Resstatw/M960093.html.
NTP. 1999. Final range-finding report. Immunotoxicity of dibromoacetic acid in female B6C3F1
mice. Submitted to the National Toxicology Program. Final Report. Richmond, VA: White. K.L.
and Munson, J.A. Medical College of Virginia.
NTP. 1998. Bromochloroacetic acid: Short term reproductive and developmental toxicity study
when administered to Sprague-Dawley rats in the drinking water. Submitted to the National
EPA/OW/OST/HECD IX-12
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Toxicology Program. Final Report. Gaithersburg, Maryland: R.O.W. Sciences, Inc. R.O.W.
Sciences Study No. 7244-301; NTP/NIEHS No.: NOl-ES-75409;NTP-RDGTNo.: 96-001. Also
available as NTIS# PB98-172414. http://ntp-server.niehs.nih.gov/htdocs/RDGT-
studies/RDGT96001 .html.
Parrish, J.M., E.W. Austin, O.K. Stevens, D.H. Kinder and R.J. Bull. 1996. Haloacetate-induced
oxidative damage to DNA in the liver of male B6C3F1 mice. Toxicology. 110: 103-111.
Moser, V.C., Phillips, P.M., Levine, A.B., McDaniel, K.L., Sills, R.C., Jortner, B.S., and Butt,
M.T. 2004. Neurotoxicity Produced by DBA in Drinking Water of Rats. Toxicological Sciences
79,112-122.
Pourmoghaddas, H., A.A. Stevens, R.N. Kinman, R.C. Dressman, L.A. Moore and J.C. Ireland.
1993. Effect of bromide ion on formation of HAAs during chlorination. J. Am. Water Works
Assn. 85: 82-87.
Randall, J.L., S.A. Christ, P. Horton Perez, G.A. Nolen, E.J. Read and M.K. Smith. 1991.
Developmental effects of 2-bromoacetic acid in the Long Evans rat. Teratology. PI22: 454.
Reckhow, D.A. and P.C. Singer. 1990. Chlorination by-products in drinking water: from formation
potentials to finished water concentrations. J. AWWA 82(4): 173-180.
EPA/OW/OST/HECD IX-13
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Reckhow, D.A., P.C. Singer and R.L. Malcolm. 1990. Chlorination of humic materials: byproduct
formation and chemical interpretations. Environ. Science and Technology 24(11):
Reinoso, R.F., Telfer, B.A. and Rowland, M. 1997. Tissue water content in rats measurd by
desiccation. J. Am. Water Works Assoc. 85:82-87.
Richard, A.M. and E.S. Hunter, in. 1996. Quantitative structure-activity relationships for the
developmental toxicity of haloacetic acids in mammalian whole embryo culture. Teratology. 53:
352-360.
Richardson, S.D. 1998. Identification of drinking water disinfection byproducts. In: John Wiley's
Encyclopedia of Environmental Analysis & Remediation, R.A. Meyers Ed. 3:1398-1421.
Saito, H., S. Isoda, M. Kato and N. Nagaoka. 1995. Mutagenic activity of indoor swimming pool
water. Environmental Mutagen Research Communications. 17(2): 169-177.
Sasaki, Y.F. and N. Kinae. 1995. Co-clastogenic effects of water pollutants. Environmental
Mutagen Research Communications. 17: 65-74.
EPA/OW/OST/HECD IX-14
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Schultz, I.R., J.L. Merdink, A. Gonzalez-Leon and R.J. Bull. 1999. Comparative Toxicokinetics
of Chlorinated and Brominated Haloacetates in F344 Rats. Toxicol. Appl. Pharmacol. 158(2):
103-114.
Schultz, I.R., A. Gonzalez-Leon, J.L. Merdink, and R.J. Bull. 1998. Comparative toxicokinetics
and metabolism of halo-acetic acids in F344 rats. Toxicol. Sci. 42(lSuppL): 212.
Shapiro, R., D. J. Strydom, S. Weremowicz and B. L. Vallee. 1988. Sites of modification of
human angiogenin by bromoacetate at pH 5.5. Biochem. Biophys. Res. Commun. 156(1): 530-
536.
Sigma-Aldrich. 2000. www.sigma-aldrich.com.
So, B. and R.J. Bull. 1995. Dibromoacetate (DBA) acts as a promoter of abnormal crypt foci in
the colon of F344 rats. Toxicologist. 15(1): 232.
Stacpoole, P.W., G.N. Henderson, Z. Yan, R. Cornett and M.O. James. 1998. Pharmacokinetics,
metabolism, and toxicology of dichloroacetate. Drug Metab. Rev. 30(3): 499-539.
Stauber, A.L, A.B. DeAngelo, and R.J. Bull. 1995. Different modes of action of chlorinated and
brominated haloacetates. Toxicol. Sci. 15(Suppl. 1): 232.
EPA/OW/OST/HECD IX-15
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Stratton, C.E., W.E. Ross and S. Chapman. 1981. Cytotoxity and deoxyribonucleic acid damage
associated with bromoacetate. Biochem. Pharmacol. 30: 1497-1500.
Takahashi, K., Taira, T., Niki, T., Seino, C., Iguchi-Ariga, S., Ariga, H. 2001 DJ-1 positevely
regulates the androgen receptor by impairing the binding of PIASxa to the receptor. J. Biol. Chem.
276, 37556-37563.
