United States
Environmental Protection
Agency
Drinking Water Addendum to
the IRIS Toxicological Review of
Dichloroacetic Acid
-------
Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
Drinking Water Addendum
to the IRIS Toxicological Review
of
Dichloroacetic Acid
U.S. Environmental Protection Agency
Office of Water (43 04T)
Health and Ecological Criteria Division
Washington, DC 20460
www.epa.gov/safewater/
EPA Document Number: 822-R-05-009
Date: November, 2005
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
Table of Contents
FOREWORD vi
Chapter I. Executive Summary 1
A. Human Exposure 1
B. Quantification of Toxicological Effects 2
Chapter II. Human Exposure 4
A. Drinking Water Exposure 4
A. 1 National Occurrence Data for DCA 4
A. 1.1 ICR Background Information 4
A. 1.2 Quarterly Distribution System Average and Highest Values for
DCA 5
A.2 Factors Affecting the Relative Concentrations of DCA in Drinking Water
5
A.2.1 Disinfection Treatment 6
A.2.1.1 Disinfection Treatment Data from the ICR
Database 7
A.2.2 Bromide Concentration 9
A.2.2.1 Bromide Concentration Data from the ICR Database
10
A.2.3 Total Organic Carbon (TOC) Concentration 12
A.2.3.1 TOC Concentration Data from the ICR Database
12
A.2.4 Seasonal Shifts 12
A.2 A. 1 Seasonal Shift Data from the ICR Database 15
B. Exposure to Sources Other Than Drinking Water 15
B. 1 Dietary Exposure 16
B.2 Air Exposure 17
B.3 Dermal Exposure 17
C. Overall Exposure 17
C. 1 Body Burden 18
D. Summary 18
Chapter III. Quantification of Toxicological Effects 20
A. Introduction to Methods 20
A.I Quantification of Noncarcinogenic Effects 20
A. 1.1. Drinking Water Equivalent Level 22
in
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
A. 1.2. Health Advisory Values 22
A.2 Quantification of Carcinogenic Effects 24
B. Noncarcinogenic Effects 24
B. 1 One-day Health Advisory 24
B.2 Ten-day Health Advisory 25
B.3 Longer-term Health Advisory 28
B.4 Reference Dose, Drinking Water Equivalent Level, and Lifetime Health
Advisory 32
C. Carcinogenic Effects 35
C.I Characterization of Carcinogenic Potential 35
D. Summary 37
Chapter IV. References 39
IV
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
List of Tables
Table 1-1. Summary of Health Advisory Values for Drinking Water 2
Table II-l. Dichloroacetic Acid Quarterly Distribution System Average and Highest Value 6
Table 11-2. DCA by Disinfection Method (Quarterly Distribution System Average) 8
Table II-3. DCA by Influent Bromide Concentration
(Quarterly Distribution System Average) 11
Table II-4. DCA by Influent Total Organic Carbon (TOC) Concentration
(Quarterly Distribution System Average) 13
Table II-5. DCA by Sample Quarter (Quarterly Distribution System Average Means) 15
Table II-6. DCA in Foods 16
Table III-1. Summary of NOAELs and LOAELs Considered for Development of One- and Ten-Day
Drinking Water Health Advisories (HA) for DCA 26
Table III-2. Summary of NOAELs and LOAELs Considered for Development of Longer-Term
Drinking Water Health Advisory (HA), RfD, and DWEL for DCA 29
Table III-3. Summary of Development of the HAs and DWEL for DC A 38
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
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 nonenforceable 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.
The documents that provide the health effects data that support the derivation of the
MCLG for Dichloroacetic Acid (DC A) are the Toxicological Review of Dichloroacetic Acid and
the Integrated Risk Information System (IRIS) Summary for DCA. These documents provide a
summary of the available data on pharmacokinetics, acute and chronic toxicity to animals and
humans, epidemiology, and mechanisms of toxicity that were used to derive the Reference Dose
(RfD) and Cancer Slope Factor for DCA. Both documents are available on line at the following
internet sites:
Toxicological Review: http://www.epa.gov/iris/toxreviews/0654-tr.pdf.
IRIS Summary: http://www.epa.gov/iris/subst/0654.htm.
This document is a supplement to the IRIS materials. It provides information on the
concentrations of DCA found in drinking water, food and other media. Most importantly it
expands the risk information found in IRIS by providing Health Advisory (HA) guidelines that
pertain to exposures through drinking water for lifetime and shorter durations (1-day, 10-day,
and longer term, approximately 10% of an individual's lifetime). The HA 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. For the convenience of the reader, the
chapter on quantification of risk in this document also includes the RfD derivation and the
quantification of the cancer risk excerpted from the IRIS documentation. They are combined
with the HA values to provide a complete summary of the health-based guidelines that are
applicable to DCA exposure via drinking water.
VI
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
Chapter I. Executive Summary
A. Human Exposure
The Information Collection Rule (ICR) database (U.S. EPA, 2000a) contains extensive
information on concentrations of DCA in drinking-water systems, and how those concentrations
vary with input-water characteristics and treatment methods. The database contains information
for six quarterly samples collected between 7/97 to 12/98, from approximately 300 large systems
covering approximately 500 plants. As shown in the database and published evaluations, the
mean concentrations of DCA were consistently lower in disinfected groundwater than in
disinfected surface water. The mean concentrations of DCA were 4.7 and 13.93 |ig/L in
disinfected groundwater and disinfected surface water, respectively.
Examination of the data that there is no consistent difference in DCA levels among the
common disinfection methods (non-ozonation) for surface water and for groundwater. It is
noted that ozonation alone does not support the production of DCA, unless followed by a
chlorine-based disinfection treatment. DCA levels can be affected by the presence of other
compounds in the treated water and by seasonal changes. The input of bromine into treated
water results in a decrease in DCA because of the increased formation of brominated analogs.
Mean concentrations of DCA have been shown to be higher in treated surface water than in
treated groundwater at all influent bromide concentrations. DCA concentrations in treated water
increase in response to TOC (total organic carbon) and are highest in spring and summer and
lowest in fall and winter.
DCA has been detected in some processed and unprocessed foods. The use of chlorine in
food processing plants and/or cooking and/or rinsing the cooked food in water containing
disinfection by-products provides an opportunity for uptake of DCA into the food matrix.
Concentrations of DCA in food can be comparable to those in water and may contribute to the
overall dose. Data are absent regarding levels of DCA in ambient air. Nevertheless, volatility of
DCA from water is very low, and inhalation of DCA volatilized from tapwater is not expected to
be a significant source of exposure. Limited data are available on the DCA concentrations in
swimming pool water, but estimated dermal absorption rates and frequency of exposure indicate
that dermal absorption of DCA through swimming, or from showering or bathing, is not
expected to contribute a significant amount toward the total dose.
DCA and its metabolites are rapidly excreted and unlikely to bioaccumulate following
environmental exposure. Limited data on plasma levels of DCA and background urinary
excretion indicate a very low body burden.
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
B. Quantification of Toxicological Effects
No suitable studies were identified for derivation of the One-day HA. A No-Observed-
Adverse-Effect-Level (NOAEL) of 25 mg/kg/day for increased liver weight in mice given DCA
in drinking water for 21 days was used to derive a Ten-day HA of 3 mg/L for a 10-kg child.
This Ten-day HA was used as a conservative value for the One-day HA. A LOAEL of 12.5
mg/kg/day, based on testicular degeneration, brain vacuolation, and liver histopathology
observed in dogs given DCA in drinking water for 90 days, was used to derive a Longer-term
HA of 0.1 mg/L for a 10-kg child and 0.4 mg/L for a 70-kg adult. The same study was used to
calculate a DWEL of 0.1 mg/L. As a matter of EPA policy (U.S. EPA, 1989), Lifetime HAs are
not determined for chemicals categorized as known or likely human carcinogens. The Health
Advisory (HA) values for DCA are summarized in Table 1-1.
Table 1-1 Health Advisory Values for Dichloroacetic Acid
One-Day HA
3
Ten-Day HA
O
Longer-Term HA
Child
0.1
Adult
0.4
Life-Time HA
ND2
In 1994, U.S. EPA performed a cancer weight-of-evidence review for DCA (U.S. EPA,
1994), and this evaluation was updated in 1996. These reviews classified DCA as a Group B2
(probable human carcinogen) in accordance with the 1986 EPA Guidelines for Carcinogen Risk
Assessment. Based on current data and in accordance with the U.S. EPA (1999) Guidelines for
Carcinogen Risk Assessment, DCA is considered a likely human carcinogen.
The tumorigenic data from a DeAngelo et al. (1999) study in mice were utilized for
cancer dose-response modeling (U.S. EPA, 2003). This study was considered to be the most
suitable for quantification among several rodent carcinogenicity bioassays because it was
specifically designed to establish a multi-point dose-response curve. The study utilized five dose
groups plus a control, and the exposure duration covered the expected lifetime of the rodent
model (mouse). Based on a Benchmark Response (BMR) of 10% extra risk, the tumorigenic
data were fit to five different models. Most of the models resulted in estimates of the
Benchmark Dose (BMD10), ranging from 6.9 to 9.36 mg/kg/day. The preferred model was the
multistage model, which yielded a BMD10 of 6.9 mg/kg/day and a BMDL10 of 2.1 mg/kg/day.
For extrapolation to low doses below the range of experimental observation, the linear
approach is used by U.S. EPA (1999) as a matter of policy if the mode of carcinogenicity is not
well understood. Therefore, linear, low-dose extrapolation was performed, yielding a cancer
slope factor for DCA of 0.048 (mg/kg/day)"1. The concentrations in drinking water equivalent to
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
a E-4 (1 in 10,000), E-5 (1 in 100,000) and E-6 (1 in 1,000,000) risk are 70 |ig/L, 7.0 |ig/L and
0.7 |ig/L, respectively.
