United States    Office of Water    EPA-822-R-04-002
     Environmental Protection  Office of Science and Technology December 2003
     Agency     4304T
4>EPA Ambient Water Quality Criterion for
     the Protection of Human Health:
     Chloroform - Revised Draft



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                             EPA-823-D-03-001
                               December 2003
Water Quality Criterion
  for the Protection of
     Human Health:

       Chloroform

       Revised Draft
    Office of Science and Technology
         Office of Water
   U.S. Environmental Protection Agency
       Washington, DC 20460

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                               NOTE TO READER

The Agency is developing streamlined criteria documents which focus on critical
toxicological and exposure-related studies only. This is a departure from the past format
in which all existing toxicological and exposure studies were presented and evaluated,
with equal emphasis placed on exposure, pharmacokinetics, toxicological effects, and
criterion formulation.  Due to limited resources and a need to update criteria as quickly
as possible, EPA has decided to develop more abbreviated versions of criteria
documents with an emphasis on using existing risk assessments (on IRIS or other EPA
health assessment documents) where available and still relevant, and to focus to a
greater extent on pertinent exposure and toxicological studies that may influence the
development of a criterion (e.g., critical effects studies which form the basis of RfD
development or cancer assessment).  EPA will continue to conduct a comprehensive
review of the literature for the latest studies,  but will not provide a summary or an
evaluation of the studies that are deemed less significant in the criteria development
process.  Where there is a significant amount of literature on an area of study (for
instance, pharmacokinetics), EPA, to the extent possible, will reference the information
or cite existing documents (e.g., IRIS or other existing EPA risk assessment documents)
that discuss the information in greater detail.

The overall objective of this change in philosophy is to allow EPA to update 1980
ambient water quality criteria (AWQC) at a greater frequency, while  still maintaining the
scientific rigor that EPA requires when developing an AWQC.  EPA  believes these
"new" criteria documents will be as informative as previous criteria documents and will
continue to serve as the key scientific basis for State and Tribal standards.  EPA also
believes the documents will provide the necessary scientific content and scope to allow
a State or Tribe to come to an appropriate technical and/or policy decision with regard
to setting water quality standards.

EPA requests that commenters identify any relevant information missing from this
criteria document which may result in a different criteria calculation or scientific
interpretation.

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                           TABLE OF CONTENTS

1.  INTRODUCTION	  1

2.  CHEMICAL AND PHYSICAL PROPERTIES	  3

3.  SUMMARY OF TOXICOKINETICS	  5
      3.1   ABSORPTION	  5
      3.2   DISTRIBUTION	  5
      3.3   METABOLISM	  6
      3.4   EXCRETION	  7

4.  TOXICOLOGICAL AND  RISK  BASES FOR CRITERIA  	  9
      4.1   NONCANCER DATA AND PREVIOUS EVALUATIONS	  9
           4.1.1  Oral Exposure	  9
                 4.1.1.1  Oral Exposure Data in Humans	  9
                 4.1.1.2  Oral Exposure Data in Animals	  9
                 4.1.1.3  Oral Reference Dose	 13
           4.1.2  Inhalation Exposure  	 15
                 4.1.2.1  Use of Chloroform as an Anesthetic 	 15
                 4.1.2.2  Inhalation Studies in the Workplace 	 16
                 4.1.2.3  Inhalation Studies in Animals	 16
                 4A.2.4  Inhalation Reference Concentration 	 18
      4.2   CANCER EVALUATION	 18
           4.2.1  Oral Exposure	 18
                 4.2.1.1  Studies in Humans  	 18
                 4.2.1.2  Studies in Animals	 19
           4.2.2  Inhalation Exposure  	 22
                 4.2.2.1  Studies in Humans  	 22
                 4.2.2.2  Studies in Animals	 22
           4.2.3  Mutagenicity	 22
           4.2.4  Cancer Evaluation Using Current Guidelines	 24
                 4.2.4.1  Mode of Action	 25
                 4.2.4.2  Quantification of Cancer Risk from  Oral Exposure  .... 28
                 4.2.4.3  Quantification of Cancer Risk from  Inhalation Exposure 28
           4.2.5  Discussion of Confidence	 29

5.  EXPOSURE ASSUMPTIONS  	 3_1
      5.1   RELATIVE SOURCE CONTRIBUTION ANALYSIS 	 3_1
           5.1.1  Population of Concern  	 3J.
           5.1.2  Overview of Potential for Exposure  	 32
                 5.1.2.1  Ambient Water	 32
                 5.1.2.2  Treated Water	 32
                 5.1.2.3  Non-Water Sources	 34

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           5.1.3 Estimates of Exposure from Non-Ambient Water Sources	 34
                5.1.3.1 Exposure from Treated Drinking Water	 34
                5.1.3.2 Exposures from Outdoor Air 	 37
                5.1.3.3 Dietary Exposures	 38
                5.1.3.4 Total Non-Ambient Water Exposures 	 38
           5.1.4 Estimates of Exposure from Ambient Water Sources	 38
                5.1.4.1 Ingestion of Ambient Water	 38
                5A.4.2 Ingestion of Freshwater and Estuarine Fish 	 4J.
     5.2   EXPOSURE DATA ADEQUACY AND UNCERTAINTY ESTIMATES .  . 41
     5.3   RSC ESTIMATES/ALLOCATION OF THE RfD	 41
     5.4   EXPOSURE ASSUMPTIONS FOR AMBIENT WATER 	 42
           5.4.1 Exposure from Ambient Water Used for Drinking 	 43
           5.4.2 Exposures from Fish Ingestion	 43

6.  BIOACCUMULATION FACTORS  	 45
     6.1   BASELINE BAFs	 45
           6.1.1 Summary of Field-derived BAF and Laboratory-measured BCF
                Data  	 46
           6.1.2 Derivation of Baseline BAFs (BAFds) 	 51
     6.2   National BAFs	 53
           6.2.1 Baseline BAFs (Baseline BAFd)  	 54
           6.2.2 Lipid Content of Consumed Aquatic Species	 54
           6.2.3 Freely-Dissolved Fraction Applicable to AWQC	 54
           6.2.4 Calculation of National BAFs  	 56

7.  AWQC CALCULATION	 57
     7.1   FOR AMBIENT WATERS USED FOR DRINKING WATER
           SOURCES   	 57
     7.2   FOR AMBIENT WATERS NOT USED FOR DRINKING WATER
           SOURCES   	 59
     7.3   AWQC SUMMARY	 59

8.  SITE-SPECIFIC OR REGIONAL ADJUSTMENTS TO CRITERIA 	 61

9.  REFERENCES 	 63
                                   IV

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                             1.  INTRODUCTION

Ambient Water Quality Criteria (AWQC) for chloroform were originally set in 1980 based
on non-threshold carcinogenic effects (45 FR 79318). Because a non-threshold
carcinogenic effect implies a risk of cancer at any concentration greater than zero, the
ideal target concentration identified for chloroform in water was zero.  However,
because a level of zero may not be attainable, criteria were set based incremental
increases in cancer risk.  The criteria for an incremental risk of 10"6 were 0.19 ug/L for
ingestion of water and organisms, and 15.7 ug/L for ingestion of organisms only.

Those criteria were revised in 1992 as part of the National Toxics Rule (57 FR 60848),
based on a revised carcinogenic slope factor published in EPA's Integrated Risk
Information System (IRIS) in 1988. The revised criteria for an incremental risk of 10"6
were 5.7 ug/L for ingestion of water and organisms, and 470 ug/L for ingestion of
organisms only. These criteria are included in EPA's 2000 compilation of AWQC
(USEPA2000a).

As required under Section 304(a) of the Clean Water Act, EPA must periodically revise
criteria for water quality to accurately reflect the latest scientific knowledge on the kind
and extent of all identifiable effects on human health from the presence of pollutants in
any body of water.  This criteria document updates the criteria for chloroform using new
methods and new information described in the Methodology for Deriving Ambient Water
Quality Criteria for the Protection of Human Health (USEPA 2000b) and in the
Methodology's accompanying Federal Register Notice (USEPA 2000c). These include
new methods to determine toxicity dose-response relationships for both carcinogenic
and noncarcinogenic effects, updated exposure factors (e.g., values for fish
consumption), new exposure assumptions used in the calculation, and new procedures
to determine bioaccumulation factors. The Risk Assessment Technical Support
Document (TSD) accompanying the Federal Register notice describes the
determination of toxicity dose-response relationships in greater detail (USEPA 2000d).
In addition to the new methods and information described above, new information on
toxicity, exposure, and bioaccumulation of chloroform is also included in this update.

Based on the most sensitive toxicity endpoint (cytolethality and regenerative hyperplasia
in liver), the revised criterion is 68 ug/L to protect against  ingestion of chloroform in
drinking water and aquatic organisms, or 2,400 ug/L to  protect against ingestion of
chloroform in aquatic organisms alone.  Based on available data on the mode of action,
these criteria will protect exposed humans against both the noncancer and cancer
effects of chloroform. The calculation is based on adults in the general population.

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The following sections describe the toxicological, exposure, and bioaccumulation factor
evaluations, and the calculation of the chloroform criteria.
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              2.  CHEMICAL AND PHYSICAL PROPERTIES
Revised AWQC are being derived for chloroform (CAS No. 67-66-3). The chemical
formula is CHCI3, and the chemical structure is:
      Cl
H

C — Cl

Cl
Synonyms include trichloromethane and methane trichloride.

      Physical and chemical properties (Howard and Meylan 1997, Montgomery and
      Welkom1989):
      Chemical formula
      Molecular weight
      Physical State (25°C)
      Boiling Point
      Density (20°C)
      Vapor pressure(25°C)
      Specific Gravity (20°C/4°C)
      Water solubility:
      Log Octanol:Water Partition Coefficient
      Conversion factor (air concentrations)
                                  CHCI3
                                  119.38
                                  Colorless, volatile liquid
                                  61.7°C
                                  1.484g/ml_
                                  197 mm Hg
                                  1.483
                                  7.95g/Lat25°C
                                  1.97
                                  1 ppm = 4.88 mg/m3
                                  1 mg/m3 = 0.205  ppm

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                   3.  SUMMARY OF TOXICOKINETICS

Data on the toxicokinetics of chloroform have been reviewed and summarized by
USEPA (2001 a). The text below is derived from USEPA (2001 a) with little modification.

3.1    ABSORPTION

Studies in animals indicate that gastrointestinal absorption of chloroform is rapid (peak
blood levels at about 1 hour) and extensive (64% to 98%) (USEPA 1997a, ILSI 1997,
USEPA 1998c). Limited data indicate that gastrointestinal absorption of chloroform is
also rapid and extensive in humans, with over 90% of an oral dose recovered in expired
air (either as unchanged chloroform or carbon dioxide) within 8 hours (Fry et al. 1972).

Most studies of chloroform absorption following oral exposure have used oil-based
vehicles and gavage dosing (USEPA 1994, 1998c). This is of potential significance
because most humans are exposed to chloroform via ingestion in drinking water, and
drinking water ingestion usually occurs intermittently throughout the day rather than as a
single dose (bolus). Although the effect of administration of chloroform in oil is to slow
absorption compared to gavage administration in water (Withey et al. 1983), the effect
of gavage (bolus) dosing  is to increase blood concentrations compared to intermittent
exposures.  Thus, extrapolation of dose-response data from non-drinking water studies
should be done with caution because of potential pharmacokinetic differences.

Dermal and inhalation absorption of chloroform by humans during showering was
investigated by Jo et al. (1990). Chloroform concentrations in exhaled breath were
measured in six human subjects before and after a normal shower, and following
inhalation-only shower exposure. Breath levels measured at 5 minutes following either
exposure correlated with tap water  levels of chloroform.  Breath levels following
inhalation exposure only were about half those following a normal shower (both
inhalation and dermal contact). These data indicate that humans absorb chloroform by
both the dermal and inhalation routes (USEPA 1994).

3.2   DISTRIBUTION

Absorbed chloroform appears to distribute widely throughout the body (USEPA 1994,
1998c). In postmortem samples from eight humans, the highest levels of chloroform
were detected in the body fat (5-68 g/kg), with lower levels (1-10  g/kg) detected in the
kidney, liver, and brain (McConnell  et al. 1975). Studies in animals indicate  rapid
uptake of chloroform by the liver and kidney (USEPA 1997a).

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3.3   METABOLISM

Chloroform is metabolized in humans and animals by cytochrome P450-dependent
pathways.  In the presence of oxygen (oxidative metabolism), the chief product is
trichloromethanol, which rapidly and spontaneously dehydrochlorinates to form
phosgene (CCI20):
2 CHCI3 + NADPH + H+ + 0  •  «2 CCIOH + NADP+
   I3OH • •CCI2
            3                  2
      CCIOH • •CCI0 + HCI
In the absence of oxygen (reductive metabolism), the chief metabolite is dichloromethyl
free radical (CHCI2 •) (USEPA 1997a, ILSI 1997).

Nearly all tissues of the body are capable of metabolizing chloroform, but the rate of
metabolism is greatest in liver, kidney cortex, and nasal mucosa (ILSI 1997). These
tissues are also the principal sites of chloroform toxicity, indicating the importance of
metabolism in the mode of action of chloroform toxicity.

At low chloroform concentrations, metabolism occurs primarily via cytochrome
P450-2E1  (CYP2E1) (Constan et al. 1999).  The level of this isozyme (and hence the
rate of chloroform metabolism) is induced by a variety of alcohols (including ethanol)
and ketones, and may  be inhibited by phenobarbital.  At high chloroform
concentrations, metabolism is also catalyzed by cytochrome P450-2B1/2 (CYP2B1/2)
(ILSI 1997; USEPA 1997a, 1998c). Because chloroform metabolism is
enzyme-dependent, the rate of metabolism displays saturation kinetics.  Under low
dose-rate conditions, nearly all of a dose is metabolized. However, as the dose or the
dose rate increases, metabolic capacity may become saturated and increasing fractions
of the dose are excreted as the un-metabolized  parent (Fry et al. 1972).

The products of oxidative metabolism (phosgene) and reductive metabolism
(dichloromethyl free radical) are both highly reactive. Phosgene is electrophilic and
undergoes attack by a  variety of nucleophiles. The predominant reaction is hydrolysis
by water, yielding carbon dioxide and hydrochloric acid. The rate of phosgene
hydrolysis is very rapid, with a half-time of less than one second (De Bruyn et al. 1995).
Phosgene also reacts with a wide variety of other nucleophiles, including primary and
secondary amines, hydroxy groups, and thiols (Schneider and Diller 1991).  For
example, phosgene reacts with the thiol group of glutathione and with nucleophilic
groups (-SH, -OH, -NH2) in cellular macromolecules such as enzymes, proteins, or the
polar heads of phospholipids, resulting in formation of covalent adducts (Pohl et al.
1977, 1980, 1981; Pereira and Chang 1981; Pereira et al.1984; Noort et al 2000).