The Merck Index. 1996. The Merck Index, 1 Oth ed. Windholz, M., S. Budavari, R.F. Blumetti
and E.S. Otterbein, eds. Rahway, NJ: Merck and Company. Inc.
Tong, Z., P.O. Board and M.W. Anders. 1998a. Glutathione transferase zeta-catalyzed
biotransformation of dichloroacetic acid and other alpha-haloacids. Chem. Res. Toxicol. 11:
1332-1338.
Tong, Z., P.O. Board and M.W. Anders. 1998b. Glutathione transferase Zeta catalyzes the
oxygenation of the carcinogen DCA to glyoxylic acid. Biochem. J. 331(2): 371-374.
U.S. EPA. 2000a. Information Collection Rule (ICR) database. Available online at
http://www.ecradlab.com/twg/mainmenu.htm Downloaded September, 2000.
U.S. EPA. 2000b. Stage 2 Occurrence and Exposure Assessment for Disinfectants and
EPA/OW/OST/HECD IX-16
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Disinfection Byproducts (D/DBPs) in Public Drinking Water Systems. Office of Ground Water
and Drinking Water.
U.S. EPA. 2000c. ICR Data Analysis Plan. Office of Water.
U.S. EPA. 2000d. Methodology for Deriving Ambient Water Quality Criteria for the Protection of
Human Health. Office of Water, Office of Science and Technology. EPA.
U.S. EPA. 1999a. U.S. Environmental Protection Agency. Preliminary assessment (1988 -
February, 1999) of Dibromoacetic acid (CAS No. 631-64-1). Submitted 5/19/99.
U.S. EPA. 1999b. U.S. Environmental Protection Agency. Guidelines for carcinogen risk
assessment. Draft. NCEA-F-0644.
U.S. EPA. 1996a. U.S. Environmental Protection Agency. Reproductive toxicity risk assessment
guidelines; notice. Federal Register. 61(212): 56274-56322.
U.S. EPA. 1996b. U.S. Environmental Protection Agency. Proposed guidelines for carcinogen
risk assessment. Office of Research and Development, Washington, DC. EPA/600/P-92/003C.
EPA/OW/OST/HECD IX-17
-------
Drinking Water Criteria Document for Brominated Acetic Acids
U.S. EPA. 1994. U.S. Environmental Protection Agency. Final draft for the drinking water
criteria document on chlorinated acids/aldehydes/ketones/alcohols. EPA contract no. 68-C2-0139
to Clement International. Office of Science and Technology, Office of Water.
U.S. EPA. 1988. U.S. Environmental Protection Agency. Recommendations for and
documentation of biological values for use in risk assessment. EPA 600/6-87-008. Cincinnati, OH:
Office of Health and Environmental Assessment, for the Office of Solid Waste and Emergency
Response.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for carcinogen risk
assessment. Federal Register. 51(185): 33992-34003.
Veeramachaneni, D.N.R., T.T. Higuchi, J.S. Palmer and C.M. Kane. 2000. Dibromoacetic acid, a
disinfection by-product in drinking water, impairs sexual function and fertility in male rabbits.
Paper presented at the annual meeting for the Society for the Study of Reproduction, Madison,
Wisconsin.
Vetter, C.M., J.E. Miller, L.M. Crawford, M.J. Armstrong, J.H. Clair, M.W. Conner, L.D. Wise
and T.R. Skopek. 1998. Comparison of motility and membrane integrity to assess rat sperm
viability. Reproductive Toxicology. 12(2): 105-114.
EPA/OW/OST/HECD IX-18
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Ward, K.W., E.H. Rogers and E.S. Hunter, IE. 2000. Comparative pathogenesis of haloacetic acid
and protein kinase inhibitor embryotoxicity in mouse whole embryo culture. Toxicological
Sciences. 53: 118-126.
Ward, K.W., E.H. Rogers and E.S. Hunter, III. 1998. Pathogenesis of haloacetic acid-induced
embryotoxicity in mouse whole embryo culture. Toxicol. Sci. 42(1 suppl.): 260-61.
Ward, K.W., W.R. Mundy, E.H. Rogers and E.S. Hunter, HI. 1997. Effect of haloacetic acids on
embryonic protein kinase C activity during neurulation (abstract). Teratology. 55(1): 52.
Webster, K.E., P.M. Ferree, R.P. Holmes and S.C. Cramer. 2000. Identification of missense,
nonsense, and deletion mutations in the GRHPR gene in patients with primary hyperoxaluria type
II(PH2). Hum. Genet. 107:176-185.
Whitney, P. L. 1970. Inhibition and modification of human carbonic anhydrase B with
bromoacetate and iodoacetamide. Eur. J. Biochem. 16(1): 126-135.
WHO. 2000. Environmental Health Criteria: 216 Disinfectant By-products. World Health
Organization, Geneva. International Programme on Chemical Safety (IPCS).
EPA/OW/OST/HECD IX-19
-------
Drinking Water Criteria Document for Brominated Acetic Acids
Williams, D.T., P.M. Benoit and G.L. Lebel. 1998. Trends in levels of disinfection by-products.
Environmetrics. 9: 555-563.
Zenick, H., E.D. Clegg, S.D. Perreault, G.R. Klinefelter and L.E. Gray. 1994. Assessment of male
reproductive toxicity: a risk assessment approach. In: A.W. Hayes, ed. Principles and methods of
toxicology, 3rd edition. New York: Raven Press, Ltd. pp 937-988.
EPA/OW/OST/HECD IX-20
------- |