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
Chapter II. Human Exposure
DCA is often formed during the disinfection of drinking water using chlorine-containing
disinfectants. In addition, it is an experimental pharmaceutical used for the treatment of
congenital lactic acidosis and can be found in some research and development laboratories.
A. Drinking Water Exposure
DCA concentrations were measured in samples of disinfected drinking water collected
under the Information Collection Rule (ICR; U.S. EPA, 2000a). A cross- section of public water
systems across the United States were required by the U.S. EPA to collect samples of treated
water and measure the levels of selected disinfection byproducts. The following sections will
present DCA data from the ICR as well as similar information from public water systems
published by other researchers.
A. 1 National Occurrence Data for DCA
A. 1.1 ICR Background Information
Under the ICR, samples of finished drinking water were collected from surface- and
ground-water systems serving at least 100,000 persons. The analytical data were entered in the
ICR database which includes information gathered for 18 months from July 1997 to December
1998. The data in sections that follow were taken from the online version of the ICR database
(U.S. EPA, 2000a). The explanation of the 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).
The ICR generated plant-level sets of data that could link system water quality
parameters and treatment methods to the seasonal variability in the levels of DCA and other
disinfectant byproducts formed. The database contains the analytical results from 18 monthly or
six quarterly samples collected from approximately 300 large systems covering roughly 500
treatment plants. The samples were tested for DBF levels, and disinfectant residuals along with
influent and finished water-quality parameters (e.g., TOC, temperature, pH, alkalinity). Samples
were collected at several locations in the distribution system to cover a range of residence times
during which DBFs can form in the finished water. Over the 18-month period, 1473 samples
were taken from 305 plants with surface water as their source, and 583 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).
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
A. 1.2 Quarterly Distribution System Average and Highest Values for DCA
This section describes the approach employed for the analysis of data for water-quality
parameters, and for DCA concentrations under the ICR Program. 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.
The Quarterly Distribution System Average and Highest Value for DCA are presented in Table
II-1. All data in these tables are presented with two decimal points (as provided in the ICR
database), however, this does not necessarily represent the actual precision of the data.
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.
The quarterly distribution-system highest value is the highest of the four distribution-system
samples collected by a plant in a given quarter.
The mean concentrations of DCA (averaged across the four sampling locations) were 4.7
and 13.93 |ig/L in disinfected drinking water derived from groundwater and surface water,
respectively. The lower mean concentration for the treated ground water is associated with a
higher percentage of non-detects. In all cases, the non-detects from the ICR data base are treated
as zero in the calculation of the mean, median, standard deviation, plO and p90 values in the
tables that follow (U.S. EPA, 2000b).
A.2 Factors Affecting the Relative Concentrations of DCA in Drinking Water
Sections A.2.1 - A.2.4 contain information on the effects of disinfection chemicals,
influent bromide concentration, influent total organic carbon (TOC) concentration, and seasonal
shifts, respectively, on DCA concentrations. Published data from other studies of factors
influencing DBF formation are also included. In a number of cases, there is a considerable
difference between the mean and median and the p90 and maximum concentration for a given
data set, indicating a skewed distribution. The standards deviation are also often large in
comparison to the mean. Accordingly, caution should be used when weighing the statistical
findings for the binary comparisons of DCA levels with a source water or treatment factor.
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
Table II-l
Dichloroacetic Acid
Quarterly Distribution System Average and Highest Value
(ICR, U.S. EPA, 2000a)
Source
SW
GW
Quarterly
Dist. Sys.
Average
High
Average
High
Plants
305
305
123
123
N
1473
1473
583
583
PctND
%
0.81
0.81
44.25
44.25
Mean
ug/L
13.93
17.52
4.70
6.53
Median
ug/L
11.80
15.00
0.65
1.50
STD
ug/L
10.26
13.04
8.69
11.44
Min
ug/L
0.00
0.00
0.00
0.00
Max
ug/L
72.50
99.00
60.25
71.00
plO
ug/L
3.58
4.80
0.00
0.00
p90
//g/L
25.40
32.00
14.30
20.00
Source:
Quarterly Dist. Sys.:
Plants:
N:
PctND:
Mean:
Median:
STD:
Min:
Max:
plO:
p90:
SW: Surface Water; GW: Groundwater
Quarterly Distribution System (DS) Samples
Average - quarterly average of 4 locations inDS
High - highest of 4 locations in DS
Number of plants sampled
Number of samples
Percent samples non-detect (detection limits not provided)
Arithmetic mean of all samples
Median value of all samples
Standard deviation
Minimum Value
Maximum Value
lOthpercentile
90th percentile
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. The occurrence of
DPBs following various disinfection treatments is, therefore, useful information. Several
published studies (Boorman et al., 1999; Lykins et al., 1994; Miltner et al., 1990) report on the
formation of DCA and other DBFs under different disinfection conditions.
In their review, Boorman et al. (1999) compared the concentrations of different drinking-
water disinfection byproducts, including the chlorinated acetic acids, formed by chlorination,
ozonation, chlorine dioxide, and chloramination. Most of the data were from surface-water
systems that used chlorination. For the systems using chlorination, DCA with a median and
maximum concentration of 15 and 74 |ig/L, respectively, was present at the highest
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
concentrations. Qualitative, but not quantitative data were provided for other disinfection
systems. The principal products formed by chloramination were similar to those formed by
chlorination. Chlorine dioxide and ozonation treatments, when followed by chlorination and
chloramination in the distribution system, produced many of the same halogenated byproducts as
seen with chlorination. However, the chlorine dioxide and ozonation treatments produced them
at lower concentrations.
Similar results were obtained by Lykins et al. (1994) who predisinfected and
postdisinfected distribution-system water with either chlorine or chloramine. Chlorine produced
the highest concentration of halogenated DBFs and, in general, the concentrations could be
reduced by adding ozone as a predisinfectant with postchlorination. Chlorine alone, and ozone
followed by chlorine produced the highest DC A concentrations (60 |ig/L and 38 |ig/L,
respectively). Chloramine treatment, and ozone treatment followed by chloramine, resulted in
the lowest concentrations of DCA (9.2 and 5.6 |ig/L respectively). Miltner et al. (1990)
reported similar results after studying 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). The amount of DCA measured in finished water and in
simulated distribution waters was lower (at p = 0.05 by Student's t-test) when ozonation was
combined with chlorination or with chloramination than when chlorination was used alone. The
authors also reported that, because less free chlorine was available, the use of ozone and
chloramine resulted in the lowest formation of halogenated DBFs, including DCA.
Jacangelo et al. (1989) examined the impact of ozonation on the formation and control of
selected DBFs in drinking water at four utilities. Modifications were made on the treatment
process train at each full or pilot-scale plant to incorporate ozone in the treatment process.
Although the sample size did not allow for statistical analysis of the data (Jacangelo et al., 1989),
in general, treatment that employed ozonation followed by chloramination were the most
effective in reducing the formation of the monitored, halogenated DBFs. Two of the treatment
plants measured the concentration of DCA in the treated water before and after the inclusion of
ozone in the treatment process. At one plant, the concentration of DCA decreased from 9.4 |ig/L
to 4.7 |ig/L after ozonation was introduced. In the other plant there was little cange in the DCA
levels (23 jig/L before ozone and 21 jig/L after).
A.2.1.1 Disinfection Treatment Data from the ICR Database
Data on the concentrations of DCA were gathered from plants using several disinfection
treatments. Those chemical-disinfection treatments most commonly used (those affecting 10%
or more of the plants evaluated), along with the ozonation treatments, are presented in Table 11-2
below.
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
Table II-2
DCA by Disinfection Method
Quarterly Distribution System Average
(ICR, U.S. EPA, 2000a)
Source
SW
GW
Disinfection
Chemicals
C12/C12
C12 CLM/CLM
O3/C12
O3/CLM
/CL2
cyci2
OVCLM
Plants
180
66
7
10
67
39
1
N
818
305
25
49
302
169
6
PctND
%
0.98
0.33
0.00
0.00
64.90
33.14
0.00
Mean
>"g/L
14.17
15.31
6.78
8.59
2.02
2.97
11.97
Median
>«g/L
12.43
13.85
6.08
5.85
0.00
1.00
11.88
STD
>"g/L
10.54
9.40
3.40
7.01
4.58
5.64
1.68
Min
/"g/L
0.00
0.00
2.50
1.43
0.00
0.00
9.58
Max
J"g/L
72.50
71.05
12.18
29.50
23.68
42.00
14.50
plO
>«g/L
3.68
5.83
2.65
2.18
0.00
0.00
9.58
p90
>"g/L
25.25
26.30
11.48
19.98
6.25
7.95
14.50
Source: SW: Surface Water; GW: Groundwater
C12/C12: Free chlorine in Water Treatment Plant (WTP) and Distribution System (DS)
Cl2_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 non-detect (detection limits not provided)
Mean: Arithmetic mean of all samples
Median: Median value of all samples
STD: Standard deviation
Min: Minimum Value
Max: Maximum Value
plO: lOthpercentile
p90: 90th percentile
Mean concentrations of DCA in the ICR database were slightly higher in surface-water
plants using chlorine followed by chloramine than in those using free chlorine alone. This is
believed to be more a function of source water characteristics that the disinfectants selected for
treatment (McGuire and Hotaling, 2002, Obolensky et al., 2002). Where groundwater was the
source, the mean concentrations of DCA in plants with no disinfectant in the treatment plant and
with free chlorine in the distribution system were lower than the mean concentrations of the
same chemicals in plants with free chlorine in both the treatment plant and the distribution
system (Table II-2). The standard deviations indicate that there was considerable variance in the
DCA concentrations in the data set reducing confidence in this observation.