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Formation of these molecular adducts can interfere with molecular function (e.g., loss of
enzyme activity), which in turn may lead to loss of cellular function and subsequent cell
death (ILSI 1997, WHO 1998).

Free radicals that are formed under conditions of low oxygen are also extremely
reactive, forming covalent adducts with  microsomal enzymes and the fatty acid tails of
phospholipids,  probably quite close to the site of free radical formation (cytochrome
P450 in microsomal membranes).  This results in a general loss of microsomal enzyme
activity, and can also  result in lipid peroxidation (ILSI 1997, USEPA 1998c).

Two lines of evidence suggest that metabolism occurs mainly via the oxidative pathway.
First, reductive metabolism of chloroform is observed only in phenobarbital-induced
animals or in tissues prepared from them, with negligible reducing activity observed in
uninduced animals (ILSI  1997). Second, in vitro studies using liver and kidney
microsomes from mice indicate that, even under relatively low (2.6%) oxygen partial
pressure (approximately  average for the liver), more than 75% of the phospholipid
binding was to the fatty acid heads. This pattern of adduct formation on phospholipids
is consistent with phosgene, not free radicals, as the main reactive species, indicating
metabolism was chiefly by the oxidative pathway (USEPA 1998c, ILSI 1997).  Addition
of glutathione to the incubation system completely  negated binding to liver microsomes,
with only residual binding remaining in kidney microsomes (ILSI 1997). This quenching
by glutathione is expected for the  products of oxidative but not reductive metabolism.
Taken together, these observations strongly support the conclusion that chloroform
metabolism in vivo occurs primarily via the oxidative pathway except under special
conditions of high chloroform doses in pre-induced animals (ILSE 1997, USEPA 1998c).

3.4    EXCRETION

Excretion of chloroform occurs primarily via the lungs (USEPA 1998c).  Results from
studies in humans indicate that approximately 90% of an oral dose of chloroform was
exhaled (either as chloroform or as carbon dioxide), with less than 0.01 % of the dose
excreted in the urine (USEPA 1994).  In mice and rats,  45 to 88% of an oral dose of
chloroform was excreted from the lungs either as chloroform or carbon dioxide, with
1-5% excreted in the urine (USEPA 1998c).

No data are available regarding the bioaccumulation or retention of chloroform following
repeated exposure. However, due to the rapid excretion and metabolism of chloroform,
combined with low levels of chloroform detected in human postmortem tissue samples,
marked accumulation and retention of chloroform is not expected (USEPA 1994).

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        4. TOXICOLOGICAL AND RISK  BASES FOR  CRITERIA

Data on the oral and inhalation toxicity of chloroform have been reviewed and
summarized by USEPA (2001 a).  The text below is derived from USEPA (2001 a) with
little modification.

4.1    NONCANCER DATA AND PREVIOUS EVALUATIONS

The noncancer effects of chloroform have been investigated in a large number of
studies in animals, with some supporting data from studies in humans.  Oral exposure
to chloroform is associated mainly with cytotoxicity to cells in the liver and kidney.
Inhalation exposure is also associated with injury to liver and kidney, with neurological
effects also occurring at high levels. Some of the studies supporting these conclusions
are summarized below.

4.1.1  Oral Exposure

4.1.1.1  Oral Exposure Data in Humans

There have been several epidemiological studies that have investigated the association
between human exposure to chloroform and other disinfection byproducts in chlorinated
water and the occurrence of adverse reproductive outcomes (Kramer et al. 1992, Bove
etal. 1995, Gallagher et al. 1998, Waller et al. 1998). Statistically significant
correlations were observed between exposure to total trihalomethanes and one or more
adverse reproductive outcomes, and in one case (Kramer et al. 1992), there was a
significant  relationship between chloroform levels and decreased intrauterine growth.  In
another case (Waller et al. 1998), an association  was noted between increased risk of
spontaneous abortion and bromodichloromethane (but not chloroform) levels.  Although
epidemiological studies of this type are useful in evaluating whether chlorinated drinking
water can increase the risk of adverse reproductive effects in exposed populations, the
studies are not adequate to establish a causal link between ingestion of chloroform and
the occurrence of adverse reproductive effects in humans, because chlorinated drinking
water contains many different potentially toxic disinfection byproducts.

4.1.1.2  Oral Exposure Data in Animals

Eschenbrenner and Miller (1945) exposed Strain  A mice (5/sex/group) to chloroform at
dose levels of 0, 150, 300, 600, 1200, or 2400 mg/kg-dat in olive oil by gavage.  The
animals were dosed  every 4 days over a period of 120 days (a total of 30 doses).  No

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males administered doses of at least 600 mg/kg and no females in the high-dose group
survived the study. All deaths occurred 24 to 48 hours after the first or second chloro-
form dose.  Liver necrosis was observed in both sexes in the three highest dose groups.
Males in all treatment groups developed kidney necrosis, whereas kidney necrosis was
not apparent in any females.  The severity of renal necrosis in males was dose related.

Palmer et al. (1979) exposed Sprague-Dawley rats (50/sex/group) to chloroform in
toothpaste by gavage at doses of 0 or 60 mg/kg-day, 6 days/week for 80 weeks. No
significant differences in mortality were observed between treated and control animals.
A marginal decrease in body weight gain (about 10%) was observed in both treated
males and females when compared with controls. A statistically significant decrease in
relative liver weight was observed in treated females. Histologic examination of the liver
revealed only minor changes, with no severe fatty infiltration, fibrosis, or marked bile
duct abnormalities reported. The incidence of moderate to severe glomerulonephritis
was reported to be slightly increased in treated males.

Heywood et al. (1979) exposed groups of eight male and eight female beagle dogs to
chloroform at doses of 15 or 30 mg/kg-day.  The chemical was given orally in a
toothpaste base in gelatin capsules, 6 days/week for 7.5 years. This was followed by a
20- to 24-week recovery period.  A group of 16 male and 16 female dogs received
toothpaste base without chloroform and served as the vehicle control group.  Eight dogs
of each sex served as an untreated group and a final group of 16 dogs (8/sex) received
an alternative non-chloroform toothpaste.  Four male dogs (one each from the low- and
high-dose chloroform groups, the vehicle control group, and the untreated control
group) and seven female dogs (four from the vehicle control group and three from the
untreated control group) died during the study. Although there was substantial
variability in  individual measurements, serum glutamate pyruvate transaminase (SGPT,
now known as alanine aminotransferase or ALT) levels tended to be about 30%-50%
higher in the low-dose group (15 mg/kg/day) than in control animals. These increases
were statistically significant for weeks 130-364.   For the high-dose group (30 mg/kg-
day), the typical increase in SGPT was about twofold, and the differences were
statistically significant for the entire exposure duration (weeks 6-372).  After 14 weeks
of recovery,  SGPT levels remained significantly increased in the high-dose group but
not in the low-dose group, when compared with the controls. After 19 weeks of
recovery, SGPT levels were not significantly increased in either treated group when
compared with the controls.  The authors concluded that the increases in SGPT levels
were likely the result of minimal liver damage. Serum alkaline phosphatase (SAP) and
SGPT levels were also moderately increased (not statistically significant) in the treated
dogs at the end of the treatment period when compared with the controls. Microscopic
examinations were conducted on the major organs.  The most prominent microscopic

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effect observed in the liver was the presence of "fatty cysts," which were described as
aggregations of vacuolated histiocytes. The fatty cysts were observed in the control
and treated dogs, but were larger and more numerous (i.e., higher incidence of cysts
rated as "moderate or marked," as opposed to "occasional or minimal") in the treated
dogs at both doses than in the control dogs. The prevalence of moderated or marked
fatty cysts was 1/27 in control animals, 9/15 in low-dose animals,  and 13/15 in high-
dose animals. Nodules of altered hepatocytes were observed in both treated and
control animals, and therefore were not considered related to treatment. No other
treatment-related nonneoplastic or neoplastic lesions were reported for the liver, gall
bladder, cardiovascular system, reproductive system, or urinary system. A NOAEL was
not identified in this study. However, a LOAEL of 15 mg/kg/day was identified, based
on elevated SGPT levels and increased incidence and severity of fatty cysts (USEPA
1998c).

Jorgenson and Rushbrook (1980) exposed seven groups of 6-week-old female B6C3F1
mice (30 mice/group) to drinking water containing either 0, 200, 400, 600, 900, 1800, or
2700 ppm chloroform for 30-90 days.  Calculated dose levels were 0, 32, 64, 97, 145,
290, or 436 mg/kg-day based on reported water intakes.  At week 1, a significant
decrease in body weight was observed in the 900, 1800, and 2700 ppm  chloroform
treatment groups; however, all body weights of the treated animals were comparable to
controls after week 1.  On days 30, 60, and 90, ten animals from each treatment group
were sacrificed for gross  and microscopic pathologic examination, as well as for
measurement of organ fatorgan weight ratios.  A 160%-250% increase in liver fat was
observed in the high-dose group. Histological examination of the liver revealed mild
centrilobular fatty changes in the 1,800 and 2,700 ppm  groups. On day 30, reversible
fatty changes  in the liver were observed at doses as low as 400 ppm chloroform.
Treatment-related atrophy of the spleen was observed at the high dose. Based on the
observation of mild effects of chloroform exposure via the drinking water on liver and
other tissues,  the LOAEL in this study was 290 mg/kg-day, while the NOAEL was 145
mg/kg-day (USEPA 1994).

Jorgenson et al. (1982) exposed male Osborne-Mendel rats and female B6C3F1 mice
to chloroform  in drinking water (0, 200, 400, 900, or 1,800 mg/L) for 1 to 6  months.  The
time-weighted average doses, based on measured water intake and body weights, were
0, 19, 38, 81,  or 160 mg/kg-day in  rats and 0, 34, 65, 130, or 263 mg/kg-day in mice.
An additional group of matched controls received the same water volume as the high-
dose groups.  In male rats, some changes were observed in body weight and in some
hematological and serum biochemical parameters, but the authors judged  these
changes to be a secondary effect of reduced water intake.  Gross and microscopic
pathology findings in the rats generally were slight or mild in severity, were not dose

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related, and either appeared adaptive (occurred in rats sacrificed after 30 or 60 days,
but not in those sacrificed after 90 days) or were sporadic (by nature and/or incidence)
and not considered treatment-related.  This study identifies a NOAEL of 160 mg/kg-day
in the male rat. In mice, mortality within the first 3 weeks was significantly increased in
the two highest dose groups (130 and 263 mg/kg-day), but was comparable to controls
after that time. Early mortality and behavioral effects (e.g., lassitude, lack of vigor) were
apparently related to reduced water consumption. A significant increase in liver fat in
mice was noted at doses of 65 mg/kg-day and higher at 3 months, and at doses of 130
and 263 mg/kg-day at 6 months. This study identifies a NOAEL of 34 mg/kg-day and  a
LOAEL of 65-130 mg/kg-day in mice, based on increased liver fat at 3 to 6 months.

In the original study reports by Jorgenson et al. (1982, 1985), histological findings
indicative of renal cytotoxicity were not included.   Recently, histological slides of rat
kidney from this study have been re-examined to assess whether evidence of renal
cytotoxicity could be detected (ILSI 1997,  Hard and Wolf 1999, Hard et al. 2000).
Based on this reexamination, it was found that animals exposed to average doses of 81
or  160 mg/kg-day of chloroform displayed low-grade renal tubular injury with
regeneration,  mainly in the mid to deep cortex. The changes included faint basophilia,
cytoplasmic vacuolation, and simple hyperplasia in proximal convoluted tubules. In
some animals, single-cell necrosis, mitotic figures, and karyomegaly were also
observed. Hyperplasia was visualized as an  increased number of nuclei crowded
together in tubule cross-sections. These changes were observable in the 160 mg/kg-
day dose group at 12, 18, and 24 months, and in the 81 mg/kg-day dose group at 18
and 24 months. Cytotoxic changes were not seen in either of the lower dose groups (19
or 38 mg/kg-day).  Based on histological evidence of renal cytotoxicity in rats,  this study
identifies a LOAEL of 81 mg/kg-day.

Bull et al. (1986) studied the effect of dose vehicle on the hepatotoxicity of chloroform
using male and female B6C3F1 mice.  Doses of 0, 60, 130, or 270 mg/kg-day in corn  oil
or  in 2% emulphor were administered via gavage for 90 days. Based on measurements
of serum enzyme levels, serum  and tissue triglyceride levels, and histological
examination of the livers, the authors concluded that hepatotoxic effects were enhanced
by the administration of chloroform via corn oil versus  chloroform administered in an
aqueous suspension.  The authors suggested that the cause may be absorption kinetics
or  interaction between chloroform and the corn oil vehicle (USEPA 1994). A LOAEL of
270 mg/kg-day was identified for chloroform when administered  in corn oil, but 270
mg/kg-day was considered a NOAEL for chloroform when administered in aqueous
vehicle (USEPA 1994).
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Larson et al. (1994a) investigated the effects of oral exposure of female B6C3F1 mice
to chloroform.  Exposure occurred either through drinking water (0, 60, 200, 400, 900, or
1800 ppm) or by gavage in corn oil (0, 3, 10, 34, 90, 238, or 477 mg/kg-day) for up to
three weeks.  All animals were examined for histological lesions in the liver and for
increases in hepatic labeling index (LI).  In animals exposed by corn oil gavage, doses
of 238 mg/kg-day or higher produced clear hepatic necrosis and increases in LI. No
histological effects were observed at 10 mg/kg-day, and no increase in LI was seen at
34 mg/kg-day. In animals  exposed to chloroform in water, there were no significant
effects on LI nor were there any significant histological lesions at any exposure  level,
even though the effective dose was as high from drinking water as from gavage.

Larson et al. (1995) studied the effects of chloroform exposure on liver, kidney,  and
nasal passage in female F344 rats. Exposure occurred by gavage in corn oil for up to
three weeks.  Dose levels  were 0, 34, 100, 200, or 400 mg/kg-day.  Mild degenerative
centrilobular changes and  increases in LI were observed in the liver at doses of 100
mg/kg-day or higher.  Degeneration and necrosis were observed in the kidney at doses
of 200 mg/kg-day and higher, and increases in LI in kidney were observed at 100
mg/kg-day and higher. Effects on the peripheral region of the nasal turbinates
(periostial hypercellularity and increased cell replication) were observed at all exposure
doses.

4.1.1.3 Oral Reference Dose

EPA used the available data on the non-cancer effects of chloroform to estimate oral
RfD values using two different approaches:  the traditional NOAEL-LOAEL approach
and the benchmark dose (BMD) modeling approach (USEPA 2001 a).  The results of
these two evaluations are  summarized below.

