EPA/635/R02/002
f/EPA
TOXICOLOGICAL REVIEW
OF
1,1-DICHLOROETHYLENE
(CAS No. 75-35-4)
In Support of Summary Information on the
Integrated Risk Information System (IRIS)
June 2002
U.S. Environmental Protection Agency
Washington, DC
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DISCLAIMER
This document has been reviewed in accordance with U.S. Environmental Protection Agency
policy and approved for publication. Mention of trade names or commercial products does not
constitute endorsement or recommendation for use. Note: This document may undergo
revisions in the future. The most up-to-date version will be made available electronically via the
IRIS Home Page at http://www.epa.gov/iris.
11
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CONTENTS —TOXICOLOGICAL REVIEW FOR 1,1-DICHLOROETHYLENE
(CAS No. 75-35-4)
FOREWORD v
AUTHORS, CONTRIBUTORS, AND REVIEWERS vi
1. INTRODUCTION 1
2. CHEMICAL AND PHYSICAL INFORMATION RELEVANT TO ASSESSMENTS .... 2
3. TOXICOKINETICS RELEVANT TO ASSESSMENTS 3
4. HAZARD IDENTIFICATION 6
4.1. STUDIES IN HUMANS—EPIDEMIOLOGY 6
4.2. PRECHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN
ANIMALS—ORAL AND INHALATION 7
4.2.1. Acute Exposure 7
4.2.2. Longer-Term Exposure 11
4.2.3. Chronic Studies and Cancer Bioassays 14
4.3. REPRODUCTIVE AND DEVELOPMENTAL STUDIES—ORAL AND
INHALATION 21
4.3.1. Direct Infusion 21
4.3.2. Oral 22
4.3.3. Inhalation 24
4.4. OTHER STUDIES 25
4.4.1. Developmental Neurotoxicity 25
4.4.2. Cardiac Sensitization 26
4.4.3. Species Specificity 26
4.4.4. Genetic Toxicity 26
4.5. SYNTHESIS AND EVALUATION OF MAJOR NONCANCER EFFECTS
AND MODE OF ACTION 27
4.6. WEIGHT-OF-EVIDENCE EVALUATION AND CANCER
CHARACTERIZATION 33
4.7. SUSCEPTIBLE POPULATIONS 35
4.7.1. Possible Childhood Susceptibility 35
4.7.2. Possible Gender Differences 36
5. DOSE-RESPONSE ASSESSMENTS 36
5.1. ORAL REFERENCE DOSE (RfD) 36
5.1.1. Choice of Principal Study and Critical Effect 36
5.1.2. Methods of Analysis 37
5.1.3. RfD Derivation 37
in
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CONTENTS (continued)
5.2. INHALATION REFERENCE CONCENTRATION (RfC) 38
5.2.1. Choice of Principal Study and Critical Effect 38
5.2.2. Methods of Analysis 38
5.2.3. RfC Derivation 39
5.3. CANCER ASSESSMENT 40
6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HAZARD AND
DOSE RESPONSE 40
6.1. HUMAN HAZARD POTENTIAL 40
6.2. DOSE RESPONSE 41
REFERENCES 42
APPENDIX A. Summary of External Peer Review Comments and Disposition 54
APPENDIX B. Benchmark Dose Calculations 62
IV
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FOREWORD
The purpose of this Toxicological Review is to provide scientific support and rationale
for the hazard and dose-response assessment in IRIS pertaining to chronic exposure to 1,1-
dichloroethylene. It is not intended to be a comprehensive treatise on the chemical or
toxicological nature of 1,1-dichloroethylene.
In Section 6, EPA has characterized its overall confidence in the quantitative and
qualitative aspects of hazard and dose response. Matters considered in this characterization
include knowledge gaps, uncertainties, quality of data, and scientific controversies. This
characterization is presented in an effort to make apparent the limitations of the assessment and
to aid and guide the risk assessor in the ensuing steps of the risk assessment process.
For other general information about this assessment or other questions relating to IRIS,
the reader is referred to EPA's IRIS Hotline at 301-345-2870.
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AUTHORS, CONTRIBUTORS, AND REVIEWERS
U.S. EPA Region 8 and the Office of Solid Waste and Emergency Response (OSWER)
were responsible for preparing the IRIS toxicological review and summary documents. A
comprehensive literature review was conducted in September 1999. The literature review was
supplemented with additional references until May 2002.