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
Only a very limited number of plants used ozonation in combination with either chlorine
or chloramine in the distribution system. Mean concentrations of DC A were significantly lower
in surface-water plants using ozone at the water-treatment plant with free chlorine or free
chloramine in the distribution system than in plants using common (non-ozonation) chemical-
disinfection processes (based on Student's t-test analysis of the data). This finding was also
presented by Lykins et al. (1994). However, there was no significant difference (p = 0.05) in
DCA concentrations between the two disinfectants used to maintain residuals in the distribution
system after ozonation in treating surface water.
Although it appears that the concentration of DCA was higher in the single plant in
which groundwater was treated with ozone/chloramine than in groundwater plants using
common disinfection-treatment methods, firm conclusions regarding trends are not possible
because of the fact that there were only data from one plant.
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 and 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.
There were no effects of pH on DCA formation between pHs 5 and 7; but there were
significant differences between pH 5 and 9.4 and pH7 and 9.4 by Tukey's test analysis.
However, DCA formation decreased significantly from pHs 5 and 7 to pH 9.4. Under all pH
conditions and reaction times, the concentration of DCA decreased rapidly with the incremental
addition of bromide ion but there was an increase in the brominated haloacids (Pourmoghaddas
etal., 1993).
In September 1987, the U.S. EPA'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 (Krasner et al., 1989). 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 US EPA study and
10 California water utilities in the CDHS study. DCA was among the DBFs measured in this
study. Chloride and bromide analyses were added to the protocol, beginning with the second
quarter (summer 1988) of sampling. Among the 35 facilities examined by AMWA, bromide
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
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 a decrease in the levels of the chlorinated acetic acids.
A.2.2.1 Bromide Concentration Data from the ICR Database
Table II-3 below presents data regarding the formation of DC A 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 301 plants using surface water
as the source (38%) 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. At all influent bromide concentrations, mean concentrations of
DC A were higher in treated surface water than groundwater.
A regression analysis of the ICR data indicates that there was no significant correlation
(at a = 0.05) between influent bromide concentrations and the mean concentrations of DCA in
surface water and groundwater (Table II-3). However, the standard deviations are large relative
to the mean value indicating considerable variability in the data and lowering the confidence in
the analysis. For surface water, the mean DCA concentration was lower with bromide levels
> 100 ppb, but relatively consistent across the lower bromide concentrations evaluated (all <100
ppb). The lower level of DCA at the higher bromide concentrations is consistent with the
concept that increasing bromide concentrations lower the chlorinated DBFs because of increased
formation of bromine-containing compounds. However, in the case of the ground water
samples, the highest mean DCA concentration was found for the plants with bromide
concentrations > 100 ppb. The distribution for these samples seems to be skewed by several
samples with high DCA concentrations causing the mean to be greater than the median. At least
two of the 208 samples had concentrations more than ten times greater than the median.
10
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
Table II-3
DCA by Influent Bromide Concentration
Quarterly Distribution System Average
(ICR, U.S. EPA, 2000a)
Source
SW
GW
Influent
Bromide
Concentration
(ppb)
= 100
= 100
Plants
114
41
48
59
39
13
11
26
32
41
N
559
198
224
281
192
66
50
109
150
208
PctND
%
1.25
1.01
0.45
0
1.04
59.09
48.00
54.13
45.33
32.69
Mean
Ag/L
14.15
14.7
15.19
14.88
9.92
0.97
2.17
4.92
2.86
7.71
Median
jUg/L
12.25
12.73
12.89
13.28
8.19
0.00
0.60
0.00
0.44
2.19
STD
Ag/L
9.91
11.03
12.54
9.26
7.99
2.40
2.64
8.96
5.99
11.11
Min
//g/L
0
0
0
0.28
0
0.00
0.00
0.00
0.00
0.00
Max
Ag/L
69.25
69.75
72.5
55.25
65.7
17.48
9.03
42.00
33.95
60.25
plO
//g/L
4
3.7
2.08
4.63
2.95
0.00
0.00
0.00
0.00
0.00
p90
jUg/L
24.5
29.25
28.6
28.5
20
2.55
5.95
20.00
7.69
24.03
Source: SW: Surface Water; GW: Groundwater
MRL: Minimum reporting limit
Plants : Number of plants sampled
N: Number of samples
PctND: Percent samples non-detect (detection limits not provided)
Mean/Median: Arithmetic mean/median of all samples
STD: Standard deviation
Min: Minimum Value
Max: Maximum Value
plO: lOthpercentile
p90: 90th percentile
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
A.2.3 Total Organic Carbon (TOC) Concentration
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 two-year DBP-monitoring study of more than 100 treatment plants of the American
Water System from 1989 to 1991, and found that only 11 and 15 percent of the variation in the
HAAS total (monochloroacetic acid, DCA, trichloroacetic acid, monobromoacetic acid and
dibromoacetic acid) could be explained by TOC for the distributed-water samples and plant
effluent, respectively.
A.2.3.1 TOC Concentration Data from the ICR Database
Table II-4 presents data from the ICR database for the concentration of DCA as a
function of influent TOC concentration. As indicated, influent TOC levels were higher, on
average, in surface water than in groundwater. For example, 83 of the 123 plants that use
groundwater as the source (67%) had TOC concentrations <1 ppb, while only 12 of the 299
plants with surface water as the source (4%) had such low TOC levels in the influent water.
Higher TOC levels in influent surface water are reasonable, because surface water
sources can contain decaying vegetation or animal matter; groundwaters do not have a
comparable risk for contamination by decaying organic matter and thus, generally have lower
TOC levels.
The concentrations of DCA in treated surface water were generally the same or higher
than DCA concentrations in treated groundwater at corresponding equivalent TOC
concentrations, up to 3 ppb. Differences in the nature of the TOC material and the levels of
disinfectant used by surface water systems compared to groundwater systems could have been
factors leading to the higher DCA levels in the surface water systems. A regression analysis of
the ICR data indicates that overall there was a significant correlation (at a = 0.05) between
influent TOC concentrations and the mean concentrations of DCA in treated surface water and
groundwater, with DCA concentrations increasing with increasing TOC concentrations (Table
II-4)
A.2.4 Seasonal Shifts
Williams et al. (1998) examined the concentrations of DBFs in winter and summer in raw
intake water, finished water, and water within the distribution-system main line at water-
treatment plants that used different disinfectant-treatment combinations. In the first survey, the
authors sampled raw water intake, finished water (water after treatment prior to distribution), and
waters near the midpoint of the distribution system for 52 Canadian water-treatment facilities.
The raw water sources included 28 rivers, eight lakes, three wells, a dam, and two sources that
12
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
Table II-4
DCA by Influent Total Organic Carbon (TOC) Concentration
Quarterly Distribution System Average
(ICR, U.S. EPA, 2000a)
Source
SW
GW
Influent TOC
Concentration
(ppb)
<1
l-<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
483
306
323
405
53
37
8
80
PctND
9.68
0.37
0.21
0.33
0.93
63.21
1.89
2.70
0.00
0.00
Mean
Ag/L
5.92
11.07
12.66
15.52
18.38
1.01
6.67
7.86
20.71
19.00
Median
yUg/L
4.04
9.74
11.13
14.19
15.50
0.00
6.23
3.35
22.18
14.60
STD
yUg/L
5.26
6.56
8.91
9.93
13.31
2.30
3.78
10.84
3.26
12.47
Min
Ag/L
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
14.3
1.28
Max
yUg/L
18.00
36.25
56.73
68.13
72.50
17.48
16.03
42.00
23.68
60.25
plO
Ag/L
0.90
3.58
3.30
4.58
5.30
0.00
2.13
0.95
14.30
5.60
p90
yUg/L
14.25
20.75
22.95
26.85
34.00
3.50
12.58
22.00
23.68
37.03
Source: SW: Surface Water; GW: Groundwater
Plants: Number of plants sampled
N: Number of samples
PctND: Percent samples non-detect (detection limits not provided)
Mean: Arithmetic mean of all samples
Median: Median value of all samples
STD: Standard deviation
Min: Minimum Value
Max: Maximum Value
plO: lOthpercentile
p90: 90th percentile
were a mixture of the aforementioned sources. Pre- and/or post-chlorination was used at 35
facilities and pre-chlorination coupled with post-chloramination was used at ten facilities. Seven
facilities used ozone coupled with chlorine or chloramine. DBFs in the raw water samples were
either not present or detected at very low levels.
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
In general, the mean concentrations of DCA were higher in the summer than in the
winter. DCA concentrations in the finished and distribution system waters were generally within
10 to 15 percent of each other. However, for facilities that used ozone coupled with chlorine or
chloramine, the concentrations of DCA in distribution-system waters were approximately 33
percent lower than in the finished waters.
To better understand the seasonal and distribution-system effects (effects of reaction
time) on DBFs, Williams et al. (1998) conducted a second survey. Because there is increased
reaction time with increasing distance from the treatment plant, the effects of reaction time may
be evaluated by sampling water throughout the distribution system. Water samples were taken at
five locations within the supply system once a month for one year at three facilities that used
different treatment combinations (pre- and post-chlorination, pre-chlorination coupled with post-
chloramination, and ozone coupled with chlorine). DBF concentrations were measured in raw
water entering the treatment system, finished water prior to entering the distribution system, and
water at three points within the distribution system (Dl - closest to the plant, D3 - the end of the
system, and D2 - a point midway between Dl and D3).
As in the previous survey, DBFs in the raw water samples were either not present or
detected at very low levels. Seasonal trends for DCA were inconsistent for all of the treatment
processes. For pre- and post-chlorination, the concentrations of DCA were similar for finished
water and water at sampling points Dl and D2, but were much lower at D3. Some investigators
attributed this decrease to bacterial degradation, but it has not been determined whether the
degradation mechanism is biological or chemical. For pre-chlorination coupled with post-
chloramination, the levels remained fairly constant throughout the year at all locations except
D2, which had a distinctly lower DCA concentration in the summer. The D2 sampling point for
this treatment facility was situated in a line that connected the main distribution-system pipe to a
system reservoir. The DCA concentrations at this sampling point would depend on whether the
direction of flow was to or from the reservoir. DCA concentrations in water from the ozone-
chlorine facility showed no clear trends with location or season, except that the finished water at
the treatment facility in the summer had a much higher DCA concentration than any other
location.