NOAEL-LOAEL Approach

For the NOAEL-LOAEL approach, the principal study selected to derive the RfD was the
report by Heywood et al. (1979),  in which there was an increase in the incidence of
moderate to marked hepatic fatty cysts in dogs. This study was selected because  it
identifies the lowest LOAEL and because it is also the longest duration study (7.5
years). The lesions observed in this study were characterized by aggregations of
vacuolated  histiocytes. Although fatty cysts were observed in the control group as  well
as all treated groups, both the size and severity of these lesions were significantly
increased in treated animals. The LOAEL for fatty cysts was 15 mg/kg-day, and a
NOAEL was not identified. The LOAEL of 15 mg/kg-day was used to derive a chronic
oral RfD for chloroform as  follows:
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         RfD __ 15 mg / Ag - day x^6 days / 7 days; __^_Q2 mg/kg_dgy
where:
      15 mg/kg-day =    LOAEL identified by Heywood et al. (1979)
      6 days/7 days =    Adjustment to account for exposure 6 days/week
      1,000 =            Uncertainty factor.  This uncertainty factor includes a factor
                        of 10 to extrapolate from a LOAEL to a NOAEL, a factor of
                        10 to extrapolate from an animal species (dog) to  humans,
                        and a factor of 10 to account for potential sensitive human
                        subpopulations.

Benchmark Dose Approach

In accord with EPA guidance (USEPA 1995), several data sets in addition to the data
set with the lowest LOAEL (Heywood et al. 1979) were selected for benchmark dose
(BMD) modeling. This is because the study that identifies the lowest LOAEL  may not
always be suitable for modeling or might not always yield the lowest BMD.  The data
sets selected for modeling included (a) incidence of fatty cysts in liver of dogs (Heywood
et al. 1979), (b) histological evidence of renal cytotoxicity  in male rats exposed via
drinking water (Hard et al. 2000), (c) increased labeling index in kidney of female mice
exposed via drinking water (Larson et al. 1994a), and  (d)  increased labeling index in
liver of female rats exposed via gavage in corn oil (Larson et al.  1995).

Based on the BMD fitting, a BMDL value of 1.2 mg/kg-day derived from the study by
Heywood et al. (1979) was selected as the most appropriate basis for the derivation of
the RfD.  Because this value is based on exposures that occurred 6 days/ week, the
value was adjusted as follows:

              BMDL = (1.2 mg/kg-day) x (6/7) = 1.0 mg/kg-day.

The RfD was derived from the BMDL by application of appropriate uncertainty factors.
In this case, an uncertainty factor of 10 is used to account for interspecies extrapolation
and a factor of 10 is used to protect potentially sensitive human subpopulations.
Additional uncertainty factors are not required  because the database for chloroform is
complete.  Bioassays are available  in the dog  (Heywood et al. 1979) and the  rat and
mouse (NCI 1976; Jorgenson et al. 1982, 1985). Developmental toxicity studies are
available in rats and  rabbits exposed via the oral route (Thompson et al. 1974), and in
rats (Baeder and Hoffmann 1988, 1991; Schwetz et al. 1974; Stanford Research

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Institute 1978) and mice (Murray et al. 1979) exposed by the inhalation route. These
studies indicate that effects on the fetus do not occur except at doses that cause
maternal toxicity. A  two-generation reproduction study (NTP 1988) found no effects on
fertility or reproduction at doses that resulted in liver histopathology.  Finally, chloroform
is rapidly metabolized and excreted and thus is not expected to bioaccumulate.  Based
on all of these considerations, a total uncertainty factor of 100 was applied and the
resulting RfD was 1E-02 mg/kg-day:

                 RfD = 1.0 mg/kg-day 1100 = 1E-02 mg/kg-day.

In general, the choice between the NOAEL-LOAEL approach and the BMD approach
depends on the quality of data and confidence in the fitting results. However, in this
particular case, the two approaches (NOAEL/LOAEL, benchmark) yield equal RfD
values.

4.1.2 Inhalation Exposure

4.1.2.1  Use of Chloroform as an Anesthetic

Chloroform was used from around 1850 to 1950 as an anesthetic in midwifery and
surgery.  It was generally preferred to ether because of its ease of use and its
non-flammable nature.  However, in the initial period of use, a number of deaths
occurred, which were mainly attributed to cessation  of the heart following over-exposure
to chloroform vapor  (Davidson 1965).  Once the potential for harm was recognized,
physicians and scientists at the time recommended that use of chloroform as an
anesthetic be restricted to airborne concentrations in the 2%-4% range (20,000-40,000
ppm). Whitaker and Jones (1965) reported on the outcome of over 1,500 patients who
had undergone chloroform anesthesia in more recent years, with the highest
concentration being  2.25% (22,500 ppm). Effects observed following administration of
chloroform included  increased respiratory rate (44%), bradycardia (the incidence
increasing from 7% at less than 30 minutes exposure to 25% at more than 120 minutes
exposure), cardiac syncope (1%), hypotension (33%), nausea and vomiting (7%),
jaundice (1 patient, probably from hepatitis), and death (one patient, probably from
infection). This body of clinical experience demonstrates that central nervous system
depression is the chief effect of acute inhalation exposure to high concentrations of
chloroform, and that clinically significant effects on target tissues such as liver are
unlikely  to be observed following brief exposures (generally less than  a few hours).
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4.1.2.2 Inhalation Studies in the Workplace

Several epidemiological studies have been performed to investigate the occurrence of
adverse effects in populations of workers exposed to chloroform vapors in the
workplace (Bomski et al. 1967, Challen et al. 1958, Phoon et al. 1983, Li et al. 1993).
All of these studies are limited by lack of detailed exposure information for the study
subjects, and hence the available  data are not adequate to define with confidence the
inhalation dose-response curve for either neurological or hepatic effects in humans.
However, the data indicate that inhalation exposure of workers to chloroform can
produce a range of neurological effects including fatigue, nausea, vomiting,  lassitude,
dry mouth, and anorexia.  Effects on the liver (jaundice, increased serum  enzyme
levels, hepatomegaly) were also observed in some studies (Phoon et al. 1983, Bomski
etal. 1967).

4.1.2.3 Inhalation Studies in Animals

Mery et al. (1994) exposed rats and mice to chloroform for 6 hours/day for 7
consecutive days. Target exposure concentrations ranged from 1 to 300  ppm.
Examination of the nasal passages revealed that chloroform caused  a complex set of
responses in the ethmoid turbinates, predominantly in rats.  These lesions were  most
severe peripherally and generally spared the tissue adjacent to the medial airways.  The
changes  were characterized by atrophy of Bowman's glands, new bone formation, and
increased labeling index in periosteal cells.  The only change noted in the mouse was
increased cell proliferation without osseous hyperplasia. The NOAEL values for these
responses ranged from 3 to 100 ppm, with histological and induced cell proliferation
being the most sensitive indices of effect.

Larson et al. (1996) exposed  male and female B6C3F1 mice to chloroform in air at
target concentrations of 0, 0.3, 2, 10,  30, or 90 ppm.  Exposure was for 6  hours/day, 7
days/week, for 4 days or for 3, 6, or 13 weeks.  Some additional animals were exposed
5 days/week for 13 weeks. All animals were examined for histological lesions of liver,
kidney, and nasal  epithelium. Some animals were administered bromodeoxyuridine
(BrdU) via osmotic pump prior to sacrifice in order to measure the labeling index (LI).
Chloroform caused treatment-related histopathological lesions in liver and nasal
passages of male and female mice and the kidneys of male mice.  Lesions in liver were
characterized by centrilobular hepatocyte swelling and lipid vacuolization  along with
scattered enlarged nuclei and individual cell necrosis.  The NOAEL for histological
effects in liver was 10 ppm, and the LOAEL was 30-90 ppm. Renal lesions  occurred in
male but not female  mice, primarily in the epithelial cells of the proximal convoluted
tubules in the cortex. Changes included mineralization, enlarged nuclei, and scattered
areas of regenerating foci.  The NOAEL for histological effects in kidney was 10  ppm,
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and the LOAEL was 30-90 ppm.  Nasal effects occurred in animals exposed for 4 days
at 10 ppm or higher. Nasal lesions were characterized by mild proliferative responses
of epithelial cells along with a thickening of bone.  Increases in LI were also observed at
10 ppm and higher. Both the histological and LI effects on nasal tissue were transient,
with little or no difference from controls after 13 weeks of exposure.

Templin et al. (1996a) exposed male and female F344 rats to chloroform in air at target
concentrations of 0, 2,  10, 30, 90, or 300 ppm. Exposure occurred for 6 hours/day,
either 5 or 7 days/week, for 4 days or 3, 6, or 13 weeks. Additional animals were
exposed for 5 days/week for 13 weeks.  All animals were examined for histological
lesions of liver, kidney, and nasal epithelium. Some animals were administered
bromodeoxyuridine (BrdU) via osmotic pump prior to sacrifice in order to measure the
LI.  Exposure to chloroform caused histopathological lesions in liver,  kidney, and nasal
epithelium of both male and female rats.  Lesions in liver were characterized by
scattered individual hepatocyte degeneration and necrosis, mitotic figures, and
midzonal vacuolization. The NOAEL for histological effects in liver was typically 30
ppm, and the LOAEL was typically 90 ppm.  Renal lesions occurred primarily in the
epithelial cells of the proximal convoluted tubules in the cortex.  Changes included
scattered vacuolation, individual tubule cell necrosis, and enlarged epithelial cell nuclei.
The NOAEL for histological effects in kidney was 30 to 90 ppm, and the LOAEL was 90
to 300 ppm. Increases in renal LI occurred at 30 ppm when exposure was 7 days/week,
and at 90 ppm when exposure was 5 days/week.  Nasal lesions were characterized by
atrophy of olfactory epithelium, mainly in the ethmoid portion of the nasal passage.
Effects were minimal at 2 ppm and increased in severity at higher exposure levels.
Increases in LI were seen in nasal epithelium at 10 ppm but not at 2 ppm. These
effects tended to be much larger after 4 days of exposure than after 3-13 weeks of
exposure.

Templin et al. (1998) exposed male BDF., mice to chloroform in air at target
concentrations of 0, 1, 5, 30 or 90 ppm. Female BDF! mice were exposed to 0, 5, 30 or
90 ppm. Animals were exposed for 6 hours/day, 5 days/week, for 3-13 weeks. All
animals were examined for histological lesions of liver and kidney.  Effects on nasal
epithelium were not investigated. Some animals were administered bromodeoxyuridine
(BrdU) via osmotic pump prior to sacrifice in  order to measure the LI.  Exposure to
chloroform caused histopathological lesions  in liver of male and female mice. Liver
lesions were characterized by centrilobular swelling along with centrilobular and
midzonal vacuolization and degeneration. The NOAEL for histological effects in liver
was 5 ppm, and the LOAEL was  30 ppm.  In males,  increases in hepatic LI were
observed at 90 ppm at 7 weeks but not at 13 weeks.  In females, an increase in LI was
observed at 90 ppm at 13 weeks. No effects on hepatic LI were seen at 30 ppm at any
time.  Renal lesions were observed in male but not female mice.  The predominant
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alteration was replacement of some or most of the epithelial cells of the proximal
convoluted tubules with regenerating cells characterized by basophilic cytoplasm and
variably sized heterochromatic nuclei.  The NOAEL for histological effects in kidney was
5 ppm, and the LOAEL was 30 ppm.  Increases in renal LI were observed at 30 ppm but
not 5 ppm.

Constan et al. (2002) exposed groups of female B6C3F1 mice (5 per group) to
chloroform vapors of 0, 10, 30, or 90 ppm for 2, 6, 12, or 18  hours/day for 7 consecutive
days.  Animals were observed for clinical signs during exposure and were subjected to
gross necropsy and histopathological examination at sacrifice. All animals were
administered bromodeoxyuridine (BrdU) via osmotic pump prior to sacrifice in order to
measure LI.  Groups of mice exposed to 90 ppm for 6 hours/day or to 30 ppm for 6 or
12 hours/day had minimal to mild hepatopathology characterized by centrilobular and
midzonal hepatocytic vacuolar degeneration. Individual necrotic cells were also
observable. Statistically significant increases in LI were observed in these three groups,
as well as animals exposed to 90 ppm for 2 hours/day.  No histological effects or
increases in LI occurred in any group exposed to 10 ppm, regardless of exposure
duration. The authors concluded that at an exposure level of 10 ppm,  the rate of
production of toxic metabolites does not exceed the  rate at which cellular detoxification
and repair reactions can occur.  Based on physiologically-based pharmacokinetic dose
modeling, the authors estimated  that the no-effect concentration for liver effects in
humans would be about 110 ppm.

4.1.2.4  Inhalation Reference Concentration

The EPA is currently working to develop an inhalation RfC for chloroform,  but a
consensus value has not yet been established.

4.2   CANCER EVALUATION

4.2.1  Oral Exposure

4.2.1.1  Studies in Humans

At present, there have been no studies of cancer incidence in humans chronically
exposed to chloroform  (alone). However, there have been a number of epidemiological
studies on cancer risk in humans who are exposed to chlorinated drinking water (e.g.,
Cantor et al. 1985, McGeehin et  al. 1993, King and Marrett 1996, Doyle et al. 1997,
Freedman et al.  1997, Cantor et  al. 1998, Hildesheim et al.  1998).  Chlorinated drinking
water typically contains chloroform, along with other trihalomethanes and a wide variety
of other disinfection byproducts (USEPA 1994).  It should be noted that humans

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exposed to chloroform in drinking water are likely to be exposed both by direct ingestion
and by inhalation of chloroform gas released from water into indoor air.

Some of these epidemiological studies have detected a weak association between
exposure to chlorinated water and cancer (mainly bladder cancer).  Based on the
studies of Cantor et al. (1985), McGeehin et al. (1993), King and Marrett (1996),
Freedman et al. (1997) and Cantor et al. (1998), EPA calculated that the population
attributable risk (the fraction of a disease which could be eliminated if the exposure of
concern were eliminated) for bladder cancer ranged from 2% to 17% (USEPA 1998f).
However, these calculations are based on a number of assumptions, including the
assumption that there is a cause-effect relationship between exposure to chlorinated
drinking water and increased risk of bladder cancer.  This  assumption  is subject to
considerable uncertainty, especially since findings are not consistent within or between
studies. Evaluation of these studies by application of standard criteria for establishing
causality from epidemiological observations (strength of association, consistency of
findings, specificity of association, temporal sequence, dose-response relation,
biological plausibility) has led EPA to conclude that the current data are insufficient to
establish a  causal relationship between exposure to chlorinated drinking water and
increased risk of cancer (USEPA 1998f).  Moreover,  even if, in the future, the weight of
evidence does become sufficient to establish  a causal link between exposure to
chlorinated water and increased risk of bladder or other types of cancer, it could not be
concluded from epidemiological studies of this type that chloroform per se is
carcinogenic in  humans, since chlorinated water contains  numerous disinfection
byproducts besides chloroform that are potentially carcinogenic (USEPA 1994).