Chemical Manager/Author
Robert Benson
Municipal Systems
Office of Regulatory Assistance
U.S. EPA Region 8, Denver, Colorado
Reviewers
This document and summary information on IRIS have received peer review both by
EPA scientists and by independent scientists external to EPA. Subsequent to external review
and incorporation of comments, this assessment has undergone an Agency-wide review process
whereby the IRIS Program Manager has achieved a consensus approval among the Office of
Research and Development; Office of Air and Radiation; Office of Prevention, Pesticides, and
Toxic Substances; OSWER; Office of Water; Office of Policy, Planning, and Evaluation; and the
Regional Offices.
Internal EPA Reviewers
NCEA-Washington NCEA-RTP
Jim Cogliano Carole Kimmel Judy Strickland
Lynn Flowers Karen Hogan
Colorado Department of Public Health and Environment
Diane Niedzwiecki
External Peer Reviewers
Melvin E. Andersen James V. Bruckner
Colorado State University University of Georgia
Ft. Collins, Colorado Athens, Georgia
Poh-Gek Forkert Sam Kacew
Queen's University University of Ottawa
Kingston, Ontario, Canada Ottawa, Ontario, Canada
VI
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Kannan Krishnan
University of Montreal
Montreal, Quebec, Canada
A summary of the external peer reviewers' comments and the disposition of their
recommendations are in Appendix A.
vn
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1. INTRODUCTION
This document presents background and justification for the hazard and dose-response
assessment summaries in the U.S. Environmental Protection Agency's (EPA's) Integrated Risk
Information System (IRIS). IRIS summaries may include an oral reference dose (RfD), an
inhalation reference concentration (RfC), and a carcinogenicity assessment.
The RfD and RfC provide quantitative information for noncancer dose-response
assessments. The RfD is based on the assumption that thresholds exist for certain toxic effects,
such as cellular necrosis, but may not exist for other toxic effects, such as some carcinogenic
responses. It is expressed in units of mg/kg-day. In general, the RfD is an estimate (with
uncertainty spanning perhaps an order of magnitude) of a daily exposure to the human
population (including sensitive subgroups) that is likely to be without an appreciable risk of
deleterious noncancer effects during a lifetime. The inhalation RfC is analogous to the oral RfD,
but provides a continuous inhalation exposure estimate. The inhalation RfC considers toxic
effects for both the respiratory system (portal-of-entry) and for effects peripheral to the
respiratory system (extrarespiratory or systemic effects). It is generally expressed in units of
mg/m3.
The carcinogenicity assessment provides information on the carcinogenic hazard
potential of the substance in question and quantitative estimates of risk from oral exposure and
inhalation exposure. The information includes a weight-of-evidence judgment of the likelihood
that the agent is a human carcinogen and the conditions under which the carcinogenic effects
may be expressed. Quantitative risk estimates are presented in three ways. The slope factor is
the result of application of a low-dose extrapolation procedure and is presented as the risk per
mg/kg-day. The unit risk is the quantitative estimate in terms of either risk per |ig/L drinking
water or risk per |ig/m3 air breathed. Another form in which risk is presented is a drinking water
or air concentration providing cancer risks of 1 in 10,000; 1 in 100,000; or 1 in 1,000,000.
Development of these hazard identification and dose-response assessments for 1,1-
dichloroethylene (DCE) has followed the general guidelines for risk assessment as set forth by
the National Research Council (1983). EPA guidelines that were used in the development of this
assessment may include the following: the Guidelines for Carcinogen Risk Assessment (U.S.
EPA, 1986a), Guidelines for the Health Risk Assessment of Chemical Mixtures (U.S. EPA,
1986b), Guidelines for Mutagenicity Risk Assessment (U.S. EPA, 1986c), Guidelines for
Developmental Toxicity Risk Assessment (U.S. EPA, 1991), Proposed Guidelines for Carcinogen
Risk Assessment (U.S. EPA, 1996a), Guidelines for Reproductive Toxicity Risk Assessment (U.S.
EPA, 1996b), and Guidelines for Neurotoxicity Risk Assessment (U.S. EPA, 1998a);
Recommendations for and Documentation of Biological Values for Use in Risk Assessment (U.S.