In the survey conducted by Krasner et al. (1989), there were no clear trends of the
concentrations of ions or chlorinated acetic acids with season in the composite analysis. DCA
concentrations were lowest in winter when the concentration of chloride ion was also at its
lowest level. Some observed shifts at utilities were seen as the result of drought conditions and
saltwater intrusion. In the summer, the concentrations of DCA were at their highest and the
concentration of bromide the lowest. In the winter, when DCA was at its lowest concentration,
the bromide concentrations were at their highest. This study suggests that as the concentration of
bromide ion increases, the concentration of DCA appears to decrease and is consistent with the
concept that the increase in bromide levels increases the concentrations of the bromine-
containing DBFs relative to those only containing chlorine.
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
A.2.4.1 Seasonal Shift Data from the ICR Database
The seasonal mean concentrations of DCA are presented in Table II-5. Based on the
mean concentrations for the different seasons, the levels of DCA formed in treated surface water
are statistically higher in the spring and summer than they are in the fall and winter. This is in
general agreement with the findings of Krasner et al. (1989), who found that DCA levels were
highest in the summer. No seasonal differences in mean DCA concentrations in treated
groundwater could be discerned.
Table II-5
DCA by Sample Quarter
Quarterly Distribution System Average Means
Sample Quarter
Summer '97
Fall '97
Winter '98
Spring '98
Summer '98
Fall '98
Dichloroacetic Acid
Surface Water (ug/L)
N
240
250
240
262
251
230
Mean
(Hg/L)
14.99
12.23
12.59
16.35
15.05
12.09
STD
(Hg/L)
11.99
8.57
9.89
10.95
11.05
7.54
Groundwater (ug/L)
N
93
88
103
105
101
93
Mean
(Hg/L)
5.21
3.33
4.66
5.09
5.08
4.68
STD
(Hg/L)
10.13
5.73
8.85
9.01
9.03
8.62
Summer '97/'98: July, August, and September
Fall '97/'98: October, November, and December
Winter '98: January, February, and March
Spring '98: April, May, and June
B. Exposure to Sources Other Than Drinking Water
DCA has been used in research and development laboratories. Between 1981 and 1983,
the National Institute of Occupational Safety and Health (NIOSH) conducted a survey of a
sample of 4,490 businesses employing nearly 1,800,000 workers (NIOSH, 1990). Data were
collected on the use of DCA and trade name products known to contain DCA. During the period
from 1981 to 1983, 1592 workers were potentially exposed to DCA in 39 plants. Exposure
levels were not reported in this survey, and more recent information on numbers of workers
exposed is not available.
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
B.I Dietary Exposure
Chlorine is used in food production and processing, such as in disinfection of chicken in
poultry plants; processing seafoods, poultry and red meats; sanitizing equipment and containers;
cooling heat-sterilized foods; oxidizing and bleaching in the flour industry. The use of chlorine
as a sanitizer is permitted by the FDA under 21 CFR 178.1010 at a maximum concentration of
100 ppm. Sodium hypochlorite is approved as a secondary direct additive for the washing of
fruits and vegetables (21 CFR 173.315) and both sodium and calcium hypochlorite can be used
for bleaching of modified food starch (21 CFR 172.892) at a concentrations of 0.0082 Ibs
chlorine equivalent/pound of starch. Therefore, DCA is likely be to be found as a disinfection
byproduct in a variety of food products.
The concentrations of DCA in several vegetables, fruits, grain, and beer samples from
Switzerland (Reimann et al., 1996a) are presented in Table II-6. DCA concentrations were
typically lower in fruits and vegetables and higher in grains and grain-based products.
Table II - 6
DCA in Foods
FOOD GROUP
Vegetables
Fruits
Grain
Flour/Bread
Beer (//g/L)
DCA
Range of Concentrations
Ag/kg
< 0.9 (tomato) - 3.5 (kohlrabi leaf)
<0.9
< 0.6 (wheat)- 11.1 (barley)
0.8 (brown bread) - 19.8 (flour)
1.5 (Mexico)- 15.2 (Peru)
4.5 (USA)
Recent investigations by Raymer and associates (2004) indicate that foods can take up
DCA from water during the cooking process. For example, when water is spiked with DCA at
150 ppb, DCA is taken up at the following percentages in these foods: carrots, 64%; green
beans, 48%; pinto beans, 85%; lettuce, 3.9% (after soaking for 5 minutes), and chicken, 64%.
The spiked concentration was higher than the levels generally found in drinking water (see Table
II-1). Uptake was determined by comparing the DCA extracted from the foods cooked in spiked
water to the levels in foods cooked in reagent water. DCA was stable to boiling of spiked water
over a 60-minute period.
Pasta is a common food that is cooked in water potentially containing DCA. Raymer and
coworkers (2004) found that pasta could take up DCA both from spiked cooking water and the
16
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
spiked water used to rinse the pasta following cooking. For example, 11% of the available DC A
in cooking water was taken up during cooking of the pasta. When rinsed with spike water after
cooking, the pasta took up an additional 7% of the DC A in the rinse water. When the pasta was
cooked in unspiked water but rinsed with spiked water it took up 7.7% of DCA from the rinse
water. These data indicate that cooking foods in water containing DCA at the 150 ppb level can
contribute to the daily dose of DCA from combined food and water sources.
B.2 Air Exposure
DCA is not volatile, and thus, is not likely to be present in indoor air as a result of water
use within a home. However, it can be present in outdoor air since it can be formed when
organic compounds are burned in the presence of a chlorine source. DCA has been detected in
rainwater at concentrations ranging from 0.05 to 4 |ig/L (Reimann et al., 1996b).
B.3 Dermal Exposure
Some exposure to disinfection byproducts can occur from swimming pools and hot tubs
(where chlorine is added to the water as a disinfectant) as well as during bathing and showering.
Clemens and Scholer (1992) reported DCA concentrations in 15 indoor and 3 outdoor swimming
pools in Germany. DCA concentrations were significantly higher in outdoor pools (mean, 119.9
|ig/L; range, 83.5 - 181 |ig/L) than in indoor pools (mean, 5.6 |ig/L; range, 0.2 - 10.6 |ig/L).
Kim and Weisel (1998) measured DCA concentrations in three indoor pools during 1995
and 1996. The mean concentration of DCA was 419 |ig/L, which is more than an order of
magnitude higher than that reported for indoor pools by Clemens and Scholer (1992), and about
four times higher than that reported for outdoor pools. Potential reasons for the different values
may be differences in chlorine amounts added to disinfect pools, or differences in sample
collection time relative to water chlorination. Kim and Weisel (1998) estimated an expected
dermal dose of 6 jig DCA following a 30-minute exposure to pool waters containing 600 |ig/L
DCA. Using a DCA drinking-water concentration of 30 |ig/L, Kim and Weisel (1998) estimated
that dermal absorption while bathing and showering are minor routes of exposure compared with
ingestion of drinking water. However, Kim and Weisel (1998) estimated that someone who
swims regularly (three times/week) could receive an additional 10% to 20% of the weekly dose
from ingestion of and dermal exposure to swimming-pool water. Experimental data on DCA
dermal absorption are unavailable, however.
C. Overall Exposure
DCA concentrations in drinking water average 13.93 |ig/L in surface water, and 4.7 |ig/L
in groundwater. Concentrations of DCA in food, based on limited data from Reimann et al.
(1996a), may contribute to the overall exposure. Studies by Raymer and associates (2004)
indicate that DCA can be taken up from cooking water when foods are cooked in disinfected
water containing DCA as a DBF. Use of chlorine as a sanitizer in food processing may also
17
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Addendum to the IRIS Toxicological Review for Dichloroacetic Acid
contribute DCA to the food supply. However there has been no dietary study that examined the
presence of DCA in prepared foods that make up a typical diet. Although DCA has been
detected in rainwater at concentrations ranging from 0.05 to 4 |ig/L (Reimann et al., 1996b),
there are no data on air concentrations of DCA.
DCA concentrations in swimming-pool water averaged 419 |ig/L in one study of three
pools (Kim and Weisel, 1998). Nevertheless, Kim and Weisel (1998) estimated that dermal
exposure to DCA would be minor compared to other exposure pathways. Inhalation dose during
showering was considered to be insignificant since there is minimal volatilization of DCA.
C.I Body Burden
The rapid metabolism and excretion of DCA and its metabolites suggests little potential
for bioaccumulation. Limited data are available on the background levels of DCA in the blood
or urine. Kim and Weisel (1998) reported a background DCA-urinary-excretion rate of
approximately 0.3 ng/min. Fisher et al. (1998) did not observe any DCA in the plasma of
unexposed human subjects.
D. Summary
The ICR database (U.S. EPA, 2000a) contains extensive information on concentrations of
DCA in drinking-water systems, and how those concentrations vary with input-water
characteristics and treatment methods. As shown in the database and published evaluations, the
mean concentrations of DCA were in general lower in disinfected groundwater than in surface
water. The mean concentrations of DCA were 4.7 and 13.93 |ig/L in drinking water derived
from groundwater and surface water sources, respectively.
Variability in DCA concentrations among the ICR samples made it difficult to draw firm
conclusions regarding the effect of disinfectant, influent bromide concentration, TOC and
seasonality of DCA formation. Some studies indicate that the levels of bromine-containing
haloacetic acids will increase and the concentrations of the chlorinated species will decrease as
the bromide concentration of the raw water increases. Mean concentrations of DCA in surface
waters seem to be higher in the summer than during the cooler seasons of the year.