4.2.1.2  Studies in Animals

There have been a number of studies in animals that demonstrate that ingestion of
chloroform  may increase the risk of cancer, at least by some types of exposure. The
most important of these studies are summarized below.

NCI (1976) evaluated the carcinogenic potential of chloroform in Osborne-Mendel rats.
Male rats were administered concentrations of 90 or 180 mg chloroform/kg-day in corn
oil, via oral  gavage, 5 days/week for 78 weeks. Female rats were administered
concentrations of  125 or 250 mg/kg-day for 22 weeks, after which the doses were
reduced to  90 or 180 mg/kg-day, with the average dose over the course of the study
being 100 or 200 mg/kg-day. Three additional groups of animals served as matched,
colony, and positive controls.  At week 111, all rats were sacrificed.  A statistically
significant increase in the incidence of kidney  epithelial tumors was observed in male
rats in the high-dose  group (12/50) when compared with males in  the control group
(0/98).  A statistically significant increase in the incidence of thyroid tumors was also
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observed in female rats, but this finding was not considered biologically significant
(USEPA1994).

NCI (1976) also evaluated the carcinogenic potential of chloroform using male and
female B6C3F1  mice. The average dose levels for the study were 138 or 277 mg/kg-
day for males and 238 or 477 mg/kg-day for females. All mice were sacrificed at weeks
92 or 93. Three additional groups of animals served as matched (20/sex/group), colony
(99 males and 98 females), and positive (100/sex/group) controls. The incidence of
hepatocellular carcinomas was significantly increased in males in  both the low-dose
(18/50) and high-dose (44/45) groups when compared to controls  (5/77). A similar
increase was observed for low-dose (36/46) and high-dose (39/41) females compared
to control (1/80) Many of the male mice in the low-dose group that did not develop
hepatocellular carcinoma had nodular hyperplasia of the liver.  The incidence of kidney
epithelial tumors was comparable between treatment and control groups.

Roe et al. (1979) reported three experiments in mice to evaluate the potential
carcinogenicity of chloroform. In three different studies, 10-week-old mice were
administered chloroform by gavage in a toothpaste base 6 days/week for 80 weeks,
followed by a 13- to 24-week observation period. The design of each study is
summarized  below:
Study
I
II
III
Strain (gender)
ICI (male, female)
ICI (male)
C57BL, CBA, CF/1 , ICI (male)
N
52/sex
52
52 per strain
Doses
(mg/kg-day)
0,17,60
0,60
0,60
In experiment I, kidney tumors were statistically higher in high-dose male mice than in
controls, while all other tumor incidences were comparable to the controls.  In
experiment II, a decrease in liver and kidney weights was observed in chloroform-
treated male mice, and the  incidence of kidney tumors was increased.  In experiment III,
treatment with chloroform was associated with increased incidence of moderate to
severe kidney lesions in CBA and CF/1  mice. No increases in liver or kidney tumors
were observed except in ICI male mice.

Heywood et al. (1979) exposed groups of 8 male and 8 female beagle dogs to doses of
15 or 30 mg chloroform/kg-day.  The chemical was given orally in a toothpaste base in
gelatin capsules, 6 days/week for 7.5 years. This was followed by a 20- to 24-week
recovery period. A group of 16 male and 16 female dogs received toothpaste base
without chloroform and served as the vehicle control group.  Eight dogs of each sex
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served as an untreated group and a final group of 16 dogs (8/sex) received an
alternative nonchloroform toothpaste. No treatment-related neoplastic lesions were
reported for the liver, gall bladder, cardiovascular system, reproductive system,  or
urinary system.

Jorgenson et al. (1985) exposed male Osborne-Mendel rats and female B6C3F1 mice
to chloroform in drinking water (0, 200, 400, 900, or 1,800 mg/L) for 104 weeks. Time-
weighted average doses, based on measured water intake and body weights, were 0,
19, 38, 81, or 160 mg/kg-day for rats and 0, 34, 65, 130, or 263 mg/kg-day for mice.  An
additional group of animals that served as controls was limited to the same water intake
as the high-dose groups. The number of animals  in the dose groups (from low to high)
was 330, 150, 50, and 50 for rats and 430, 150, 50, and 50 for mice. A statistically
significant dose-related increase in the incidence of kidney tumors (tubular cell
adenomas and adenocarcinomas) was observed in male rats in the high-dose group
(160 mg/kg-day). Chloroform in the drinking water did not increase the incidence of
hepatocellular carcinomas in female B6C3F1  mice. The combined incidence of hepato-
cellular adenomas and carcinomas was 2% in the high-dose group compared with 6%
in the control groups.  The authors speculated that the differences observed between
this study and the NCI (1976) bioassay may be related to differences in the mode of
administration (in drinking water versus in corn oil by gavage).

Tumasonis et al.  (1987) exposed male and female Wistar rats to chloroform  in drinking
water at concentrations of 0 or 2,900 mg/L for 72 weeks. Concentrations of chloroform
were then reduced to 1,450 mg/L for an additional 113 weeks until all animals had died
(approximately 185 weeks). The average dose for males and females was
approximately 200 and 150 mg/kg-day, respectively (USEPA 1994). Treated females
(but not males) showed a statistically significant increase in the incidence of hepatic
neoplastic nodules, and both males and females had a statistically significant increase
in the incidence of hepatic adenofibrosis. It is unclear if the nodules and adenofibroses
were considered  to be tumors (USEPA 1994).

DeAngelo (1995) exposed male F-344 rats to chloroform in drinking water for 100
weeks. Exposure levels were 0,  900, or 1,800 ppm.  Assuming ingestion of  about 0.05
L/day of water  per kg  body  weight, this corresponds to doses of approximately 45 and
90 mg/kg-day,  respectively.  Exposure began when the animals were 8 to 10 weeks of
age. Interim sacrifices of groups of 6 animals were performed at 26, 52, and 78 weeks,
and the final sacrifice at 100 weeks included 50 animals per group. At each time point,
liver and  kidney were examined for gross and microscopic lesions.  In the  liver, there
were borderline significant (p = 0.05 to 0.10) increases in the prevalence of
hepatocellular proliferative lesions at 100 weeks.  In addition, there was a  statistically
significant increase (p < 0.05) in the multiplicity of hepatic adenomas and carcinomas in
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the group exposed to 1,800 ppm, and a significant dose trend (p < 0.05) for hyperplastic
nodules, neoplasia, and total proliferative lesions.  No renal neoplasms were observed
in any of the chloroform-exposed groups.

4.2.2  Inhalation Exposure

4.2.2.1 Studies in Humans

No studies were located in humans on inhalation exposure to chloroform and increased
risk of cancer.

4.2.2.2 Studies in Animals

Nagano et al. (1998) evaluated the effects of chronic exposure of F344 rats and BDF1
mice to chloroform vapor.  This study has also been summarized in abstract form by
Yamamoto et al. (1994, and described in a letter report by Matsushima (1994). Groups
of  male and female rats and mice were exposed to target chloroform vapor
concentrations of 0, 10, 30, or 90 ppm (rats) or 0,  5, 30, or 90 ppm (mice), 6 hours/day,
5 days/week for 104 weeks.  To avoid lethality in the high-dose groups, mice in the 30-
ppm group were exposed to chloroform concentrations of 5 ppm for the first 2 weeks, 10
ppm for the next 2 weeks,  and then exposed to 30 ppm for the remaining period.  For
the 90 ppm group, rats were exposed to chloroform initially at 10 ppm for 2 weeks, then
30 ppm for 2 weeks, and then 90 ppm for 98 weeks. The time-weighted average for the
30-ppm group was 29.1 ppm and for the 90-ppm group 85.7 ppm. Statistically
significant increases in the incidence of renal cell adenoma and renal cell carcinoma
(combined) were observed in male mice in the 30 (7/50) and 90 (12/48) ppm groups
when compared with controls (0/50).  The overall incidence rates of renal cell carcinoma
were statistically significantly increased in males in the 90-ppm group (11/48) when
compared with controls (0/50).  There were no statistically significant changes in tumor
incidence for female mice or for male or female rats in any exposure group.

4.2.3  Mutagenicity

A large number of studies  have been performed to evaluate the mutagenicity of
chloroform. These data have been reviewed and evaluated by several groups of
experts, including the International Commission for Protection against Environmental
Mutagens and  Carcinogens (ICPEMC), ILSI (1997), and WHO (1998).

ICPEMC applied a comprehensive, quantitative weight-of-evidence approach to
characterize the genotoxic potential of more than 100 chemicals with large genetic
toxicity databases (Lohman et al. 1992). In this approach, scores are developed for

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relative DMA reactivity.  For a particular chemical, the maximum possible score is 100
and the minimum possible score is -100.  The highest actual score obtained using this
approach was 49.7 (triazaquone) and the lowest score was -27.7 (ethanol).  When this
approach was applied to chloroform, the score based on the results of more than 40
studies was -14.33. Thus, ICPEMC concluded that the weight of evidence indicates
that chloroform should be classified as nongenotoxic (Brusick et al. 1992, Lohman et  al.
1992).

ILSI (1997) performed a review of the available data on the mutagenicity of chloroform.
The committee noted that phosgene is highly reactive and might be expected to have
the capacity to interact directly with DMA,  but that phosgene has not been tested in any
standard mutagenicity test system. The committee also noted that, because of its high
reactivity, phosgene formed in the cytosol following chloroform metabolism would likely
react with cellular components prior to reaching the cell nucleus, and concluded that
direct effects on DMA would be unlikely.  Based on their review of the available data,  the
ILSI committee concluded that no subset of observations points unequivocally to a
specific genotoxic mode of action associated with chloroform,  and  that the
preponderance of the evidence indicates that chloroform is not strongly mutagenic (ILSI
1997).  Based on this, the committee concluded that chloroform would not be expected
to produce rodent tumors via a genotoxic mechanism.

WHO (1998) noted that studies on the mutagenicity of chloroform must be considered in
light of the fact that (1) chloroform is volatile, so tests that do not prevent volatilization
are unreliable, and  (2) most chloroform contains ethanol, which may react with
phosgene generated from chloroform metabolism to yield ethyl or diethyl carbamates
(potentially causing false positive results).  The WHO committee noted that largely
negative results have been obtained in Salmonella typhimurium and Escherichia coli
(with and without activation), in gene mutation tests in Chinese hamster ovary (CHO)
cells and human lymphocytes, in mouse micronucleus tests, and in tests of
unscheduled DMA synthesis both in vitro and in vivo. Given the large number of
sensitive assays that have been used to investigate the genotoxicity of chloroform, the
committee considered it noteworthy that the positive responses were so few, and that
the positive results were randomly distributed among the various assays. Taken
together, WHO (1998) concluded that the weight of evidence indicates that neither
chloroform nor its metabolites appear to interact directly with DMA or possess genotoxic
activity.

EPA considered all of the available data and noted that the results of the mutagenicity
assays that have been conducted with chloroform are mixed (USEPA 2001 a).  By
number, the majority of tests are negative, and many of the positive studies have been
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conducted under high exposure conditions that resulted in severe cytotoxicity.  As
expressed by EPA's Science Advisory Board (SAB 2000):

      Genotoxicity endpoints have to be interpreted cautiously when used as evidence
      for potential carcinogenicity. In vitro clastogenicity can be a product of severe
      cytotoxicity resulting from lysosomal or other releases (Brusick 1986). This may
      be important with substances such as chloroform, where there is evidence of
      cytotoxicity and cell proliferation in target tissues. Also, cycles of cytotoxicity and
      cell proliferation could cause the expression of preexisting genetic damage in
      target tissues which, under normal conditions, have low mitotic indices.

Consequently, EPA concluded that the relevance of many of the positive studies is
questionable. Therefore, based on the preponderance of negative findings and the
uncertain relevance of the positive findings, EPA concluded that the weight of evidence
indicates that even though a role for mutagenicity cannot be excluded with certainty,
chloroform is not a strong mutagen and that neither chloroform nor its metabolites
readily bind to DMA (USEPA 2001 a). Based on these results and the results of studies
that evaluated other endpoints of DMA reactivity, it seems  likely that chloroform does
not produce carcinogenic effects primarily by a specific mutagenic mode of action.

4.2.4  Cancer Evaluation Using Current Guidelines

In accord with proposed EPA guidelines  for cancer risk assessment (USEPA 1996a),
the method used to characterize and quantify cancer risk from a chemical depends on
what is known about the mode of action  of carcinogenicity and the shape of the cancer
dose-response curve for that chemical. A default assumption of linearity is appropriate
when evidence supports a mode of action of gene mutation due to DMA reactivity or
supports another mode of action that is anticipated to be linear. The linear approach is
used as a matter of policy if the mode of action of carcinogenicity is not understood. A
default assumption of non-linearity is appropriate when there is no evidence for linearity
and sufficient evidence to support an assumption of nonlinearity.  Alternatively, the
mode of action may theoretically have a  threshold, e.g., the carcinogenicity may be a
secondary effect of toxicity that is  itself a threshold phenomenon (USEPA 1996a).

USEPA  (2001 a) has completed an evaluation of the carcinogenic hazard of chloroform,
using EPA's new cancer assessment guidelines.  The following sections summarize this
evaluation, taken from USEPA (2001 a) with little modification.
                                      24

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                                REVISED DRAFT

4.2.4.1 Mode of Action

Noncancer Mode of Action

The exact mode of action by which chloroform produces toxic effects in liver and kidney
is not yet certain, but it is evident that metabolism of chloroform to toxic metabolites by
cytochrome P450-dependent pathways plays a critical role (USEPA 1994, USEPA
2001 a). For example,  Brown et al. (1974) reported that pretreatment of rats with
phenobarbital (a cytochrome  P-450 inducer) resulted in increased hepatic toxicity
following chloroform exposure.  Similarly, Gopinath and Ford (1975) indicated that
chloroform hepatotoxicity in rats was increased by phenobarbitone, phenylbutazone,
and chlorpromazine, all inducers of microsomal enzymes.  Conversely, inhibitors of
microsomal enzymes, such as SKF-525A, sodium diethyl-dithiocarbamate, and carbon
disulfide, decreased the hepatic toxicity of chloroform. Constan et al. (1999) showed
that 1-aminobenztriazole, which is a general cytochrome P450 inhibitor, prevented
chloroform-induced toxicity in liver and kidney of mice following inhalation exposure.