EPA, 1988); (proposed) Interim Policy for Particle Size and Limit Concentration Issues in
Inhalation Toxicity (U.S. EPA, 1994a); Methods for Derivation of Inhalation Reference
Concentrations and Application of Inhalation Dosimetry (U.S. EPA, 1994b); Peer Review and
Peer Involvement at the U.S. Environmental Protection Agency (U.S. EPA, 1994c); Use of the
Benchmark Dose Approach in Health Risk Assessment (U.S. EPA, 1995); Draft Revised
1
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Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999); Science Policy Council
Handbook: Peer Review (U.S. EPA, 1998b, 2000a); Memorandum from EPA Administrator,
Carol Browner, dated March 21, 1995, Policy for Risk Characterization; and Science Policy
Council Handbook, Risk Characterization (U.S. EPA, 2000b).
Literature search strategies employed for this compound were based on the CASRN and
at least one common name. At a minimum, the following databases were searched: RTECS,
HSDB, TSCATS, CCRIS, GENETOX, EMIC, EMICBACK, DART, ETICBACK, TOXLINE,
CANCERLINE, MEDLINE, and MEDLINE backfiles. Any pertinent scientific information
submitted by the public to the IRIS Submission Desk was also considered in the development of
this document.
EPA has previously reviewed 1,1-DCE (U.S. EPA, 1985a, b). This review replaces those
assessments.
2. CHEMICAL AND PHYSICAL INFORMATION RELEVANT TO ASSESSMENTS
1,1-DCE does not occur naturally. It is produced commercially by the
dehydrochlorination of 1,1,2-trichloroethane in the presence of excess base. 1,1-DCE is used
principally for the production of polyvinylidene chloride polymers (PVDC). PVDC is used
principally in the food packaging industry as cast and extruded film (Saran and Velon wraps)
and as a barrier coating for paper, cellulose, polypropylene, and other plastics. Extruded
filaments of PVDC are also used in the textile industry for furniture and automobile upholstery,
drapery fabric, and outdoor furniture. 1,1-DCE enters in the environment though release during
its manufacture and use, from the breakdown of PVDC products, and from the biotic or abiotic
breakdown of 1,1,1-trichloroethane, tetrachloroethylene, 1,1,2-trichoroethene, and 1,1-
dichloroethane (ATSDR, 1994; IARC, 1999; and U.S. EPA, 1985 a, b).
The chemical and physical properties of 1,1-DCE (ATSDR, 1994; IARC, 1999) are
presented below.
CAS name: 1,1-dichloroethene
CAS number: 75-35-4
IUPAC name: 1,1-dichloroethylene
Primary synonyms: 1,1-DCE; vinylidene chloride, vinylidene dichloride
Chemical formula: C2H2C12
Chemical structure: C12C=CH2
Molecular weight: 96.94
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Boiling point:
Melting point:
Specific gravity:
Vapor pressure:
Solubility:
Odor:
Odor threshold:
Partition coefficients:
Flash point:
Autoignition:
Conversion factor:
31.6 °C
-122.5 °C
1.218
67 kPa at 20 °C
Practically insoluble in water; soluble in acetone, ethanol, and
many organic solvents; very soluble in diethyl ether.
Mild, sweet, resembling chloroform
500 ppm in air; no data in water
Log Kow
LogKoc
1.32
1.81
-19 °C, closed cup; -15 °C, open cup
570 °C
1 ppm = 3.97 mg/m3
3. TOXICOKINETICS RELEVANT TO ASSESSMENTS
1,1-DCE is rapidly absorbed following inhalation and oral exposures. Because of its low
molecular weight and hydrophobic nature, dermal absorption is also likely, but no relevant data
were found in the literature. In rats treated with 1,1-DCE by gavage in corn oil, complete
gastrointestinal absorption was found to occur at <350 mg/kg (Jones and Hathway, 1978a, b;
Putchaetal., 1986). 1,1-DCE is easily transported across the alveolar membrane. At constant
<750 ppm concentration in the air, equilibrium or near steady-state is reached in the blood in rats
in approximately 45 minutes (Dallas et al., 1983). Continued uptake in rats reflects to some
extent continuing deposition in fatty tissues, but this is primarily a result of metabolism of 1,1-
DCE.