In addition to DCA concentrations in drinking water, there are some limited data on DCA
concentrations in rain, food, and swimming pools. Very limited data are available on the
concentrations of DCA in foods, but the concentrations in the foods measured are comparable to
those in water and may contribute to the overall dose. Although there are data on the
concentration of DCA in rainwater, there are no data on air concentrations of DCA. Data
available on DCA concentrations in swimming pool water and estimated dermal absorption rates
and frequency of exposure indicate that dermal absorption of DCA through swimming and
showering/bathing is not expected to contribute a significant amount toward the total dose.
However, as stated previously, these observations are based on an incomplete set of data.
18
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
Chapter III. Quantification of Toxicological Effects
The quantification of toxicological effects of a chemical consists of separate assessments
of noncarcinogenic and carcinogenic health effects. Unless otherwise specified, chemicals which
do not produce carcinogenic effects are believed to have a threshold dose below which no
adverse, noncarcinogenic health effects occur, while carcinogens are assumed to act without a
threshold unless there are mode of action data that indicate otherwise.
A. Introduction to Methods
A. 1 Quantification of Noncarcinogenic Effects
In quantification of noncarcinogenic effects, a Reference Dose (RfD) is calculated. 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 adverse effect level (NOAEL), lowest observed adverse effect level
(LOAEL), or a NOAEL surrogate such as a benchmark dose identified from a subchronic or
chronic study, and divided by a composite uncertainty factor(s). The RfD is calculated as
follows:
RfD= NOAEL or LOAEL
UF
where:
NOAEL = No-observed-adverse-effect level expressed as mg/kg/day from a high-
quality toxicological study of an appropriate duration.
LOAEL = Lowest-observed-adverse-effect level expressed as mg/kg/day 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
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),
19
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
uncertainty is broken down into its components, and each dimension of uncertainty is given a
quantitative rating. The total uncertainty factor is the product of the component uncertainties.
The individual components of the uncertainty factor are as follows:
UFH A 1, 3, or 10-fold factor used when extrapolating from valid data in studies
using long-term exposure to average healthy humans. The intermediate factor
of 3 is approximately !/2 Iog10 unit, i.e., the square root of 10. 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 to adjust for the absence of data on
toxicological endpoints considered critical for assessing human risk.
Frequently, it is applied if data for endpoints that need to be experimentally
addressed in specialized studies (e.g. reproductive and developmental toxicity)
are lacking. The 3-fold factor is often used when there is a single data gap
exclusive of chronic data.
On occasion, EPA also uses a modifying factor in the determination of the RfD. A
modifying factor (MF) is an additional uncertainty factor that is greater than zero and less than or
equal to 10. There are a limited number of situations where a modifying factor has been applied.
One situation in which a MF might be appropriate is where structure function relationships
indicate that a hazard may exist that has not been studied using the compound of interest. The
magnitude of the MF depends upon the professional assessment of scientific 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.
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
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 product of UF and MF that exceeds 3000,
then the database does not support development of an RfD. The quantification of toxicological
effects of a chemical consists of separate assessments of noncarcinogenic and carcinogenic
health effects. Unless otherwise specified, chemicals that do not produce carcinogenic effects
are believed to have a threshold dose below which no adverse, noncarcinogenic health effects
occur.
A. 1.1. Drinking Water Equivalent Level
The drinking water equivalent level (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:
DWEL = (Rfm x BW
WI
where:
BW = 70 kg adult body weight
WI = Drinking water intake (2 L/day)
A. 1.2. 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 of a duration
consistent with the HA duration and identify a sensitive noncarcinogenic endpoint of toxicity.
The HAs are derived as follows:
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
HA = NOAEL or LOAEL x BW
UF x WI
where:
NOAEL or LOAEL = No- or lowest-observed-adverse-effect level in mg/kg/day
BW = Assumed body weight of a child (10 kg) or an adult (70 kg)
UF = Uncertainty factor, in accordance with EPA or OW
guidelines
WI = Assumed daily water consumption 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 the 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
RSC = Relative source contribution. The fraction of the total exposure
allocated to drinking water.
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
A.2 Quantification of Carcinogenic Effects
The evaluation of carcinogenic potential involves two levels of assessment. The first is a
qualitative judgment based on the weight of evidence of the likelihood that the agent is a human
carcinogen and the conditions under which the carcinogenic effects may be expressed. The
second level provides a quantitative estimate of cancer potency (slope factor) when the data
available are suitable for modeling of the dose-response. The cancer 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 DCA IRIS assessment (US EPA, 2003) and this addendum also include the drinking water
concentration equivalent to cancer risks of one-in-ten-thousand (10"4), one-in-one-hundred-
thousand (10"5), and one-in-one-million (10"6).
Cancer assessments conducted before 1996 used a five-category, alphanumeric system
for classifying carcinogens (U.S. EPA, 1986). After 1999, assessments were conducted based on
the revised versions of the Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1996, 1999).
Although the 1999 version of the Guidelines for Carcinogen Risk Assessment had not been
finalized by the agency, use of the 1986 guidelines ceased in 2001 with the publication of a
directive from the administrator (U.S. EPA, 2001) specifying that the 1999 draft guidelines were
to be used on an interim basis for any new cancer assessment by the Agency. The U.S. EPA
assessment of the carcinogenicity of DCA for the Integrated Risk Information System (U.S.
EPA, 2003) provides both a qualitative weight of evidence descriptor for DCA and a quantitative
estimate of the cancer potency. (See Section 5.3 [page 89] of the Toxicological Review.)
B. Noncarcinogenic Effects
B. 1 One-day Health Advisory
No studies were identified that were appropriate for the derivation of a One-day HA
value (see Table III-l). Several studies were evaluated for their potential to serve as the basis
of a One-day HA but were found to be unsuitable.
There are a number of studies of human subjects exposed to DCA through its use in
treatment of congenital lactic acidosis and other disease conditions (Section 4.1, [page 17] of the
IRIS Toxicological Review). Some of the subjects that took part in these studies did not suffer
from disease and were classified as healthy. The numbers of healthy subjects tested and their
durations of exposure varied. In one case (Stacpoole et al., 1998a), 14 subjects were given oral
doses of 2.5, 25, or 250 mg/kg/day for 5 consecutive days for evaluation of in vivo metabolism
of DCA in humans. The authors reported that about 50% of the healthy subjects receiving 25
mg/kg/day and above experienced neurological effects (sedation, diminished anxiety) initiating
in the 60-minutes after dosing and lasting for several hours. Accordingly the 25-mg/kg/day dose
can be considered a LOAEL in humans.
23
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
Neurobehavioral deficits were observed after a single gavage dose of 0, 300, 1000, or
20,000 mg/kg/day DC A in Long-Evans rats (Moser et al., 1999). Hind limb grip strength was
decreased in all three groups at 4 hours post-dosing but there was no dose-response relationship.
At 24 hours after exposure, only the 300 mg/kg group exhibited decreased hind limb grip
strength significantly different from controls. These effects were reversible, with recovery
occurring 7-14 days after dosing. This study was judged to be inappropriate for the derivation of
a One-Day HA due to a lack of a clear dose-response and the absence of data on other toxicity
endpoints.
In a single gavage dose study by Linder et al. (1997; page 38 of the Toxicological
Review), there was some evidence of mild effects on spermiation in DCA-treated animals 2, 14,
and 28 days after exposure to a dose of 3000 mg/kg. Testicular histopathology was indicative of
retention of mature spermatids in the lumen of the testicular tubules. However, use of a
reproductive-toxicity endpoint seen in sexually mature rats is not appropriate for the
development of a One-Day HA for a 10-kg child.
Austin et al. (1996; page 50 of the Toxicological Review) showed changes in lipid
peroxidation and the formation of oxidatively modified DNA bases in animals treated with a
single gavage dose of 300 mg/kg DCA. A single-dose acute study by Davis (1990; page 22 of
the Toxicological Review) only measured metabolic changes. In both of these cases it is not
clear whether the observed effects after the exposure to a single dose persist leading to
permanent physiological damage; thus, neither endpoint is considered appropriate for use in
derivation of a One-Day HA.
The effect level in the animal studies occurs at doses higher than those reported in
healthy humans. However, the human data come from third-party deliberate dosing studies. The
U.S. EPA has not yet finalized its policies with regard to the use of deliberate human-dosing
studies in risk assessment. Accordingly the Stacpoole et al. (1998a) data cannot presently be
used to derive a One-day HA value. Each of the animal studies are also inappropriate for the
determination of a One-day HA as described above. Accordingly, the Ten-Day HA of 3 mg/L,
developed below, is recommended as a conservative One-Day HA.
B.2 Ten-day Health Advisory
Several studies of suitable duration for calculation of the Ten-day HA have been
conducted and are listed in Table III-l. A number of studies were judged to be unsuitable
because they were designed to measure metabolic changes which may or may not have been
adverse effects (e.g., Davis, 1986; Ribes et al., 1979; Evans and Stacpoole, 1982; Section 4.2.1
of the Toxicological Review). The study by Stacpoole et al. (1978) with a LOAEL of 43
mg/kg/day was a clinical study with only one dose level, which was a LOAEL.
24
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
Table III-l. Summary of NOAELs and LOAELs Considered for Development of One- and
Ten-Day Drinking Water Health Advisories (HA) for DCA1
Study
Type
Acute
Acute repro.
Acute
Acute
Clinical
studies
Subacute
Subacute
Subacute
Subacute
Subacute
Subacute
repro.
Subacute
Subacute
Develop.
Develop.
Species
Rat
Rat
Rat
Mouse
Human
Mouse
Rat
Dog
Rat
Mouse
Rat
Rat
Mouse
Rat
Rat
Duration/
No. Dose
Groups2
1 -day/3
1 -day/3
1 -day/4
1 -day/3
6-7 days/1
3 -weeks/4
2-weeks/5
7-days/2
7-days/2
2-weeks/4
2-weeks/5
2-weeks/4
2-weeks/4
GD 6-15/4
GD 9-11, or 12-
15, or GD 9,
10, or 12 /I
dose per group
Critical Effects
Metabolic changes
Mild spermiation
Neurobehavioral deficits
Changes in lipid
peroxidation, DNA
adducts
Sedation, mild, occas.
periph. neuropathy, I
serum glucose/ lactate
T absolute/ relative liver
wt
Metabolic changes
I serum metabolites
I serum lactate
T glycogen, T liver wt.,
focal necrosis
Impaired sperm
formation
I body wt.