Further evidence of the role of metabolism  is derived from the finding that variations in
toxicity between tissues,  genders, and species generally correlate with differences in
cytochrome P450 metabolic rate.  For example, male mice are more sensitive to
chloroform-induced renal toxicity than female mice, and this difference in toxicity is
paralleled in a difference in metabolism in proximal tubular cells (llett et al. 1973).
Renal cytochrome levels in mice are increased by testosterone (Mohla et al. 1988,
Henderson et al. 1989, Hong  et al.  1989), and female mice treated with testosterone
have increased  renal toxicity along with increased covalent binding of chloroform
metabolites (Taylor et al. 1974, Smith et al. 1979, Pohl et al. 1984).  Conversely, male
mice that were castrated had  lower levels of chloroform-derived radioactivity
accumulated in  the kidneys (Eschenbrenner and Miller 1945, Culliford and Hewitt 1957,
Taylor et al. 1974, Smith et al.,  1984).  Constan et al. (1999) compared the toxicity of
chloroform in three strains of  mice: B6C3F1, Sv/129 wild type, and Sv/129 CYP2E1
knockout mice.  Exposure to 90 ppm chloroform for 6 hours/day for 4 days produced
clear hepatotoxicity and renal toxicity (histopathology, increased labeling index) in the
B6C3F1 mice and the Sv/129 wild type, but not in the Sv/129 CYP2E1 knockout mice.
The authors concluded that metabolism of chloroform by CYP2E1 was obligatory for
toxicity, at least  at the dose tested.

Cancer Mode of Action

As noted above, studies in animals reveal that chloroform can cause an increased
incidence of kidney tumors in male rats and an increased incidence of liver tumors in
male and female mice. However,  not all exposure regimens have been observed to
                                      25

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                                REVISED DRAFT

produce increases in cancer incidence.  In particular, increase in tumors appear to
occur only in those cases where the exposure regimen resulted in cytotoxicity, and did
not occur under conditions when cytotoxcity did not occur (Butterworth and Bogdanffy
1999). This observation lead to the hypothesis that chloroform-induced tumor
responses are secondary to cytotoxicity, most likely as a consequence of the
regenerative hyperplasia which occurs following cytotoxicity in liver and kidney.

Regenerative hyperplasia can lead to an increased probability of cancer by one or both
of two alternative modes of action.  First, cells that are undergoing cell division are
inherently more susceptible to initiation than are slowly growing or nondividing cells.
This is because DMA undergoing replication is more exposed to nucleophilic attack than
DMA that is covered with histones and arranged in nucleosomes (Ames and Gold,
1991a,b). Also, any gene damage that occurs in a cell undergoing division has less
time to be repaired before mitosis than in a slowly growing cell, so a larger fraction of
DMA alterations could be converted into mutations. Second, chemicals that promote
cell division may convey a selective growth advantage to preexisting initiated cells in
comparison with normal cells,  thereby facilitating clonal expansion of initiated  cells.  This
could occur because initiated cells are more responsive than normal cells to growth
stimuli, because they are less  susceptible to the toxicity of the chemical, or because
they are less susceptible to endogenous regulatory signals that trigger programmed  cell
death (apoptosis). In any case, the ratio of cell birth to cell death of initiated cells
increases compared with normal cells, leading to increased  likelihood that a clone of
initiated cells will form and survive.  A key characteristic of this mode of action is that the
effect is reversible: the clones of induced cells will tend to regress if the promoter
(mitogen, cytotoxicant)  is withdrawn (Pitot et al.  1987, Schulte-Hermann et al. 1993).

EPA reviewed the strength and consistency of the association between cytotoxicity and
regenerative hyperplasia and the occurrence of increased tumor frequency in  exposed
animals (USEPA 2001 a). This analysis revealed two main points:

      There are numerous cases where exposure to chloroform causes an increase in
      cellular regeneration (as reflected in  an increase in LI) without any observable
      increase in  cancer incidence. These data indicate that chloroform exposures  that
      are adequate to cause cytotoxicity and regenerative cell proliferation do not
      always lead to cancer.

      There are no cases in which a tumorigenic response  has been observed where
      evidence of cell regeneration is not also observed at the same or lower dose as
      that which caused an increase in tumors.  This consistency of evidence (i.e., cell
      regeneration is detected in all cases of tumorigenicity) is strong evidence
                                       26

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                                REVISED DRAFT

      supporting the conclusion that cell regeneration is a mandatory precursor for tumorigenicity.

In male Osborne-Mendel rats exposed to chloroform in water for 2 years (Jorgenson et
al. 1985), a statistically significant increase in renal tumors was observed at a
concentration of 1,800 ppm (160 mg/kg-day). A reanalysis of the histopathological
slides from this study (Hard et al. 2000) revealed evidence for sustained cytotoxicity and
cell proliferation in the kidney at exposures of 900 ppm (81 mg/kg-day) or higher.
Likewise, in BDF1 mice exposed to chloroform by inhalation at 5, 30, or 90 ppm for 6
hours/day, 5 days/week (Nagano et al. 1998), increased incidence of renal tumors was
observed in male mice at the two higher doses while females showed no significant
tumor response. Templin et al. (1998) duplicated this exposure regimen to study
whether the treatment caused cytotoxicity and regenerative hyperplasia. These authors
observed cytotoxicity and hyperplasia in the kidneys of male mice exposed to 30 or 90
ppm throughout a 90-day exposure period, but not in exposed females.  This
observation is consistent with the hypothesis that sustained cytotoxicity and
regenerative hyperplasia are key events in the neoplastic response of the kidney to
chloroform.

Available data also indicate that cytotoxicity and regenerative hyperplasia are required
for liver cancer, although the strength of this  conclusion is somewhat limited because
most of the observations are based on short-term rather than long-term histological or LI
measurements.  For example, in B6C3F1 mice, exposure by corn oil gavage at the
same doses that resulted in liver tumors in the study by NCI (1976) also caused hepatic
cytolethality and a cell proliferative response at 4 days and 3 weeks (Larson et al.
1994a,b).  Similarly, exposure of female B6C3F1  mice to chloroform in drinking water at
levels that did not induce liver tumors (Jorgenson et al. 1985) also did not induce
hepatic cytolethality or cell proliferation at 4 days or 3 weeks (Larson et al. 1994a).  This
consistency of the data (i.e., evidence of cytolethality and/or regenerative hyperplasia is
always observed in cases of increased liver tumors) supports the conclusion that this
liver cancer also occurs via a mode of action involving regenerative hyperplasia.

In summary, based on a review of all available data, EPA concluded that the weight of
the evidence in animals supports the conclusion that chloroform-induced tumors in liver
and kidney are only produced at dose levels  that result in repeated or sustained
cytotoxicity and regenerative cell proliferation, and that direct genotoxicity is unlikely to
play an important role (USEPA 2001 a).  No data  exist that indicate the mode of action
observed in rodents does not also apply to humans.
                                       27

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                                REVISED DRAFT

4.2.4.2  Quantification of Cancer Risk from Oral Exposure

The Proposed Guidelines for Carcinogenic Risk Assessment (USEPA 1996a) state that
when the mode of action analysis based on available data indicates that "the
carcinogenic response is secondary to another toxicity that has a threshold, the margin-
of-exposure analysis performed for toxicity is the same as is done for a  non-cancer
endpoint, and an RfD for that toxicity may be considered in the cancer assessment".

This is the case for chloroform. That is, available evidence indicates that chloroform-
induced carcinogenicity is secondary to cytotoxicity and regenerative  hyperplasia, and
that doses below the RfD do not result in cytolethality (and hence do not result in
increased risk of cancer). Accordingly, the RfD developed above (1E-02 mg/kg-day) for
protection against non-cancer effects (including cytolethality and regenerative
hyperplasia) is also judged to be protective against increased risk of cancer.

4.2.4.3  Quantification of Cancer Risk from Inhalation Exposure

Only one study was located that is a candidate for quantification of cancer risks
following inhalation exposure to chloroform (Nagano et al. 1998, also  reported in
Yamamoto  et al. 1994 and Matsushima 1994). However, this study is not considered
suitable for quantitative cancer dose-response modeling because the study had to rely
on an increasing step-wise exposure protocol for male mice  in the two highest exposure
groups (30  and 90 ppm) in order to avoid lethality. This indicates that the 30 and 90
ppm doses  exceed the maximum-tolerated dose (MTD), making it difficult to interpret
these observations with respect to potential cancer effects at lower doses.  Moreover,
this study has not been published in full and the data are not available for detailed
review.

In the absence of reliable quantitative inhalation cancer dose-response  data for
chloroform, risk of cancer from inhalation exposure was evaluated using a non-linear
RfC-type approach,  similar to that described above for oral exposure. Although data on
the mode of chloroform action following inhalation  exposure  are less robust than for oral
exposure, the available data suggest the mode of action is likely to be the same by both
exposure routes. That is, both oral exposure and inhalation  exposures  of animals cause
effects in the same target tissues (liver, kidney, nasal epithelium),  and the histological
and biochemical nature of the adverse effects observed in these tissues are the same in
both cases.  Further, in analogy with what is well-established for oral exposure, Templin
et al. (1998) demonstrated that the inhalation exposure levels (30 and 90 ppm) that had
been shown to produce renal tumors in male mice (Nagano et al. 1998) also produced
evidence of renal cytotoxicity and cellular regeneration, while no evidence of renal
cytotoxicity  or cellular regeneration was noted in groups that did not display a
                                       28

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                                REVISED DRAFT

tumorigenic response (males at 5 ppm, females at all exposure levels).  Based on this,
EPA concluded that inhalation exposures that do not cause cytolethality and
regenerative hyperplasia in liver or kidney will not cause increased risk of cancer
(USEPA 2001 a). That is,  inhalation exposures below the RfC of 0.1  mg/m3
(corresponding to an inhalation RfD of 0.03 mg/kg-day) do not result in cytolethality, and
hence do not result in increased risk of cancer.

4.2.5  Discussion of Confidence

Available evidence is strong that chloroform can induce increased risk of cancer in
animals. A large number of studies have been performed to investigate the hypothesis
that this increased risk of cancer requires occurrence of cytotoxicty and regenerative
cell proliferation.  To date, the strength of this association is strong: there are no known
cases of increased cancer incidence in animals that were not accompanied by clear
evidence of cytotoxicity and regenerative hyperplasia in the target organ. While
compelling, it is important to realize that a correlation of this type can never prove the
proposed mode of action is correct, since the hypothesis could be disproved by only a
single counter example.  In addition, it is important to recall that even though the weight
of evidence indicates that chloroform is not strongly mutagenic, there are a number of
mutagenicity studies that yielded positive results, and it remains plausible that
chloroform may have a weak potential to cause mutagenic effects under at least some
dose conditions.  On this basis, confidence in the cancer assessment is rated as
medium. This confidence level may increase as further studies are performed on the
mode of action.
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   30

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                      5. EXPOSURE ASSUMPTIONS

5.1    RELATIVE SOURCE CONTRIBUTION ANALYSIS

When an ambient water quality criterion is based on noncarcinogenic effects (as is the
case for chloroform), anticipated exposures from non-occupational sources (e.g., food,
air) are taken into account.  As described in the 2000 Human Health Methodology
(USEPA 2000b),  EPA has, in the past, used a "subtraction" method to account for
multiple sources of exposure to pollutants.  However, EPA has also previously used a
"percentage" method for the same purpose.  In this approach, the percentage of total
exposure typically accounted for by the exposure source for which the criterion is being
determined, referred to as the relative source contribution (RSC), is applied to the RfD
to determine the maximum amount of the RfD "apportioned" to that source. The
underlying objective is to maintain total exposure below the RfD while generally avoiding
an extremely low limit in a single medium that represents just a nominal fraction of the
total exposure. To meet this objective, all proposed numeric limits lie between 80% and
20% of the RfD.  EPA uses the Exposure Decision Tree in the 2000 Human Health
Methodology (Figure 4-1 in USEPA 2000b) when deriving its AWQC values.

When more than  one criterion is relevant to a particular chemical, as in the case of
chloroform, apportioning the RfD via the percentage method of the Exposure Decision
Tree is considered appropriate to ensure that the combination of criteria, and thus the
potential for resulting exposures, do not exceed  the RfD. In following the Exposure
Decision Tree,  it is necessary that adequate data exist for the relevant
sources/pathways of exposure if one is to avoid  using default procedures.  The
adequacy of data is a professional judgement for each individual chemical of concern,
but EPA recommends that the minimum acceptable data are exposure distributions that
can be used to determine, with an acceptable 95%  confidence interval, the central
tendency and high-end exposure levels for each source. If there are not sufficient
data/information to characterize exposure, particularly to the media of concern, EPA
intends to use the "default" assumption of 20% of the RfD when deriving or revising the
RfD.

5.1.1  Population of Concern

For chloroform, the population of concern for setting national criteria is assumed to be
the adults of the general population of the United States. Based on a review of CYP2E1
levels in the fetus and in children, EPA concluded that there is no basis to conclude that
the mode of action of chloroform would differ between children and adults, and that
                                      31

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                               REVISED DRAFT

neither the fetus nor the child appears to more sensitive than adults based on level of
CYP2E1 activity (USEPA 2001 a).

5.1.2  Overview of Potential for Exposure

Humans may be exposed to chloroform from a number of different media and by
several different exposure routes. Figure 5-1 is a conceptual model that summarizes
these exposure sources and routes.

5.1.2.1 Ambient Water

The principal pathways by which humans may be exposed to chloroform in ambient
water include ingestion of the water as drinking water and ingestion offish and other
aquatic organisms collected from the water.

Exposure to chloroform  in ambient water may also occur by incidental ingestion during
recreational activities, but this pathway is generally so minor that it is not included in the
derivation of the AWQC (USEPA 2000b). Likewise, dermal contact during recreational
exposures to ambient water are likely to be so minor that they are not included.

5.1.2.2 Treated Water

When water is treated with a strong oxidant such as chlorine, chlorine dioxide, or
chloramine to kill potentially pathogenic organisms that may be present in the water,
chlorine is formed as a disinfection byproduct (DBP) by oxidation of organic matter that
is present in the water.  Residents in homes supplied with treated drinking water may be
exposed to chloroform by three main routes:  ingestion, inhalation of chloroform vapors
that have been released from the water into indoor air, and dermal contact with the
water while showering or bathing.