The major route of excretion for unchanged 1,1-DCE is through the lung (Jones and
Hathway, 1978a). However, the majority of 1,1-DCE is rapidly metabolized to nonvolatile
compounds and covalently bound derivatives (McKenna et al., 1997, 1978a, b). Mice
metabolize more 1,1-DCE than do rats. For example, when given 50 mg/kg by oral gavage in
corn oil, mice excreted 6% and rats excreted 28% of the dose as unchanged 1,1-DCE through the
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lungs (Jones and Hathway, 1978b). When exposed to 10 ppm for a single 6-hour episode, mice
excreted 0.65% and rats excreted 1.63% of the absorbed dose as unchanged 1,1-DCE through the
lungs (McKenna et al., 1977). Intraperitoneal (i.p.) administration of 125 mg/kg 14C-1,1-DCE to
mice resulted in the highest concentrations of covalent binding (based on protein content) in the
kidney, lung, and liver (Okine et al., 1985; Okine and Gram, 1986a, b). The covalent binding
and cellular damage in kidney, lung, and liver correlated with the high concentration of CYP2E1
in certain cell populations in these tissues.
The proposed metabolic pathways for 1,1-DCE are summarized in Figure 1. These
pathways were determined from experimental studies in laboratory animals. It is not known
whether the metabolism of 1,1-DCE is the same in humans, although in vitro microsomal
preparations from human liver and lung form the same initial products (Dowsley et al., 1999).
Oxidation of 1,1-DCE by CYP2E1 should produce three metabolites: DCE epoxide, 2-
chloroacetyl chloride, and 2,2-dichloroacetaldehyde. All of these metabolites react with
glutathione (GSH) and/or water. In the kidney, further metabolism of-S'-(2,2-dichloro-l-
hydroxy)ethylglutathione could form another toxic compound, dicholorothioketene. The GSH
conjugates formed are catabolized in the kidney to a variety of urinary excretion products. The
epoxide, and perhaps to a lesser extent the chloroacetaldehyde, are believed to be associated with
the tissue reactivity and toxic effects in tissues that ensue after significant depletion of GSH.
The primary metabolites of 1,1-DCE formed in rat hepatic microsomal incubations are
DCE epoxide, 2,2-dichloroacetaldehyde, and 2-chloroacetyl chloride (Liebler et al., 1985, 1988;
Costa and Ivanetich, 1982). These metabolites were also identified from mouse microsomal
incubations (Dowsley et al., 1995). All these electrophilic metabolites undergo secondary
reactions, including oxidation, conjugation with GSH, and hydrolysis. The major products
formed are GSH conjugates, 2-(,S'-glutathionyl)acetyl glutathione [B], and 2-,S'-glutathionyl
acetate [C], which are believed to be derived from the DCE epoxide (Fig. 1). ,S'-(2,2-Dichloro-l-
hydroxy ethyl glutathione [A], the GSH conjugate formed from reaction of GSH with 2,2-
dichloroacetaldehyde, was not observed in rat liver microsomal incubations containing GSH
(Dowsley et al., 1995). The acetal, together with chloroacetic acid and ,S'-(2-chloroacetyl)-
glutathione [D]—the hydrolysis and GSH-conjugated products of 2-chloroacetyl chloride,
respectively—was detected at levels much lower than those for the DCE epoxide-derived
conjugates [B] and [C].
In human liver and lung microsomal incubations, the DCE epoxide-derived GSH
conjugates [B] and [C] were the major metabolites detected (Dowsley et al., 1999). 2,2-
Dichloroacetaldehyde was detected at low levels. Liver microsomes from three out of five
human samples metabolized 1,1-DCE to the epoxide-derived GSH conjugates at levels that were
2.5- to 3-fold higher than in mouse liver microsomes, based on milligrams of microsomal
protein. These GSH conjugates were also the major products formed in lung microsomes from
eight human samples; only low levels of 2,2-dichloroacetaldehyde were formed. The mean level
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ci
Tissue Targets
& Resultant Toxicity
[D]
GSH
GS
Cl
DCE-Epoxide GSH/
GS
\H,0
H,0
GS
V-/
-A
OH
[B]
Tissue Targets
& Resultant Toxicity
[C]
Figure 1: Proposed pathways for 1,1-DCE metabolism and toxicity.