T relative liver wt.
I fetal wt, incr.
resorptions
Fetal malformations,
including cardiopathy
NOAEL
mg/kg/d
120
1500
(equivocal)
300
100
ND
25
150
ND
ND
75
18
ND
ND
14
ND
LOAEL
mg/kg/d
ND3
3000
—
300
43
125
ND
150
100
250
54
166
90
140
1990 to
3500
Reference
Davis, 1990
Linderatal.,
1997
Moseretal.,
1999
Austin et al.,
1996
Stacpoole et
al., 1978
Parrishetal.,
1996
Davis, 1986
Ribes et al.,
1979
Evans and
Stacpoole,
1982
Sanchez and
Bull, 1990
Linderetal.,
1997
DeAngelo et
al., 1989
DeAngelo et
al., 1989
Smith etal.,
1992
Epstein etal.,
1992
Abbreviations: GD = gestational day; T, increased; I, decreased; ND = not determined
No. of dose groups includes both treated and control groups.
25
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
Several studies reported adverse reproductive or developmental effects in rodents after short term
exposures (Linder et al., 1997; Smith et al., 1992; Epstein et al., 1992; Section 4.3 of the
Toxicological Review). However, the critical effect in the reproductive study with the lowest
LOAEL, increased sperm retention and residual bodies (Linder et al., 1997), is relevant only to
sexually mature animals and not to children, and so is not appropriate for the derivation of a Ten-
Day HA for a 10-kg child. Similarly, the developmental endpoints for the fetus observed by
Smith et al. (1992) and Epstein et al. (1992) are not appropriate for a Ten-Day HA.
A 21-day drinking-water study by Parrish et al. (1996), which demonstrated statistically
significant increases in absolute and relative liver weights in the absence of body-weight changes,
is judged to be suitable for the development of a Ten-Day HA. The IRIS Toxicological Review
does not provide a detailed description of the systemic toxicity seen in this study. Accordingly,
the details are presented below.
Male B6C3F1 mice were exposed to DC A in drinking water at concentrations of 0, 100,
500, or 2000 mg/L for either 3 or 10 weeks (Parrish et al., 1996). The study authors did not
estimate the doses, however, based on a default water intake of 0.25 L/kg/day for male B6C3F1
mice (U.S. EPA, 1988), the corresponding doses were 0, 25, 125, and 500 mg/kg/day. No
significant changes were observed in body weight at the end of 3 or 10 weeks. Both absolute and
relative liver weights were statistically increased at the mid and high doses in a dose-related
manner at both time points. DCA produced a small but statistically significant increase in acyl-
CoA oxidase activity, a biomarker for peroxisome proliferation, only at the high dose at 3 weeks.
Extension of DCA treatment to 10 weeks produced small, but statistically significant, increases in
acyl-CoA activity at the two lowest doses of DCA but no response at the high dose. Based on the
increased liver weight (a characteristic response from DCA exposure), 125 mg/kg/day is LOAEL,
and the NOAEL is 25 mg/kg/day.
The Parrish et al. (1996) study was designed to evaluate lipid peroxidation and oxidative
DNA damage in liver tissue, accordingly other organs and end points were not evaluated.
However, the liver is known to be a primary target of DCA-induced toxicity and thus the data are
suitable for derivation of the Ten-day HA. The LOAEL and NOAEL for the Parrish et al. (1996)
study were 125 and 25 mg/kg/day, respectively. Other studies of similar exposure duration
showed similar liver effects but at higher doses (DeAngelo et al., 1989; Sanchez and Bull, 1990;
discussed in Section 4.2.1 of the Toxicological Review), with corresponding higher LOAELs and
NOAELs. The LOAELs for increases in liver weights in 10-week and 3-month studies of rats by
Toth et al. (1992) and Mather et al. (1990) were 31 and 32.5 mg/kg/day, respectively.
Sanchez and Bull (1990) exposed B6C3F1 mice to estimated DCA doses of 0, 75, 250, or
500 mg/kg/day in drinking water based on the default water intake and body weight for male
B6C3F1 mice (U.S. EPA, 1988). A dose-related increase in liver weight was observed beginning
at 75 mg/kg/day, but did not reach statistical significance until 250 mg/kg/day. Increases in
hepatic cell size were significant at 250 and 500 mg/kg/day. Focal necrosis was also observed at
26
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
250 and 500 mg/kg/day, indicating a progression of hepatic effects with increasing dose. The
NOAEL in the Sanchez and Bull (1990) study was 75 mg/kg/day, based on increased liver
weight.
The LOAEL from Parrish et al. (1996) is lower than the LOAELs for developmental
effects seen in the Smith et al. (1992) and Epstein et al. (1992) studies. Therefore, the Ten-Day
HA derived from the Parrish et al. (1996) study should also protect against potential
developmental effects of DC A.
T ™ UA (25 mg/kg/day) (10 kg) n , , , n.
Ten-Day HA = (100) (1 L/day) m§ (rounded to 3 m§/L)
where:
25 mg/kg/day = NOAEL, based on increased absolute and relative liver weights in the
absence of body weight changes in rats dosed in drinking water for 21 days
10 kg = assumed body weight of a child
100 = composite default uncertainty factor, chosen to account for extrapolation from
animals and for protection of sensitive subpopulations
1 L/day = assumed daily water consumption by a 10-kg child
This HA for a 10-kg child is adequate to protect adults from the spermatotoxic effects
observed in the Linder et al. (1997) study, which had a NOAEL of 18 mg/kg/day.
B. 3 Longer-term Health Advi sory
Studies of suitable duration for derivation of the Longer-term HA are summarized in
Table III-2. The U.S. EPA (1994) Longer-term HA was derived from a subchronic study
conducted by Cicmanec et al. (1991; page 28 of the Toxicological Review). Although several
new studies are available, they do not replace this principal study. Briefly, DCA was
administered orally via gelatin capsules to four month-old male and female dogs daily for 90
days. At necropsy, the following histopathology was observed in all dosed groups: (1)
degeneration of testicular germinal epithelium and syncytial giant-cell formation in the testes; (2)
hepatic vacuolation and chronic hepatitis; and (3) vacuolization of myelinated white tracts of the
cerebrum, cerebellum and/or spinal cord. The study identified a LOAEL of 12.5 mg/kg/day,
based on testicular, neurological and hepatic toxicity; no NOAEL could be determined.
27
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
Table III-2. Summary of NOAELs and LOAELs Considered for Development of Longer-
Term Drinking Water Health Advisory (HA), RfD, and DWEL for DCA 1
Study Type
Species
Duration/
No. Dose
Groups 2
Critical Effects
NOAEL
mg/kg/d
LOAEL
mg/kg/d
Reference
Subchronic Studies
Subchronic
toxicity
Subchronic
toxicity
Subchronic
toxicity
Subchronic
toxicity
Subchronic
toxicity
Subchronic
toxicity
Develop.
toxicity
Subchronic
toxicity
Repro.
toxicity
Clinical
studies
Mouse
Rat
Mouse
Dog
Rat
Rat
Rat
Dog
Rat
Human
10 weeks/4
3 months/2
8 weeks/ not
given
13 weeks/4
3 months/2
90 days/4
GD 6-15/4
3 months/4
10 weeks/4
5-16
weeks/ 1
T Absolute/relative liver
wt
Altered serum
metabolites, peripheral
neuropathy, hepato-
megaly, testic. degenerat.
T Glycogen
I Serum metabolites,
prostate, atrophy,
testicular changes, brain
vacuolation
I Serum lactate and
glucose, T liver wt,
histological brain lesions;
at higher doses, testic.
degeneration
I Serum protein, T liver
wt, T glycogen
deposition, liver histopath
T Liver wt. in pregnant
females
T Liver wt,
inflammation, brain
vacuolation,
testicular degeneration
T Liver wt, impaired
sperm formation
I Cholesterol, peripheral
nervous system effects
25
ND3
16
(estimate)
ND
ND
36.5
14
ND
ND
ND
125
323
80
(estimate)
50
125
355
140
12.5
31
50
Parrish et
al., 1996
Yountetal.,
1982
Kato-
Weinstein et
al., 1998
Katz et al.,
1981
Katz et al.,
1981
Mather et
al., 1990
Smith et al.,
1992
Cicmanec et
al., 1991
Toth et al.,
1992
Moore et
al., 1979
28
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
Study Type
Clinical
studies
Neurobehav.
toxicity
Neurobehav.
toxicity
Species
Human
Adult Rat
Weanling
Rat
Duration/
No. Dose
Groups 2
<5 years/1
12 weeks/4
13 weeks/4
Critical Effects
Sedation, occasional mild
peripheral neuropathy
Neurotox., including gait
abnormalities; at higher
doses, righting reflex
deficits, I motor activity,
I limb grip strength,
tremors
Neurotox. including gait
abnormalities; at higher
doses, I limb grip
strength, tremors,
hypotonia, righting reflex
deficits, inhib. pupil
reflex
NOAEL
mg/kg/d
ND
ND
ND
LOAEL
mg/kg/d
43
16
16
Reference
Stacpoole et
al., 1998b
Moser et al.,
1999
Moser et al.,
1999
Chronic Studies
Chronic tox./
carcinogen.
Chronic tox./
carcinogen.
Chronic tox./
carcinogen.
Chronic tox./
carcinogen.
Chronic tox./
carcinogen.
Chronic tox./
carcinogen.