Exposure to chloroform  may also occur during swimming in a pool that is disinfected
with chlorine.  Exposure during swimming may occur  by ingestion, inhalation, and
dermal pathways.  However, the frequency of such exposures is relatively small for the
general population, and estimated exposures via these pathways are small compared to
other pathways (USEPA 2000b). Therefore, exposures during swimming are not
included in the derivation of the AWQC.
                                      32

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                         REVISED DRAFT
   FIGURE 5-1  EXPOSURE PATHWAYS FOR CHLOROFORM
 Ambient Water
                           Fish
                       Drinking Water
                         (untreated)
      Ingestion
     Ingestion as
    drinking water
Swimming Pools
Swimming
pool air
^
w
Inhalation
 Dermal exposure while
	swimming	
                                              Incidental ingestion while
                                              	swimming	
 Treated Water
Drinking water

Indoor air
^
^

Ingestion

Inhalation
                         Showering
                                                   Inhalation
                                                    Dermal
\ \
Other Sources \ \
\ \

\ <<
V
\b
T
\
+
^
^
w
^

1
Diet
(non-fish)

Beverages
(non tap water)

Indoor Air

Outdoor Air
hi
w
^
^
^


Ingestion

Ingestion

Inhalation

Inhalation
               Minor pathway, not quantified in derivation of RSC or AWQC
                                33

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                               REVISED DRAFT

5.1.2.3 Non-Water Sources

Chloroform may also be found in various dietary items, and this can contribute to oral
exposure of humans.  It should be noted that some of the chloroform present in dietary
items may be attributable to chloroform that was present in treated water used in the
preparation of the food as well as from the use of chlorine disinfectants in the
processing of foods or the cleaning of food contact surfaces. Chloroform may also be
released to indoor and outdoor air from a variety of industrial sources and from
consumer products, including household bleach, that contain chloroform, and this can
lead to inhalation exposure of humans.

5.1.3  Estimates of Exposure from Non-Ambient Water Sources

EPA has published an extensive review of human exposure to chloroform from a
variety of environmental media (USEPA 2001 b).  Information presented below on
concentration levels and exposures from various non-ambient water related pathways
are taken from that report. It should be noted that all exposure estimates for these
exposure pathways are based on absorbed doses and are intended to estimate median
exposures for the general population.

5.1.3.1 Exposure from Treated Drinking Water

Ingestlon Exposure

The basic equation for calculating the average daily absorbed dose (ADD) of chloroform
from ingestion of treated water as drinking water is as follows:

            ADD(ingestion of treated water) = CTW  x IRTW*AF0/BW

Input values and the resultant ADD value are summarized below:
                                      34

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                                 REVISED DRAFT
Parameter
CTW
IRT]A/BW
AF0
ADD
Description
Mean concentration of chloroform in
treated drinking water
Average ingestion rate of treated
water
Oral absorption fraction for
chloroform
Average daily absorbed dose from
ingestion of treated drinking water
Value
24
0.019
1.0
0.46
Units
ug/L
L/kg-day
~
ug/kg-
day
Source
USEPA2001b
USEPA2001b
USEPA2001b
Calculated from
inputs
General Inhalation Exposure

The basic equation for calculating the average daily absorbed dose of chloroform from
inhalation of chloroform that is released from treated water into general indoor air  is as
follows:

    ADD(inhalation of vapors from treated water) = C(air)general x BR x ET x /\f./B\N

Data on the concentration of chloroform in general indoor air indicate that the average
value is about 3.0 ug/m3 (USEPA 2001 b). This measurement includes chloroform from
all sources and does not distinguish the amount of chloroform due to releases from
water from that which is due to releases from other sources. However, a screening
level calculation of the fraction due to water is possible, as follows.

      The release of radon from water to indoor air has been well-studied, and the average
      transfer factor for radon from water to indoor air is about 0.1 L/m3 (USEPA 2001c).
      Based on physical-chemical properties, the release of chloroform from water into air is
      expected to be about 75% that of radon (McKone 1987).  Thus,  the predicted average
      increment in general indoor air due to releases from water is as follows:

                   C(air)general = 24 pg/L x 0.075 Urn3 = 1.8 pg/m3

      Adding the average concentration in outdoor air (about 1.6 ug/m3) (USEPA 2001 b) to the
      water-related increment (1.8 ug/m3) yields a predicted total indoor air mean
      concentration of about 3.4 ug/m3, a value that is slightly higher than the measured mean
      value (3.0 ug/m3) (USEPA 2001 b). This calculation indicates  that measured levels of
      chloroform in indoor air are likely due mainly to releases from treated water plus the
      contribution from outdoor air, and releases from other (non-water) indoor sources  are
      likely to be so small that they can be ignored.

Based on this, input values and the resultant ADD value are summarized below:
                                        35

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                               REVISED DRAFT
Parameter
C(air) general
BR
ET
AFi
BW
ADD
Description
Mean concentration of chloroform in
general indoor air
Average breathing rate while indoors
Average exposure time indoors
Inhalation absorption fraction
Mean body weight
Average daily absorbed dose from
breathing chloroform in indoor air
Value
3.0
0.67
24
0.63
70
0.43
Units
ug/m3
m3/hour
hours/day
~
kg
ug/kg-day
Source
USEPA2001b
USEPA2001b
USEPA2001b
USEPA2001b
USEPA 2000b
Calculated
from inputs
Inhalation Exposure While Showering

The basic equation for calculating the average daily absorbed dose of chloroform from
inhalation of chloroform  in bathroom air that is released from treated water during
showering is as follows:

     ADD(inhalation of vapors while showering) = C(air)shower x BR x ET x /\f./B\N

Input values and the resultant ADD value are summarized below:
Parameter
C(air)shower
BR
ET
AFt
BW
ADD
Description
Mean concentration of chloroform in air
during showering
Average breathing rate while showering
Average exposure time during and after
showering
Inhalation absorption fraction
Mean body weight
Average daily absorbed dose from
breathing chloroform in shower air
Value
190
0.66
0.12
0.63
70
0.14
Units
ug/m3
m3/hour
hours/day
~
kg
ug/kg-day
Source
USEPA 2001 b
USEPA 2001 b
USEPA 2001 b
USEPA 2001 b
USEPA 2000b
Calculated
from inputs
Dermal Exposure While Showering

The basic equation for calculating the average daily absorbed dose of chloroform from
dermal contact with treated water used for showering or bathing is as follows:
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                               REVISED DRAFT

           ADD(dermal contact with treated water) = CTW xD/Ax ED x SAR

Input values and the resultant ADD value are summarized below:
Parameter
CTW
DA
ED
SAR
ADD
Description
Mean concentration of
chloroform in treated water
Dermal absorption rate from
water
Showering exposure duration
Surface area to body weight
ratio
Average daily absorbed dose
from dermal contact with treated
water
Value
24
3.52E-06
5
290
0.12
Units
ug/L
ug per ug/L
per cm2-min
m in/day
cm2/kg
ug/kg-d
Source
USEPA2001b
USEPA2001b
USEPA2001b
USEPA2001b
Calculated
from inputs
5.1.3.2 Exposures from Outdoor Air

The basic equation for calculating the average daily absorbed dose of chloroform from
inhalation of chloroform in outdoor air is as follows:

          ADD(inhalation in outdoor air) = C(air)outdoor x BR x ET x AF,/BW

Concentration values of chloroform in outdoor air vary substantially depending on the
location of measurement. Based on data tabulated in USEPA (2001 b), the mean
concentration for locations in the United States is about 1.6 ug/m3.  Assuming an
average outdoor exposure of 2 hours/day (USEPA 1997b), the resultant ADD value is
as summarized below:
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                               REVISED DRAFT
Parameter
C(air)outdoor
BR
ET
AFi
BW
ADD
Description
Mean concentration of chloroform in
outdoor air (United States)
Average breathing rate
Average exposure time outdoors
Inhalation absorption fraction
Mean body weight
Average daily absorbed dose from
breathing chloroform in indoor air
Value
1.6
0.67
2
0.63
70
0.019
Units
ug/m3
m3/hour
hours/day
~
kg
ug/kg-day
Source
USEPA2001b
USEPA2001b
USEPA1997b
USEPA2001b
USEPA 2000b
Calculated
from inputs
5.1.3.3 Dietary Exposures

USEPA (2001 b) summarized data on the concentration levels and estimated exposure
levels to chloroform in a wide variety of dietary items. These data are summarized in
Table 5-1. As shown, the estimated absorbed dose of chloroform from all dietary
sources (excluding freshwater and estuarine fish) is about 0.46 ug/kg-day, with dairy
products and grain contributing the largest intakes.

5.1.3.4 Total Non-Ambient Water Exposures

The total exposure from all  non-ambient water sources is summarized in Table 5-2.  As
seen, the estimated total average daily absorbed dose is about 1.6 ug/kg-day if
exposures from treated drinking water are included, and about 0.5 ug/kg-day if
exposures from treated water are excluded.

5.1.4 Estimates of Exposure from Ambient Water Sources

5.1.4.1 Ingestion of Ambient Water

USEPA (2001 b) summarizes data from a number of studies on chloroform levels in
ambient surface water bodies (mainly streams and rivers).  Average levels typically
ranged from 0.2 to 5 ug/L.  However, nearly all of these data (15 out of 16 observations)
are from locations outside of the  United States (mainly Finland). The only study with
data from within the US  (Sheldon and Hites 1978) reported that chloroform levels in
water samples from the  Delaware River were below the detection limit.  However, the
detection limit was not reported.  Thus, it is concluded that data on chloroform  levels in
ambient waters of the US are too limited to allow reliable quantitation of exposure
through ingestion.
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      TABLE 5-1. DIETARY EXPOSURES TO CHLOROFORM

Category
Fruits/Vegetables



Meat/Dairy





Grain
Marine Fish
TOTAL

Type
Fruits
Exposed veg
Protected veg
Root veg.
Beef
Pork
Lamb
Sausage
Poultry
Dairy
All
All

C
ug/g
0.010
0.025
0.019
0.016
0.090
0.038
0.060
0.017
0.038
0.079
0.045
0.052

IR
g/kg-day
1.6
1.2
0
0.78
0.54
0.07
0.0056
0.12
0.30
2.60
2.6
0.154

ADD
ug/kg-day
0.016
0.030
0.000
0.012
0.049
0.003
0.0003
0.002
0.011
0.205
0.117
0.008
0.454
All data are from USEPA (2001 b)
                          39

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TABLE 5-2. SUMMARY OF NON-AMBIENT WATER EXPOSURES
Source
Treated Water
Ambient Air
Diet
Total (Including treated water)
Total (Excluding treated water)
Pathway
Ingestion
General Inhal
Shower Inhal
Shower Dermal
Outdoor Inhal
Fruits
Veg
Meat
Dairy
Grain
Marine fish

ADD
ug/kg-day
0.46
0.43
0.14
0.12
0.02
0.02
0.04
0.07
0.21
0.12
0.01
1.62
0.47
                       40

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                               REVISED DRAFT

5.1.4.2 Ingestion of Freshwater and Estuarine Fish

No data were located on the concentration of chloroform in freshwater or estuarine fish
or other aquatic species. Thus, data are not sufficient to quantitate exposure from
ingestion of fish from ambient water sources.

5.2    EXPOSURE DATA ADEQUACY AND UNCERTAINTY ESTIMATES

As noted above, the main source of human exposure to chloroform is through the use of
treated (disinfected) water for drinking and other indoor uses.  Data on the level of
chloroform in treated water are extensive, and all public drinking water systems are
subject to national regulations that limit the levels of chloroform and other
trihalomethanes that may be present in treated water.  Human exposure to chloroform
in treated water is highly variable, depending both on human behavioral factors (amount
of water ingested, time spent in the shower, time spent at home, etc.), and on house
construction variables (size of the home, ventilation rate, water use rate, etc.).
However, reliable data are available to characterize all of these exposure variables, and
hence estimates of human exposure to chloroform from treated water are reasonably
certain.

Data on chloroform levels in the diet are less extensive than for water, but are sufficient
to establish that exposure through the diet is relatively small compared to the sum of all
exposure pathways (ingestion, inhalation, dermal) related to use of treated water.

Data on chloroform levels in ambient water are very limited, and no data were located
on chloroform levels in fish from freshwater or estuarine sources.  Thus, data are
inadequate to allow reliable quantitation of typical exposure levels from ambient water
sources.

5.3    RSC ESTIMATES/ALLOCATION OF THE RfD

RSC for Exposure from Ingestion of Fish Only

In the case of an individual who is exposed to ambient water only  by ingestion of
freshwater/estuarine fish, the equation for calculating the RSC is as follows:

                     RSC = ADD(freshwater fish) /ADD(total)

As discussed above, when total exposure includes exposures to treated water, the total
ADD is about 1.6 ug/kg-day (see Table 5-2).  The precise fraction of this total dose that
is contributed by ingestion of freshwater or estuarine fish is unknown because no data
                                      41

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                               REVISED DRAFT

could be located on chloroform levels in freshwater fish.  In accord with the Exposure
Decision Tree approach described in Figure 4-1 of USEPA (2000b), when data are not
available to derive reliable quantitative estimates of all relevant exposure distributions, a
default "floor" value of 20% is recommended for the RSC (Box 8B). Based on this
guidance, 20% is selected as the RSC for this exposure  scenario.

RSC for Exposure from Ingestion of Fish and Use of Ambient Water for Drinking

In the case on an individual who is exposed to ambient water both by ingestion of
freshwater/estuarine fish and by use of the ambient water for drinking water,  it is
assumed that the water used for ingestion is not treated1. In this case, the equation for
calculating the RSC is as follows:

        RSC = ADD(freshwater fish + Ingestion of ambient water) /ADD(total)

As discussed above, when exposure does not include treated water, the total ADD
about 0.5 ug/kg-day (see Table 5-2). The precise fraction of this total dose that is
contributed by ingestion of freshwater or estuarine fish is unknown because no data
could be located on chloroform levels in freshwater fish.  Likewise, the fraction of the
dose contributed by ingestion of ambient water is also unknown, since no reliable data
could be located on chloroform levels in ambient surface waters of the United States. In
accord with the Exposure Decision Tree approach described in Figure 4-1 of USEPA
(2000b), when data are not available to derive reliable quantitative estimates of all
relevant exposure distributions, a default "floor" value of  20% is recommended for the
RSC (Box 8B).  Based on this guidance,  20% is selected as the RSC for this exposure
scenario.

5.4   EXPOSURE ASSUMPTIONS FOR AMBIENT WATER

As noted above, humans may be exposed to chloroform  in ambient water by ingestion
as drinking water and by ingestion of fish or other aquatic food items collected from the
ambient water body.

In order to ensure that the AWQC value is protective, the combinations of exposure
parameter values selected for quantifying exposure to chloroform from these pathways
are intended to estimate exposures of individuals who are at the upper end of the range
of possible exposure levels.  This is  achieved by selecting one or two of the key
variables in each exposure equation and setting them  at  their high-end values (e.g., 90th
            If the ambient water were treated before use in the house, exposure would be
            evaluated as described for Scenario 1.

                                      42

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                               REVISED DRAFT

to 95th percentile), while other values are held at their median or average value. The
selection of these exposure parameters is described below.