Source: Adapted from Forkert, 1999a, b
in lung microsomes from humans was about 50% of the amount formed in lung microsomes
from mice. In both animal and human tissues, cytochrome P450 CYP2E1 catalyzes the
formation of the DCE epoxide (Dowsley et al., 1996).
The significance of the metabolic pathway in the liver involving 2,2-
dichloroacetaldehyde is unclear. Existing evidence, however, suggests that this pathway is of
minor toxicological importance. In addition to 2,2-dichloroacetaldehyde and the GSH conjugate,
potential metabolites include the acetal (the hydration product of the aldehyde), dichloroacetic
acid, and dichloroethanol. An initial study with rat liver microsomes found a trace level of 2,2-
dichloroacetaldehyde but no detectable dichloroacetic acid (Costa and Ivanetich, 1982). A later
report using isolated rat hepatocytes detected dichloroacetic acid and trace levels of 2,2-
dichloroacetaldehyde, 2,2-dichloroethanol, and chloroacetic acid (Costa and Ivanetich, 1984).
Forkert (1999a) and Forkert and Boyd (2001), using intact mice, found no acetal in liver cytosol;
however, acetal was detected in the bile in the first study but was not mentioned as being found
in the bile in the second study. In early studies on the metabolism of 1,1-DCE, none of the
potential metabolites from this pathway were reported as being found in the urine of rodents
using techniques that readily identified chloroacetic acid (Jones and Hathway, 1978a, b;
McKenna et al., 1977, 1978a, b). A pharmacokinetic analysis showed that any dichloroacetic
acid formed in the liver is rapidly metabolized in the liver to two carbon, nonchlorinated
chemicals and carbon dioxide (Merdink et al., 1998).
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The oxidative metabolism of 1,1 -DCE has been found to reach saturation in rats at an
oral exposure of 10-50 mg/kg and an inhalation exposure of 200 ppm (794 mg/m3) (Andersen et
al., 1979; D'Souza and Andersen, 1988; Dallas et al., 1983; McKenna et al., 1977).
Because 1,1-DCE is lipophilic and has a blood-to-air partition coefficient of 5 in rats
(D'Souza and Andersen, 1988), any 1,1-DCE not metabolized following oral or inhalation
exposure is rapidly exhaled unchanged when exposure is terminated. Because of its low
octanol:water partition coefficient, 1,1-DCE will not bioaccumulate in tissues to a significant
extent. The major metabolites found in urine of rodents include oxalic acid, thiodiglycolic acid,
thioglycolic acid, dithioglycolic acid, N-acetyl-S-(2-carboxymethyl) cysteine, N-acetyl-S-(2-
hydroxyethyl) cysteine, other -acetyl-S-cysteinyl derivatives, and
methylthioacetylaminoethanol.
D'Souza and Andersen (1988) developed physiologically based pharmacokinetic (PBPK)
models for 1,1-DCE in the rat for both oral and inhalation exposure. No validated model is
available for humans. D'Souza and Andersen (1988) used allometric scaling to estimate
comparative amounts of epoxide formed (mg/kg) in rats and humans. Cardiac output and
pulmonary ventilation were scaled by (body weight)0'7, Vmax was scaled by (body weight)0'74,
and body fat was estimated at 7% in the 200 g rat and 20% in the 70 kg human. When the oral
exposure was less than 5 mg/kg, the estimated amount of epoxide formed was about the same in
rats and humans. When the inhalation exposure was less than 100 ppm, the estimated amount of
epoxide formed was fivefold lower in humans than in rats.
El-Masri et al. (1996a, b) used a combination of gas uptake experiments in Sprague-
Dawley rats and PBPK modeling to assess the potential for interaction between 1,1-DCE and
trichloroethylene. Both substrates are activated by CYP2E1. Thus, there is a potential for
competitive inhibition when simultaneous exposure to both substrates occurs. The results of the
gas uptake experiments confirmed a model based on competitive inhibition. There was,
however, no evidence of competitive inhibition when exposure to both substrates was 100 ppm
or less. As environmental exposures to these chemicals are expected to be less than 100 ppm,
there is little potential for reduced toxicity from 1,1-DCE when individuals are also exposed to
tri chl oroethy 1 ene.