Mouse
Mouse
Mouse
Rat
Rat
Mouse
1 year/3
60 weeks/5
104 weeks/2
104
weeks/4
103 weeks/2
90-100
weeks/75
T Relative liver wt,
hepatocytomegaly, liver
glycogen deposition, liver
focal necrosis
T Relative liver wt.
T Absolute and relative
liver wt, liver
enlargement, chronic
inflammation, necrosis
T Absolute testes wts.
without I body wt.
I Body wt., I absolute
testes wt., T liver
vacuolation
T Liver weight
137
7.6
ND
3.6
ND
8
270
77
88
40.2
139
84
Bull et al.,
1990
DeAngelo et
al., 1991
Daniel et
al., 1992
DeAngelo et
al., 1996
DeAngelo et
al., 1996
DeAngelo et
al., 1999
1 Abbreviations: GD - gestational day, T, increased; I, decreased
2 No. of dose groups includes both treated and control groups
3 ND: could not be determined
4 High dose discontinued at 60 weeks.
5 Two control groups
29
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
A number of other studies have also demonstrated that the testes, nervous system, and
liver are target organs of DCA-induced toxicity (e.g., Katz et al., 1981; Bull et al., 1990; Bhat et
al., 1991 [Section 4.2.1 of the Toxicological Review]; DeAngelo et al., 1991, 1999; Daniel et al.,
1992 [Section 4.2.2 of the Toxicological Review]; Toth et al., 1992 [Section 4.3 of the
Toxicological Review]; Parrish et al., 1996; Moser et al., 1999; Linder et al., 1997; and Spruijt et
al., 2001 [Section 4.1 of the Toxicological Review]). The Cicmanec et al. (1991) study is the
most appropriate for derivation of a Longer-term health advisory because the LOAEL for this
experiment is the lowest (i.e., most sensitive) among all studies, and the available data do not
indicate the appropriate experimental animal species for extrapolation to humans. The LOAEL
from this study also serves as the basis for the derivation of the RfD in the next section (B4) of
this report.
The choice of the Cicmanec et al. (1991) study is supported by other studies with similar
quantitative data on dose-response. For example, Moser et al. (1999; page 34 of the
Toxicological Review) identified a LOAEL of 16 mg/kg/day, based on gait impairment in both
weanling and adult male rats exposed to DC A in drinking water for 12-13 weeks; no NOAEL was
identified. Other animal studies identified similar toxic endpoints but at higher LOAELs, ranging
from 31 mg/kg/day (Toth et al., 1992; page 39 of the Toxicological Review) to 323 mg/kg/day
(Yount et al., 1982; page 32 of the Toxicological Review). In these studies, the LOAEL was the
lowest dose tested and a NOAEL could not be determined. A benchmark dose for a 10% response
level (BMD10) of 3.2 mg/kg/day, and a 95% lower bound (BMDL10) of 2.1 mg/kg/day was
calculated in the DC A Toxicological Review (pg. 87) for the incidence of testicular degeneration
in the Cicmanec et al. (1991) study. However, there was low confidence in this value as there
was an 80% response at the low dose, meaning that there is little information on the shape of the
dose-response curve in the range of the benchmark response (BMR). Nonetheless, it is of interest
that the BMDL10 is within a factor of two of the LOAEL/10. The data on altered gait in the
Moser et al. (1999) study were not presented in a form amenable to BMD modeling (i.e., no
incidence data, and no information on variability for the continuous data). Accordingly, the
LOAEL was used as the point of departure for the Longer-term HA derivations.
The composite uncertainty factor is 1000 (10 x 10 x 3 x 3), based on a full default
uncertainty factor of 10 each for extrapolation from a LOAEL to a NOAEL and to account for
inter-individual variability, and 3-fold factors to account for interspecies extrapolation and
database deficiencies. This composite uncertainty factor is used in conjunction with the LOAEL
of 12.5 mg/kg/day identified by Cicmanec et al. (1991) for calculation of the Longer-term HA.
The Longer-term HA for a 10-kg child consuming 1 L/day of water is calculated as follows:
T T TTA t u-u\ (12.5 mg/kg/day) (10 kg) A10c n t JJ^AI n\
Longer-Term HA (child) = , ,7 -,, - •' °* = 0.125 mg/L (rounded to 0.1 mg/L)
^luuuj ^1 L/day)
where:
30
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
12.5 mg/kg/day = LOAEL, based on histopathology observed in the testes, brain, and
liver of dogs orally exposed to DCA for 90 days.
10 kg = assumed body weight of a child.
1000 = composite uncertainty factor, chosen to account for interspecies extrapolation
from the dog to the human, inter-individual variability in humans, extrapolation from a
LOAEL to a NOAEL from an animal study, and database insufficiencies.
1 L/day = assumed daily water consumption by a 10-kg child.
The longer-term HA for a 70-kg adult consuming 2 L/day of water is calculated as
follows:
T T UA t A u\ (12.5 mg/kg/day) (70 kg) n f , , ^ A „ n\
Longer-Term HA (adult) = nnnnvo T IA \ =0.44 mg/L (rounded to 0.4 mg/L)
(1000)(.Z L/day)
where:
12.5 mg/kg/day = LOAEL, based on histopathology observed in the testes, brain, and
liver of dogs orally exposed to DCA for 90 days.
70 kg = assumed body weight of an adult.
1000 = composite uncertainty factor, chosen to account for interspecies extrapolation
from the dog to the human, inter-individual variability in humans, extrapolation from a
LOAEL to a NOAEL from an animal study, and database insufficiencies.
2 L/day = assumed daily water consumption by a 70-kg adult.
The Cicmanec et al. (1991) LOAEL is also the point of departure for the derivation of the RfD in
the following section. The full rationale for each of the individual components of the composite
uncertainty factor is presented in full with the description of the RfD below. The duration
adjustment used in calculating the RfD is not necessary in the derivation of the Longer-term HA
B.4 Reference Dose, Drinking Water Equivalent Level, and Lifetime Health Advisory
The Reference Dose developed for the Toxicological Review (page 88) and the derivation
of the DWEL are presented in this section. The RfD was developed using the LOAEL of 12.5
mg/kg-day identified in the Cicmanec et al. (1991) study, with the application of a composite
uncertainty factor of 3000. The uncertainty factor includes a factor of 10 to account for potential
interhuman variability in susceptibility to DCA, a factor of 3 to account for extrapolation from
31
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
animal data to humans, a factor of 10 to account for the use of a LOAEL, a factor of 3 to account
for the use of a less-than-lifetime study in which frank effects were noted, and a factor of 3 to
account for deficiencies in the database. The RfD derivation is taken directly from the IRIS
Toxicological Review (US EPA, 2003).
Step 1: Determination of an RfD for Dichloroacetic Acid:
RfD = 12.5 mg/kg/dav = 0.0042 mg/kg/day (rounded to 0.004 mg/kg/day)
3000
where:
12.5 mg/kg/day = LOAEL, based on testicular toxicity, neuropathology, and liver toxicity
in dogs.
3000 = composite uncertainty factor for extrapolation from animals to humans, inter-
individual human variability, the use of a LOAEL instead of a NOAEL, the use of a less-
than-lifetime study, and database insufficiencies. The justification for the uncertainty
factors from the IRIS document is presented below:
A factor of 10 was applied for intrahuman variability because of the observation that the
most frequent human GSTZ variant (GSTZ Ic-lc) is one that has a low activity toward DCA and
is also impacted by DCA inhibition to a greater extent than the most active, but less frequent
human variant (GSTZ la-la). Accordingly, one might expect poor clearance of DCA from
human plasma via oxidative dechlorination when exposure is continuous.
A threefold factor was applied for interspecies variability. There are several reasons for
this choice and the resulting partial reduction of the UF from the default of 10. First, death
occurred at a dose of about 75 mg/kg-day DCA (90 day study) in 3/10 and 1/3 dogs after 51 and
74 days of dosing and 50 days of dosing, in the principle study by Cicmanec et al. (1991) and the
study by Katz et al. (1981) respectively. Conversely, Stacpoole et al. (1998b) reported on cases
of five children with lactic acidosis who received 25-60 mg/kg-day orally for two months to four
years without clinical signs of DCA toxicity (elevation of liver enzymes and neuropathy).
Although two of the children died during treatment, death was the result of infection and not from
the lactic acidosis or DCA treatment. Annual mortality in patients with congenital lactic acidosis,
even with treatment, is 20%.
Additional support for this conclusion is provided by the fact that metabolic effects of
DCA on serum lactate and glucose in dogs (Katz et al., 1981; Ribes et al., 1979) parallel those in
humans (Stacpoole et al., 1998a, b). Stacpoole et al. (1998b) reported that lactate concentrations
decreased by at least 20% within 24 hours after oral doses of 25 to 100 mg/kg in humans. In the
study by Katz et al. (1981), there was an approximate 40% reduction in serum lactate
concentrations of dogs (male and female) after 13 weeks of exposure to 50 mg/kg-day DCA.
32
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
Limited toxicokinetic data suggest that dogs metabolize DCA at a slower rate than humans and
rodents, supporting the concept of their increased sensitivity (Section 3.3; Lukas et al., 1980;
Curry et al., 1991; Lin et al., 1993; Larson and Bull, 1992; James et al., 1998). A single
intravenous dose of 100 mg/kg in two dogs led to peak plasma levels that were twice as high as
the same levels in rats (Lukas et al., 1980). Lastly, the structure of GST Zeta appears to be highly
conserved across species making it unlikely that the metabolic differences in humans will differ
from dogs by a full order of magnitude, also taking into consideration that the full UF=10 has
been applied for intrahuman variability. Under these circumstances an interspecies uncertainty
factor of 3 rather than the default 10-fold value is justified.