5.4.1  Exposure from Ambient Water Used for Drinking

The basic equation for calculating the average daily dose of chloroform from ingestion
of ambient water that contains a concentration of CAW (ug/L) as drinking water is as
follows:

                  Average Daily Intake(water) = CAW * DIAW /BW

Exposure parameters needed to calculate the daily intake from this pathway are
summarized below:
Parameter
DIAW
BW
Description
daily intake rate
of ambient
water
Body weight
Value
2
70
Units
L/day
kg
Comment
High end value
(AWQC default)
Typical value
(AWQC default)
Source
USEPA
2000b
USEPA
2000b
5.4.2  Exposures from Fish Ingestion

The basic equation for calculating the average daily dose of chloroform from ingestion
of fish collected from an ambient freshwater or estuarine water body that contains a
concentration of CAW (ug/L) is as follows:

                 Average Daily Intake(fish) = CAW x BAF * FI/BW

Because both intake rates and BAFs vary among different types offish, the basic
equation is stratified  into three trophic levels, as follows:

              Average Daily Intake(fish) = CAW x  • 1BAFTU x FITU/BW)

Exposure parameters needed to calculate the ADD from this pathway are summarized
below:
                                      43

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REVISED DRAFT
Parameter
BAFTLI
Flru
BW
Description
Bioaccumulati
on factor for
trophic level i
Ingestion rate
for fish from
trophic level i
Body weight
Value
See
Section 6
TL2 =3.8
TL3 = 8.0
TL4 = 5.7
70
Units
ug/kg tissue
per ug/L
g tissue/day
kg
Comments
Best estimate
based on available
data
High end (90th
percentile) values
(AWQC defaults)
Typical value
(AWQC default)
Source
See
Section 6
USEPA
2000b
USEPA
2000b
     44

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                               REVISED DRAFT
                    6. BIOACCUMULATION FACTORS

This section describes the procedures and data sources used to calculate the
bioaccumulation factors (BAFs) used for deriving an AWQC for chloroform.  Details and
the scientific basis of EPA's recommended methodology for deriving BAFs are
described in USEPA (2000b) and USEPA (2003).

When determining BAFs for use in deriving AWQC for nonionic organic chemicals, two
general steps are required. The first step consists of calculating baseline BAFs for
organisms at appropriate trophic levels using available field and laboratory studies of
the bioaccumulation or bioconcentration of the chemical of interest.  Since baseline
BAFs are normalized by important factors shown to affect bioaccumulation (e.g., the
lipid content of aquatic organisms on which they are based, the freely dissolved
concentration of the chemical in water), they are more generally applicable than BAFs
not adjusted for these factors.  Once baseline BAFs have been calculated for the
appropriate trophic levels, the second step involves adjusting the baseline BAFs to
reflect the expected conditions at the sites that are applicable to the AWQC  (e.g.,  lipid
content of consumed organisms and the freely dissolved fraction of the chemical in the
site water).  Application  of both of these steps to the derivation of BAFs for chloroform is
described below.

6.1   BASELINE BAFs

In EPA's  framework for deriving national BAFs,  several different procedures are
presented by which to derive national BAFs, the applicability of which depends on the
properties of the chemical of interest (USEPA 2000b, 2003).  According to the guidance
provided, nonionic chemicals with log Kow values less than 4.0 should be classified as
exhibiting low hydrophobicity.  The log K^ for chloroform is less than 2, hence it is
considered  to have low hydrophobicity.  Based on the hydrophobicity of chloroform,
national BAFs for chloroform are calculated according to Procedure  3,  as discussed in
detail in USEPA (2000b, 2003). Within Procedure 3, two methods are recommended
for determining the baseline BAFs, depending on the type of bioaccumulation data
available. The data preference for deriving a BAFs for non-polar organics with low
hydrophobicity is (in order of preference):

      Method 1 :   calculation of a  baseline  BAF from a reliable field-measured BAF or
                  laboratory-measured  BCF;

      Method 2:   calculation of a  baseline  BAF from the chemical's  octanol-water
                  partition coefficient
                                      45

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                                REVISED DRAFT

For nonionic organic chemicals that exhibit low hydrophobicity, available information
indicates that non-aqueous exposure to these chemicals is not likely to be important in
determining chemical bioaccumulation in aquatic organisms (USEPA, 2000b).  For this
group of chemicals, Baseline BAFs derived from laboratory-measured BCFs and K^ do
not require adjustment with food-chain multipliers (FCMs) for determining the national
BAFs.

Fish consumption rates determined from the USDA's Continuing Survey of Food Intakes
by Individuals (CSFII) indicate that on a national, average per capita basis, individuals in
the United States consume significant quantities offish and shellfish at trophic levels
two (e.g., clams, oysters), three (e.g., crab, shrimp, flounder) and four (e.g., trout, pike,
certain catfish species) (USEPA 1998e). Therefore, the national AWQC for chloroform
requires that BAFs be derived to reflect bioaccumulation in aquatic organisms at each of
these three trophic levels.

6.1.1  Summary of Field-derived BAF and Laboratory-measured BCF Data

A total of 15  literature reports were located that contained data potentially useful in the
derivation of BAFs for chloroform.  Each was reviewed to determine the relevance and
adequacy of the data reported for application in deriving Baseline BAFs.  Table 6-1
summarizes  the results of this review. As seen,  only one study (Anderson and Lusty
1980) was judged to have potentially useful data on which to base chloroform BAFs  in
freshwater/estuarine species.

Table 6-1 summarizes the data from Anderson and Lusty (1980). Laboratory BCF
values were  not reported by the authors, but data (water and tissue concentrations)
necessary for calculating them were provided. Rainbow trout, bluegill sunfish,
largemouth bass, and channel catfish were exposed to aerated water collected from the
Columbia River as part of a 24-hour static exposure study. Concentrations of
chloroform were measured in whole body tissue  of each fish and in the exposure water.
A semi steady-state condition appeared to be reached within 24 hours for the bluegill,
trout and bass,  but not the catfish.  Maximum concentrations of chloroform in trout and
bass tissues were reached in 4 hours. Measured concentrations of chloroform in fish
tissue were only slightly greater than in the exposure water, indicating that chloroform
has a relatively  low tendency to bioconcentrate in fish tissue.

Further examination of the study by Anderson and Lusty (1980) revealed that the lipid
content of the fish used in the study was not reported. According to the data quality
considerations provided in USEPA (2003), the lipid content of study organisms must be
measured or reasonably estimated in order for a laboratory-measured BCF to be
                                      46

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                     REVISED DRAFT




TABLE 6-1  SUMMARY OF STUDIES ON CHLOROFORM UPTAKE BY FISH
Reference
Baumann-Ofstad et al. 1981. Science
of the Total Environment. 20: 205-215.
Anderson D.R. and E.B. Lusty. 1980.
Acute Toxicity and Bioaccumulation of
Chloroform to Four Species of
Freshwater Fish. U.S. Nuclear
Regulatory Commission
Diamond M.L. 1994. Journal of Great
Lakes Research 20(4): 643-666.
Pearson, C.R. and G. McConnell.
1975. Proc. R. Soc. Lond. 189:
305-332.
Roose, P. and U.A.Th. Brinkman.
1 998. Journal of Chromatography A.
799: 233-248.
Study
Type
Field
Lab
Model
Field
Field
Water Type or
Location
Iddefjord,
Norway
Aerated
Columbia River
Water
Predicted
conditions in
Bay of Quinte,
Lake Ontario
Liverpool Bay,
United
Kingdom
Belgian
Continental
Shelf
Habitat
Type
Marine
NA
Lentic
Marine
Marine
Species
Name
species not specified
Lepomis macrochirus
Oncorhynchus mykiss
Micropterus salmoides
Ictalurus punctatus


Perca flavescens
Perca flavescens
Stizostedion vitreum

Mytilus edulis
Cerastoderma edule
Cancer pagurus
Platycthys flesus
Scomber scombrus
Limanda limanda
Merlangius merlangus
Common
Name
eel
bluegill sunfish
rainbow trout
largemouth bass
channel catfish
plankton
sculpin
Y-of-Y perch
perch
walleye
zooplankton
mussel
cockle
crab
flounder
mackerel
dab
whiting
Trophic
Level
4
3
4
4
4
1-2
3
2-3
3
4
2
2
2
3
4
4
4
3
Data
Evaluation
Inadequate
Accept
Accept
Accept
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Comments on Relevance and Data Adequacy
Article lacks detail on the measurement of
chloroform in water. Only a range of water
concentrations is given - 0.3 to 13 ug/L.
24-hour static exposure. A semi steady-state
condition appeared to be reached withing 24 hrs for
bluegill, trout and bass, but not catfish. Maximum
concentrations of chloroform in trout and bass were
reached in 4 hrs. Chloroform in water was
measured.
Tissue and water chloroform values are estimates
generated using a fugacity model. Water data were
modeled using input from water chloroform values
reported in Comba and Kaiser (1984), J. Gr. Lakes
Res. 1 9:375. Authors state that uncertainty in the
estimates could vary by as much as 10-fold.
Article lacks detail regarding the measurement of
chloroform in water. Only a range of water
concentrations is given - 0.01 to 1 ug/L.
The concentration of chloroform in water was not
measured with the fish samples. The authors,
however, do refer to an article by DeWulf and Van
Langenhove (1995). Internal. J. Environ. Anal.
Chem. 61 :35 who did measure water chloroform
concentrations in water in proximity to the biota
sample sites, and apparently at a similar date.
                          47

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REVISED DRAFT
Reference
Toussaint et al. 2001 . Environmental
Health Perspectives. 109(1):35-40
Voss, R.H. 1983. Environmental
Science Technology. 17(9): 530-537.

Yoshida, K. 1993. Chemosphere.
27(4)' 621-630


Barrows et a I 1980 Toxic Chemicals
pp 379-392


Darnerud et al. 1989. Journal of
Toxicology and Environmental Health
26: 209-221 .

Veith et al. 1980. Aquatic Toxicology.
116-128


Hendriks, A.J. 1995. Ecotoxicology
and Environmental Safety. 32:103-130.



Hiatt et al. 1 981 . Analytical Chemistry.
53:1541-1543.

Paasivirta et al. 1983. Chemosphere.
12(2):239-252
Study
Type
Lab
Lab

Model



Lab



Lab


Lab


Model



Lab

Field
Water Type or
Location
Processed
well water
Bleached
Kraft Mill
Effluents





Well water

Aerated
City of
Uppsala,
Sweden
tapwater
Well water








Lakes in
Central Finland
Habitat
Type
NA
NA

NA







NA


NA


NA



NA

Lentic
Species
Name
Oryzias latipes






Lepomis macrochirus



Oncorhynchus mykiss


Lepomis macrochirus


Oncorhynchus mykiss





Rutilus rutilus
Common
Name
Japanese medaka






bluegill sunfish



rainbow trout


bluegill sunfish


rainbow trout



Unspecified
fish species

roach
Trophic
Level
2






3



4


3


4





3
Data
Evaluation
Inadequate
Inadequate

Inadequate



Inadequate



Inadequate


Inadequate


Inadequate



Inadequate

Inadequate
Comments on Relevance and Data Adequacy
9 month flow-through exposure. Chlorform was not
measured in edible or whole-body tissue.
No tissue data reported.
No original tissue or water data provided. Model
predicted BCF of 15.7 could be calculated based on
fish and drinking water values reported in Table 6 -
Calculated exposure concentrations and doses for
humans.
14-day flow-through exposure. Parent compound
was not quantified in biological tissue. Half-life of
chloroform in tissue estimated to be less than 1 day.
8-hr static exposure. Parent compound was not
quantified in edible biological tissue. Contrary to
rodents, the data indicated that most chloroform
binding in liver and kidney was not irreversibly
bound in the form of electrophilic metabolites.
Data is the same as reported in Barrows et al.
(1980), reference #9, this spreadsheet
Tissue chloroform concentrations were estimated
using a model of lethal response doses from
Hodson et al. (1984). Environ. Toxicol. Chem.
3:243-254 and Hodson (1985). J. Appl. Toxicol.
5:220-226. No corresponding values in water are
provided.
Paper described an analytical methods
development. Chloroform in tissue, sediment, and
water were given as percent recoveries from a
matrix spike (25 ppb).
Chloroform was not measured in water, nor was it in
roach at levels sufficient for quantification.
     48

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REVISED DRAFT
Reference
Young et al. 1 983. Chapter 60 In:
Water Chlorination, Book 2, Vol. 4, pp.
871-884.
Study
Type
Field
Water Type or
Location
San Pedro
Bay,
Los Angeles,
CA
Habitat
Type
Marine
Species
Name
Sicyonia ingentis
Genyonemus lineatus
Scorpaena guttata
Microstomus pacificus
Citharichthys sordidus
Common
Name
ridgeback prawn
white croaker
Calfornia scorpion fish
Dover sole
Pacific sanddab
Trophic
Level
3
3
4
4
4
Data
Evaluation
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Comments on Relevance and Data Adequacy
Tissue chloroform concentrations were reported as
less than values, with fish liver and shrimp muscle
values (< 2 ug/kg ww) below the analyical
quantification limit in tissue of 10 ug/kg ww.
Representative fish tissue values were for a
non-edible tissue (liver). Biota samples were
collected up-coast of the municipal effluent outfall
where chloroform in water was measured.
     49

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                                               REVISED DRAFT
                                TABLE 6-2 SUMMARY OF RELIABLE BCF DATA
Citation



Anderson
and Lusty
(1980)
Habitat
Type


Freshwater

Exposed
Organism


rainbow
trout
Scientific
Name


Oncorhynchus
mykiss
Trophic
Level


4

C(water)
(mg/L)


1
1
C(tissue)
(mg/kg)


3.34
10.35
BCF}
(L/kg)


3.34
10.35
f..
(a)


0.051
0.051
BAFS
(b)


45.9
183.1
Species
BAFS



yi ./
Trophic
Level
Baseline
BAF™

yi . /
(a)
  Lipid content offish tissue was not reported in the study; used species-mean value for rainbow trout from USEPA (2003).
(b) Where fH is assumed to be 1.0.
                                                      50

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                                REVISED DRAFT

acceptable for deriving a national BAF. Although the lipid contents of the organisms
studied by Anderson and Lusty (1980) are not available, USEPA (2003) provides
species-specific average lipid values for fish species commonly consumed by the U.S.
population. This compilation of lipid values includes an average lipid percentage for
rainbow trout (5.1 %), but not for the other species used in the Anderson and Lusty
study. Therefore, a baseline BAF for trophic level 4 will be derived using this average
percent  lipid value for rainbow trout from USEPA (2003) and the rainbow trout data from
Anderson and Lusty (1980).

Due to their low hydrophobicity (i.e., log Kow <4.0), nonionic organic chemicals to which
Procedure 3 is applied are expected to remain almost entirely in the freely dissolved
from in waters containing dissolved and particulate organic carbon concentrations
typical of laboratory BCF studies (USEPA 2000b). Therefore, the freely dissolved
fraction  is usually assumed to  be 1.0.  For the calculation of a baseline BAF from the
BCF study of Anderson and Lusty (1980), the ffd is assumed to be 1.0.