4. HAZARD IDENTIFICATION
4.1. STUDIES IN HUMANS—EPIDEMIOLOGY
Ott et al. (1976) investigated the health records of 138 employees who were
occupationally exposed to 1,1-DCE in processes not involving vinyl chloride. The individuals
included in the study had worked in experimental or pilot plant polymerization operations, in a
monomer production process as tankcar loaders, or in a production plant manufacturing a
monofilament fiber. Time-weighted average concentrations (8 hours) of 1,1-DCE in the
workplace were estimated from job descriptions and the results of industrial hygiene sampling.
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The subjects were grouped into three exposure categories: less than 10 ppm, 10-24 ppm, and
greater than 25 ppm. The researchers estimated career exposure by taking into account average
duration of employment. Results of the most recent health inventory for individuals in the
exposed cohort were compared with findings for matched controls. An analysis of mortality in
the cohort indicated no statistically significant differences. Overall, there were no significant
differences in hematology and clinical chemistry parameters between the exposed cohort and the
controls.
Three reports suggest an association between exposure to dichloroethylenes and birth
defects. The California Department of Health Services (Swan et al., 1985) reported an increase
in the number of cardiac congenital anomalies during 1980 and 1981 in an area served by a
public water supply contaminated with 1,1,1-trichloroethane and dichloroethylene. The public
water supply also contained chlorinated disinfection by products. Goldberg et al. (1990)
reported an increase in congenital cardiac malformations between 1969 and 1987 in an area of
Arizona where the drinking water was contaminated with trichloroethylene and dichloroethylene
(isomer not specified). The dichloroethylene concentration in the drinking water was usually 5%
to 10% of the trichloroethylene concentration. The paper does not specify whether the drinking
water was chlorinated. Finally, Bove et al. (1995) reported increased odds ratio for oral cleft
defects (1.71), for central nervous system defects (2.52), and for neural tube defects (2.60)
associated with exposure to total dichloroethylenes of more than 2 |ig/L from public drinking
water supplies in an area of northern New Jersey. The period of time studied was 1985 to 1988.
The drinking water also contained chlorinated disinfection by-products. It is not clear from the
paper whether there was also co-exposure to other chlorinated solvents also reported on,
including trichloroethylene, tetrachloroethylene, 1,1,1-trichloroethane, carbon tetrachloride, and
1,2-dichloroethane. As all of these situations involved exposure to multiple contaminates, a
cause-and-effect relationship between the reported birth defects and exposure to 1,1-DCE cannot
be established.
4.2. PRECHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN
ANIMALS—ORAL AND INHALATION
4.2.1. Acute Exposure
Mice are more sensitive than rats to acute toxicity from 1,1-DCE. The National
Toxicology Program (NTP) (NTP, 1982) conducted a study to determine lethality in five male
and five female F344 rats and five male and five female B6C3Fj mice (all animals 9 weeks old)
after a single exposure to 1,1-DCE by gavage in corn oil at 0, 10, 50, 100, 500, or 1000 mg/kg.
By day 14 postexposure, mortality was 0/10, 1/10, 0/10, 0/10, 1/10, and 2/10 in the rats and 0/10,
0/10, 1/10, 0/10, 8/10, and 10/10 in the mice, respectively. Other representative lethality data are
presented in Tables 1 and 2.
Table 1. Representative lethality (LD50) from oral exposure to 1,1-DCE
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Species
Rat
male
male
male, adrenalectomized
female
Mouse
male
female
Dose
(mg/kg)
1550
1800
84
1500
217
194
Effect
LD50
LD50
Reference
Jenkins et al., 1972
Ponomarkov and Tomatis, 1980
Jenkins et al., 1972
Ponomarkov and Tomatis, 1980
Jones and Hathway, 1978b
Table 2. Representative lethality (LC50) or time for 50% lethality (LT50) from
inhalation exposure to 1,1-DCE
Species
Rat
male, fed
male, fasted
Mouse
male
female
Exposure
(ppm)
63 50 for 4 hr
200 for 4.1 hr
400 for 3. 6 hr
500 for 3.0 hr
1000 for 2.4 hr
2000 for 1.4 hr
98 for 22-23 hr
105 for 22-23 hr
Effect
LC50
LT50
LC50
Reference
Siegaletal., 1971
Andersen et al., 1979
Short etal, 1977a
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Toxicity is enhanced by fasting (Andersen and Jenkins, 1977; Chieco et al., 1981; Jaeger
et al., 1974, 1975, 1977a, b; McKenna et al., 1978a, b; Moslen et al., 1985), by GSH depletion
(Andersen et al., 1980; Jaeger et al., 1974, 1977a, b; Kanz et al., 1988; Moussa and Forkert,
1992), and by administration in oil vehicles compared to administration in aqueous Tween
(Chieco et al., 1981). Toxicity is decreased by agents that decrease metabolism by the P450
system (Andersen et al., 1978; Moslen et al., 1989) or by hypothyroidism, which increases
intracellular GSH (Kanz et al., 1991).