The factor of 10 for the use of a LOAEL is justified by the observed effects of DCA on the
nervous system in sensitive humans (those under treatment for lactic acidosis and other disorders)
at doses of 25-50 mg/kg-day. These doses are within the same order of magnitude as the LOAEL
in the Cicmanec et al. (1991) study and the LOAEL for neurotoxicity in adult and weanling F344
and LE rats in the Moser et al. (1999) study. There are no human data on testicular effects from
DCA.
Threefold factors were applied for both database inadequacies and the use of data from a
less-than-lifetime study. The database for DCA lacks a multi-generation study of reproductive
toxicity and a developmental neurotoxicity study, thus meriting an uncertainty factor of 3 for
database insufficiency. Otherwise the database is comprehensive with information from
subchronic and lifetime animal studies, studies in three animal species, and over 25 years of
experience with the use of DCA as a experimental pharmaceutical in the treatment of several
human disorders.
The richness of the data base does not abrogate all concern associated with using a
subchronic study as the basis of the RfD, but is sufficient to reduce the uncertainty factor from a
10 to a 3. The neurological effects of DCA in the principal study are a concern as is the fact that
there are no data on the mechanism for the observed neurological or testicular effects.
Neurological effects were seen in humans and other animal species (rats, dogs) at doses
comparable to the LOAEL in the Cicmanec et al. (1991) study. They were severe enough in
human patients to alter the clinical treatment regime. About 20 to 50% of patients with lactic
acidosis experience sedative effects with single and repeated oral or intravenous doses of 25 to 50
mg/kg-day. The effects were reversed with the withdrawal of DCA, but in some patients reversal
was slow (Stacpoole et al., 1998a). The effects on the nervous system seen in dogs involved
vacuolization of myelin. This observation can be mechanistically linked to the decreased nerve
conduction velocity observed in human subjects (Spruijt et al., 2001) because nerve impulses
travel faster in myelinated nerves. Therefore, the use of an uncertainty factor of 3 to extrapolate
from subchronic to chronic exposures is appropriate. There are no data that permit an assessment
of the clinical progression of the neurological effects. The data on testicular effects could be
more robust, but are mitigated by the lack of testicular histopathology in the DeAngelo et al.
(1996) rat cancer study. Rats are susceptible to testicular effects as a result of DCA exposure
33
-------
Addendum to the IRIS lexicological Review for Dichloroacetic Acid
(Linder et al., 1997; Toth et al., 1992), but the data indicate they are less sensitive to this effect
than dogs.
Step 2: Determination of a Drinking Water Equivalent Level (DWEL) for Dichloroacetic Acid
DWEL = (0.004 mg/kg/dav^) (JOke.} =0.14 mg/L (rounded to 0.1 mg/L)
(2 L/day)
where:
0.004 mg/kg/day = RfD
70 kg = assumed body weight of an adult
2 L/day = assumed drinking water consumption of a 70-kg adult
A lifetime HA has not been determined for DCA because it has been shown to cause liver
tumors in rats and mice exposed to high concentrations in drinking water. Based on the
Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999), EPA considers DCA as likely to be
carcinogenic to humans. Lifetime HA values are not provided for chemicals, like DCA, that
appear to lack a threshold for the tumorigenic response.
C. Carcinogenic Effects
C.I Characterization of Carcinogenic Potential
EPA considers that DCA is likely to be carcinogenic to humans. This descriptor is based
on the strength of the evidence in animal bioassays. In particular, the following characteristics
were considered and found to increase the overall weight of evidence for this descriptor: the
number of independent studies reporting consistently positive results and at roughly comparable
doses, site concordance for tumor formation between two species, consistent observations in
different species and sexes, and clear evidence of a dose-response relationship (U.S. EPA, 2003).
The reader is referred to the Toxicological Review (page 89) for the details of the cancer
assessment. The text of the summary provided below is that from the on-line IRIS summary for
DCA.
Multiple studies in B6C3FJ mice were performed which investigated the carcinogenic
response to oral exposures of DCA and established that increased incidence of hepatic tumors
resulted. Of these studies, the best one for cancer dose-response was deemed to be that of
DeAngelo et al. (1999), because the study was specifically designed to establish a multipoint
dose-response curve and data were available for five dose groups plus a control. The duration of
the study spanned the expected lifetime of a mouse. The administered doses in the mouse study
were converted to human equivalent doses using a dose scaling factor of BW075.
34
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
The dose-response curve was modeled using the Benchmark Dose-Response (BMD)
method. The high-dose was excluded from the modeling because the dose approached the MTD
of the study, as evidenced by the significantly increased hepatic necrosis indices throughout the
study and the decreases in body weight gain from 52 weeks on (a sensitivity analysis on the BMD
results indicated that no significant changes resulted from exclusion of the highest dose). Nine
different models were used to develop BMD values corresponding to a 10% benchmark response
(BMR). The 95% lower confidence limit on the central tendency estimate was designated the
Benchmark Dose Low (BMDL10). This value was used to develop the Cancer Slope Factor (CSF).
The multistage and quantal-quadratic models provide essentially identical fits to the data.
The multistage model estimate was selected for dose-response extrapolation because the quantal-
quadratic model has no first-order term and therefore may predict zero slope at zero dose. Given
the uncertainty surrounding the carcinogenic mechanism of DCA, it was decided that the zero
slope assumption was not justified. The multistage model gives a BMD10 of 6.9 mg/kg-day and a
BMDL10 estimate of 2.1 mg/kg-day (2.05, rounded to two significant figures). This is defined as
the point of departure (POD). Consistent with EPA's Guidelines for Carcinogen Risk Assessment
(U.S. EPA, 1999), extrapolation from the POD to low dose is performed by assuming a linear
dose-response curve between the POD and the origin. Based on this, the slope factor for DCA is
calculated as follows:
Slope factor = BMR / POD (BMDL10) = 0.1 / (2.1 mg/kg-day) = 0.048 (mg/kg-day)-1
The slope factor derived from the central tendency estimate of the cancer response is simply the
BMR divided by the BMD.
Slope Factor = BMR / BMD = 0.1 / (6.9 mg/kg-day) = 0.015 (mg/kg-day)'1
The drinking water unit risk is 1.4E-06 per |ig/L. This value is based on an ingestion rate
of 2L/day and a body weight of 70 kg. The concentrations in drinking water equivalent to a E-4
(1 in 10,000), E-5 (1 in 100,000) and E-6 (1 in 1,000,000) risk are 70 |ig/L, 7.0 |ig/L and 0.7
|ig/L, respectively.
The CSF for DCA was determined from the dose-response data on adenomas and
carcinomas from a study in male B6C3Fj mice by DeAngelo et al. (1999). The study is
representative of a larger cancer database in mice and provides data for five dose groups in
addition to the control. It is accordingly well suited to modeling using the BMD approach. There
is uncertainty regarding the suitability of mice as a model for human carcinogenesis because of
the high spontaneous liver tumor response in this species.
DCA has been used therapeutically in humans at doses as low as 25 mg/kg-day (Stacpoole
et al., 1998a,b; Spruijt et al., 2001). However, carcinogenic endpoints following these exposures
have not been evaluated. To extrapolate mouse tumor data for DCA to the human situation, it is
35
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Addendum to the IRIS lexicological Review for Dichloroacetic Acid
assumed that humans will respond similarly to the mouse. It is not clear that this is the case.
Nevertheless, the robust tumor response in male rats, following ingestion exposures to DCA,
increases the confidence that DCA can elicit liver tumors across different species. In the absence
of human data, the choice for using the mouse data from DeAngelo et al. (1999) is a reasonable
assumption and is protective of human health.
Overall, there are inadequate data to support any conclusive mode of carcinogenic action.
Some studies indicate that DCA is genotoxic (Fuscoe et al., 1996; Harrington-Brock et al., 1998;
Leavitt et al., 1997) but the doses required to induce these effects are very high and far exceed
exposure concentrations that are relevant to humans. Studies are conflicting regarding lipid
peroxidation and the production of oxygenated radicals; further, cancer is induced at doses lower
than those required to induce lipid peroxidation in the rat. Data suggest that tumors can originate
from several different cell lines and through more than one pathway (Carter et al., 2003). Findings
related to glycogen accumulation and necrosis as they apply to the tumorigenic response are not
clear. Given the uncertainty regarding the mode of action, the dose-response was modeled using
a BMD approach which assumed linearity at low doses. This decision is protective of public
health. The cancer risk estimation presented for DCA is considered to be protective of
susceptible groups, including children.
D. Summary
Table III-3 summarizes HA and DWEL values that have been derived from available
toxicological dose-response data for DCA.
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Addendum to Drinking Water Criteria Document for Dichloroacetic Acid
Table III-3. Summary of Development of the HAs and DWEL for DCA
Study
NOAEL/
LOAEL
(mg/kg/day)
UFH
UFA
UFS
UFL
UFD
Composite
Factor
RfD
Equivalent
(mg/kg/day)
Final
Value
(mg/L)
One-day HAa
Default to Ten-day HA
3b
Ten-day HAa
Parrish et
al. (1996)
25/125
10
10
-
-
-
100
3
Longer-Term HA
Cicmanec
etal. (1991)
DWEL
Cicmanec
etal. (1991)
ND/12.5
ND/12.5
10
10
3
3
-
3
10
10
3
3
1000
3000
0.01
0. 1/0.4 c
0.004
0.1
ND, not determined
a Database uncertainty factors are not applied for One-day or Ten-day Health Advisories per Office of Water policy.
b The One-day Health Advisory value was derived from the Ten-day Health Advisory value.
c Child/Adult
37
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Addendum to Drinking Water Criteria Document for Dichloroacetic Acid
Chapter IV. References
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Addendum to Drinking Water Criteria Document for Dichloroacetic Acid
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Addendum to Drinking Water Criteria Document for Dichloroacetic Acid
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Addendum to Drinking Water Criteria Document for Dichloroacetic Acid
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Addendum to Drinking Water Criteria Document for Dichloroacetic Acid
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Addendum to Drinking Water Criteria Document for Dichloroacetic Acid
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