6.1.2  Derivation of Baseline BAFs (BAF^s)

According to the data preference hierarchy specified above, method 1 will be used for
determining the baseline BAFs for trophic level 4. In accordance with this method, each
laboratory-measured BCF (expressed as total concentration in tissue divided by total
concentration in water) was adjusted to a baseline BAF (expressed as lipid-normalized
concentration in tissue divided by freely-dissolved concentration in water) using
Equation 6.1.1 below:
                  Baseline BAF!d =
       BCF/
             -1
                                       'fd
                                                  1
[Equation 6.1.1]
where:
      Baseline BAF[?


      BCFj


      f..
BAF expressed on a freely-dissolved and
lipid-normalized basis (L/kg lipid)

laboratory BCF based on total concentration of
chemical in tissue and water (L/kg tissue)

fraction of the tissue that is lipid (kg/kg)

fraction of the total chemical that is freely-dissolved in
the water of the BAF study
                                       51

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                               REVISED DRAFT

Based on these inputs, the baseline BAF calculations are as shown in the right side of
Table 6-2. For trophic level four, based on a BCF derived for rainbow trout, the baseline
BAFis91.7L/kg-lipid.

Because no acceptable measured BAF or BCF data were located for trophic levels 2
and 3, Baseline BAFs for these trophic levels are estimated by method 2 within
Procedure 3 of the BAF Derivation Framework. The 2000 Human Health Methodology
recommends that individual Baseline BAFs be calculated using as many of the methods
in Procedure 3 as possible, and the selection of the final baseline BAF for each trophic
level be determined from the individual baseline BAF[?s by considering the data
preference hierarchy defined in Procedure 3 and uncertainty in the data. Therefore,
method  2 will also be used to calculate a baseline BAF for trophic level 4.  In
accordance with this method, baseline BAFs for trophic levels 2, 3, and 4 are calculated
from the K^ for chloroform using Equation 6.1.2 below:

                      Baseline BAFefd   = Kow                   [Equation 6.1.2]

where:

      Baseline BAF"      =     BCF expressed on a freely-dissolved and
                              lipid-normalized basis

      K^                =     octanol-water partition coefficient

The log  Kow for chloroform  is 1.97 (see  Section 2). As mentioned previously, food-chain
multipliers (FCMs) are not applicable for chemicals with low hydrophobicity (USEPA
(2000b). Based on these inputs, the baseline BAFs for trophic level 2, 3, and 4 are
calculated as follows:

      Baseline BAFl? (TL2)     = (K^)
                              = (101'97)
                              = 93.3 L/kg-lipid

      Baseline BAFl? (TL3)     = (K^)
                              = (101'97)
                              = 93.3 L/kg-lipid

      Baseline BAFl? (TL4)     = (K^)
                              = (101'97)
                              = 93.3 L/kg-lipid
                                      52

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                                REVISED DRAFT

Due to lack of appropriate field-measured BAFs and laboratory-measured BCFs, there
is only one Baseline BAF™ value for trophic levels 2 and 3.  Therefore, the final Baseline
BAF™ for trophic level 2 is 93.3 L/kg and for trophic level 3 is 93.3 L/kg. For trophic
level 4, there was one laboratory-measured BCF value available from which a Baseline
BAF™ of 91.7 L/kg was derived.  Uncertainties associated with this value include the
fraction lipid value, which was estimated from species mean data rather than from
study-specific fish lipid content determination, and the ffd, which was assumed to be 1.0
based on general characteristics of chemicals with low hydrophobicity rather than from
measured POC and DOC values. The Baseline BAF™s for trophic level 4 calculated by
methods 1 and 2 are essentially the same.  Because both trophic level 4 Baseline BAFs
have moderate to low uncertainty associated with them, the Baseline BAF™ calculated
by method 1 (from laboratory-measured BCF) will be used as the final Baseline BAF™ in
accordance with the data preference hierarchy  outlined for Procedure 3 (Section 6.1;
USEPA2000b).

6.2   National BAFs

After the derivation of trophic level-specific baseline BAFs for chloroform (described in
the previous section), the next step is to calculate BAFs that will be used in the
derivation of AWQC. This step is necessary to adjust the baseline BAFs to conditions
that are expected to affect the bioavailability of  chloroform at the sites applicable to the
AWQC. Derivation of the National BAFs requires information on: (1) the baseline BAF
at appropriate trophic levels, (2) the percent lipid of the aquatic organisms consumed by
humans at the site(s) of interest (trophic-level specific), and (3) the freely dissolved
fraction of the chemical in ambient water at the site(s) of interest.  For each trophic
level, the equation for deriving a BAF to  use in  deriving AWQC is:

      National BAF(TLn) = [(Baseline BAFf?)TL„ ~(f.}TL„ + 17 -ffj       [Equation  6.2.1]

where:

      National BAF(TLn)   =      BAF at trophic level "n" used to derive AWQC based
                               on site conditions for lipid content of consumed
                               aquatic organisms for trophic level "n" and the freely
                               dissolved fraction in the site water

      (Baseline BAF™)TLn =      BAF expressed on a freely dissolved and
                               lipid-normalized basis for trophic level "n"

      (f.)TLn              =      fraction lipid of aquatic species consumed at trophic
                               level "n"
                                       53

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                               REVISED DRAFT

      ffd                 =     fraction of the total chemical in water that is freely
                              dissolved

Each of the components of the National BAF equation are discussed below.

6.2.1  Baseline BAFs (Baseline BAF?)

The derivation of baseline BAFs at specific trophic levels is described in Section 6.1.
For chloroform, a baseline BAF of 93.3 L/kg was derived from K^ for aquatic organisms
in trophic levels 2 and 3.  For organisms in trophic level 4, a baseline BAF of 140.2 L/kg
was derived from a laboratory-measured BCF and a baseline BAF of 93.3 L/kg was
derived from Kow.

6.2.2  Lipid Content of Consumed Aquatic Species

Accumulation of non-polar organic chemicals in aquatic organisms has repeatedly been
shown to be a function of lipid content (e.g., Mackay 1982, Connolly and Pedersen
1988, Thomann 1989). Therefore, baseline BAFs, which are lipid normalized for
comparative purposes, need to be adjusted to reflect the lipid content of aquatic
organisms consumed by the target population.   As discussed in the 2000 Human
Health Methodology (USEPA 2000b) and the Bioaccumulation TSD (USEPA 2003),
EPA recommends that where possible, lipid content of consumed aquatic species be
determined on a consumption-weighted average basis.

For the purposes of deriving national AWQC, EPA has established national default
consumption-weighted lipid content values of 1.9% at trophic level two, 2.6% at trophic
level three, and 3.0% at trophic level four (USEPA, 2003). These national default lipid
content values are based on a  national survey offish and shellfish consumption rates
and information on their lipid content (see USEPA 2000b or 2003 for details of the
determination of national default lipid content values). As discussed in the 2000 Human
Health Methodology, EPA considers the use of national default lipid values appropriate
in situations where local or regional data on lipid content and consumption rates are
unavailable for the site(s) applicable to the AWQC. However, if local or regional data
are available for the site(s) of interest, EPA recommends that States and Tribes use the
local or regional data instead of the national default values because the type and
quantity of consumed aquatic organisms and their lipid content may vary from one
location to another.

6.2.3  Freely-Dissolved  Fraction Applicable to AWQC

Information on the freely-dissolved fraction of the chemical expected at the site(s)
applicable to the AWQC is important because the freely dissolved form of nonionic

                                      54

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                               REVISED DRAFT

organic chemicals is considered to represent the most bioavailable form in water and
thus, the form that best predicts bioaccumulation (USEPA 2000b, 2003).  Freely
dissolved chemical is defined as the portion of the chemical dissolved in water,
excluding the portion sorbed onto  particulate and dissolved organic carbon. The
freely-dissolved fraction is estimated from the octanol-water partition  coefficient and the
dissolved and particulate organic carbon concentrations as shown below.

   f	i	
    '"' [1 +(POC  .  Kow) +(DOC  • 0.08 . Kow)]              [Equation6.2.2]

where:

      ffd                =     freely-dissolved fraction of chemical in water
                              applicable to the AWQC

      POC              =     concentration of particulate organic carbon applicable
                              to the AWQC (kg/L)

      DOC              =     concentration of dissolved organic  carbon applicable
                              to the AWQC (kg/L)

      K^               =     n-octanol-water partition coefficient for the chemical

In this equation, the terms "K^" and "0.08 ••Kow" are used to estimate the partition
coefficients to POC and DOC, respectively, which have units of L/kg, the scientific basis
of which is explained in USEPA (2000b and 2003). Based on national default values of
2.9 mg/L (2.9 x  1Q-06 kg/L) for DOC, 0.48 mg/L (4.8 x 10'07 kg/L) for POC, and 93.3 for
the Kow (log K^  of 1.97), the freely dissolved concentration of chloroform is calculated to
be 0.9999 (expressed as four significant digits for convenience), as follows:

           fa =-	T= 0.9999
               1+(4.8  x10"7 • 93.3J + (2.9 x10"6« 0.08 •93.3H

The national default values for POC and DOC used here are based on the median
value of POC and DOC concentrations observed in numerous water  bodies across the
United States and are described further in USEPA (2000b and 2003). For the purposes
of deriving national AWQC, EPA believes that the use of national default values is
appropriate.  In  addition,  EPA considers the use of national default values of POC and
DOC as being appropriate in situations where local or regional data on POC and DOC
are unavailable  for the site(s) applicable to the AWQC. However, if local or regional
data are available for the site(s) of interest, EPA recommends that States and Tribes
                                      55

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                                REVISED DRAFT

use the local or regional data instead of the national default values because the POC
and DOC can vary on a local basis, thus affecting the freely dissolved fraction.

6.2.4  Calculation of National BAFs

The last step in deriving a national BAF for a given trophic level is to convert the final
baseline BAF™ determined in Section 6.1.2 to a BAF that reflect conditions to which the
national AWQC will apply.  Using Equation 6.2.1 and the inputs described above,  the
BAFs appropriate for calculating national AWQC are calculated as follows:

              National BAF(TLn) = [(Baseline BAFf?)TLn ~(f.)TLn + 1; f fj

      National BAF for Trophic Level Two
            = [(93.3 L/kg-lipid) "(0.019 kg-lipid/kg-tissue) +1] "(0.9999)
            = 2.8 L/kg-tissue

      National AWQC BAF for Trophic Level Three
            = [(93.3 L/kg-lipid) "(0.026 kg-lipid/kg-tissue) +1] "(0.9999)
            = 3.4 L/kg-tissue

      National AWQC BAF for Trophic Level Four
            = [(91.7 L/kg-lipid) "(0.030 kg-lipid/kg-tissue) +1] "(0.9999)
            = 3.8 L/kg-tissue2
      2Use of the Baseline BAF" that was calculated by method 4 (i.e., from f^J results in the same
National BAF as calculated by method 3 here).
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                        7.  AWQC CALCULATION

7.1    FOR AMBIENT WATERS USED FOR DRINKING WATER SOURCES

The basic equation used to calculate the AWQC that is protective of an individual who is
exposed to ambient water both by ingestion of fish and by use of the ambient water as
an indoor drinking water source is as follows:


                AWQC = RfD x RSC x          BW
                                           ^(FITUxBAFTLi)]
where:
 RfD  =     Oral reference dose for chloroform (ug/kg-day)
 RSC =     Fraction of total exposure attributable to ambient water
 BW  =     Body weight (kg)
 IR   =     Ingestion rate of ambient water used for drinking (L/day)
 FITLi =     Fish intake rate for trophic level "i" (g/day)
 BAFTLI =   Bioaccumulation factor for chloroform in fish of trophic level "i" (ug/g per
            M9/L)

Table 7-1 summarizes the input values needed to calculate the AWQC.  Based on these
inputs, the resulting AWQC value for this scenario is 68 ug/L.
                                     57

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                   REVISED DRAFT




TABLE 7-1 SUMMARY OF TOXICITY AND EXPOSURE PARAMETERS
Pathway
All
Drinking
water
ingestion
Fish
Ingestion
Variable
RfD
RSC
IR
RSC
Fl (TL 2)
Fl (TL 3)
Fl (TL 4)
BAF (TL 2)
BAF (TL 3)
BAF (TL 4)
Units
ug/kg-d
~
L/day
~
g/day
g/day
g/day
L/kg
L/kg
L/kg
Value
10
0.20
2
0.2
3.8
8.0
5.7
2.8
3.4
3.8
Comment
Agency consensus value
Default "floor" value derived using decision tree approach
AWQC default value
Default "floor" value derived using decision tree approach
AWQC default value
AWQC default value
AWQC default value
Calculated from available data
Calculated from available data
Calculated from available data
Source
See Section 4
See Section 5
USEPA 2000b
See Section 5
USEPA 2000b
USEPA 2000b
USEPA 2000b
See Section 6
See Section 6
See Section 6
                        58

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                               REVISED DRAFT

7.2   FOR AMBIENT WATERS NOT USED FOR DRINKING WATER SOURCES

The basic equation used to calculate the AWQC that is protective of an individual who is
exposed to ambient water by ingestion of fish is as follows:
                                               BW
                    AWQC=RfDxRSCx
where:

 RfD =     Oral reference dose for chloroform (ug/kg-day)
 RSC =     Fraction of total exposure attributable to ambient water
 BW =     Body weight (kg)
 FITLi =     Fish intake rate for trophic level "i" (g/day)
 BAFTLj=    Bioaccumulation factor for chloroform in fish of trophic level "i" (ug/g per
            M9/L)

Table 7-1 summarizes the input values needed to calculate the AWQC.  Based on these
inputs, the resulting AWQC  value for this scenario is  2,400 ug/L.

7.3   AWQC SUMMARY

Based on the equations and input parameters described above, the AWQC values for
chloroform are as follows:
AWQC Type
Ingestion of drinking water plus ingestion of
organisms
Ingestion of organisms only
AWQC Value (a)
68 ug/L
2,400 ug/L
            (a) AWQC values are shown to two significant figures
                                      59

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    60

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                              REVISED DRAFT
    8.  SITE-SPECIFIC OR REGIONAL ADJUSTMENTS TO CRITERIA

Several parameters in the AWQC equation can be adjusted on a site-specific or
regional basis to reflect regional or local conditions and/or specific populations of
concern. These include fish consumption; incidental water consumption as related to
regional/local recreational activities; BAF (including factors used to derive BAFs such as
POC/DOC, percent lipid offish consumed by target population, and species
representative of given trophic levels); and the relative source contribution. States and
Tribes are encouraged to make adjustments using the information and instructions
provided in the 2000 Human Health Methodology (USEPA 2000b) and its supporting
Technical Support Documents for Exposure Assessment (Volume II) and
Bioaccumulation (Volume III) when they become available.
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Draft
    62

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                              REVISED DRAFT
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