The target organs for toxicity after acute oral or inhalation exposure are the liver, the
kidney, and the Clara cells of the lung. The effects in the liver include an increase in liver
enzymes in the serum (Jenkins et al., 1972; Jaeger, 1977a, b; Short et al., 1977a; Jenkins and
Andersen, 1978; Reynolds et al., 1980); severe histopathological damage, including disruption of
bile canaliculi, cytoplasmic vacuolization, and hemorrhagic necrosis (Short et al., 1977a; Kanz
and Reynolds, 1986; Reynolds et al., 1984); an increase in covalent binding of 1,1-DCE (Forkert
and Moussa, 1991, 1993; Jaeger et al., 1977a, b); and a decrease in GSH (Forkert and Moussa,
1991, 1993; Kanz et al., 1988; Reichert et al., 1978, 1979) mediated by CYP2E1 metabolism of
1,1-DCE to intermediates that react with GSH (Kainz et al., 1993; Lee and Forkert, 1994).
Several researchers have investigated the hepatotoxicity of 1,1-DCE. In a study by
Jenkins and Andersen (1978), four female Sprague-Dawley rats (body weight, 223 g) received a
single oral exposure by gavage in corn oil at 400 mg/kg. Four to 8 hours after exposure, there
was a significant increase in aspartate aminotransferase (approximately 75-fold), alanine
aminotransferase (approximately 70-fold), lactate dehydrogenase (approximately 110-fold), and
sorbitol dehydrogenase (approximately 320-fold). The serum enzymes returned to
approximately normal values within 82 hours after exposure.
Reynolds et al. (1984) administered a single oral exposure by gavage in mineral oil at
200 mg/kg 1,1-DCE to fasted male Sprague-Dawley rats (body weight, 225-375 g). Within 2
hours after exposure, the livers showed evidence of dilatation and disruption of bile canaliculi,
plasma membrane invagination and loss of microvilli, cytoplasmic vacuolization, and loss of
density in mitochondrial matrices. One hour after a single inhalation exposure at 250 ppm 1,1-
DCE for 4 hours, Sprague-Dawley rats showed a significant decrease (/X0.05) in GSH
concentration in the liver (Jaeger, 1977). Four hours after exposure there was an increase in the
serum concentration of sorbitol dehydrogenase (approximately 230-fold) and ornithine
carbarnoyl transferase (approximately 380-fold).
Short et al. (1977a) studied CD-I male mice (Charles River) and CD male rats (Charles
River) exposed by inhalation for 22-23 hrs/day for 1-5 days at 0, 15, 30, or 60 ppm 1,1-DCE
(mice) or for 1-3 days at 0 or 60 ppm (rats). In male mice exposed to > 15 ppm, serum enzymes
(alanine aminotransferase and aspartate aminotransferase) were significantly increased (four- to
sixfold), and hepatocellular degeneration was observed in one of five mice after the first
exposure. In two of five male rats exposed to 60 ppm, mild centrilobular degeneration and/or
necrosis was observed after the first exposure, but serum enzymes (alanine aminotransferase and
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aspartate aminotransferase) were not significantly increased (four- to sixfold) until after the
second exposure.
Reynolds et al. (1980) found that after a single 4-hour exposure by inhalation at 200 ppm
1,1-DCE, the liver of fasted male Sprague-Dawley rats (body weight, 150-200 g) showed
catastrophic morphological alterations of the parenchymal cells, including retraction and central
rarefaction of nuclei with peripheral displacement of chromatin to nuclear margins, progressing
to frank hemorrhagic centrilobular necrosis. GSH concentrations were also depleted. After the
extensive hepatocellular damage, cytochrome P450 and oxidative -demethylase were
deactivated.
Toxic effects of 1,1-DCE exposure in the kidney include increased kidney weight,
increased blood urea nitrogen and creatinine (Jackson and Conolly, 1985; Jenkins and Andersen,
1978), and histopathological changes, including vacuolization, tubular dilatation, and nephrosis
and necrosis of the proximal tubules (Short et al., 1977a; Jackson and Conolly, 1985; Jenkins
and Andersen, 1978). These changes were correlated with metabolic activation of 1,1-DCE by
CYP2E1 in the proximal tubules, decreased GSH concentration, increased covalent binding of
1,1-DCE, and the presence of a relatively high concentration of p-lyase activity in rodent kidney
tissue (Brittebo et al., 1993; Dekant et al., 1989; Dekant, 1996). In addition, renal toxicity can
be inhibited by pretreatment of mice and rats with aminooxyacetic acid, an inhibitor of renal
cysteinyl-B-lyase (Ban et al., 1995; Cavelier et al., 1996).
Jenkins and Andersen (1978) investigated the nephrotoxicity of 1,1-DCE in Sprague-
Dawley rats after a single oral exposure by gavage in corn oil. Fasted male rats (two to six per
group; body weight, 300 g) were administered 0, 50, 100, 200, 400, or 600 mg/kg. At 600
mg/kg, there was a fivefold increase in blood urea nitrogen. Histopathological examination was
not conducted in animals treated at 600 mg/kg. In male rats at 400 mg/kg, there was a
statistically significant increase (p<0.05) in blood urea nitrogen (fourfold) and in creatinine
(threefold). The increases became apparent 8 hours after exposure, reached a peak 24 hours after
exposure, and returned to normal 96-144 hours after exposure. In male rats at 400 mg/kg, there
was also a twofold increase in relative kidney weight 48 hours after exposure. The relative
kidney weight had nearly returned to normal 144 hours after exposure. In female rats at 400
mg/kg, there was no substantial increase in blood urea nitrogen, creatinine, or relative kidney
weight. Histopathological lesions (tubular dilatation and tubular necrosis) were observed in both
sexes at 400 mg/kg. No significant effects were seen at 200 mg/kg and below.
Short et al. (1977a) studied CD-I male mice (Charles River) after inhalation exposure for
22-23 hrs/day for 1-5 days at 0, 15, 30, or 60 ppm 1,1-DCE. Tubular nephrosis was observed at
> 15 ppm after the first exposure. Jackson and Conolly (1985) reported that in male Sprague-
Dawley rats (body weight, 225-275 g) exposed continuously for 4 hours to 0, 200, 250, 300,
375, or 400 ppm, mortality was 0/22, 1/4, 1/16, 3/14, 3/12, and 3/6, respectively. At >250 ppm
there were significant increases (p<0.05) in kidney-to-body weight ratios (approximately 1.4-
fold), serum urea nitrogen (approximately fourfold) and creatinine (approximately threefold).
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Histopathological examination revealed severe tubular necrosis with calcium deposits at >300
ppm.
Using autoradiographic methods, Brittebo et al. (1993) investigated the mechanism of
nephrotoxicity in C57BL6 mice (body weight, 18-22 g) following i.p. injection of 0.4 mg/kg of
14C-labeled 1,1-DCE. Selective covalent binding of radioactivity occurred in the proximal
tubules, in the midzonal parts of the liver lobules, and in the mucosa of the upper and lower
respiratory tract. Treatment with buthionine sulphoximine (BSO), an irreversible inhibitor of
y-glutamylcysteine synthetase and a GSH-depleting agent, caused a threefold increase in
covalent binding of 1,1-DCE. Histopathological examination of kidneys in BSO-pretreated male
mice given single i.p. injections of 25 and 50 mg/kg 1,1-DCE showed necrosis in the proximal
tubules (Sj and S2 segments). In mice given 1,1-DCE only, no significant lesions in the kidneys
were observed. The authors concluded that the severe renal toxicity of 1,1-DCE in BSO-
pretreated mice is related to metabolic activation of 1,1-DCE in the proximal tubules, resulting in
GSH depletion and covalent binding.
The effects in the Clara cells of the lung in mice include extensive histopathological
changes (Forkert and Reynolds, 1982; Forkert et al., 1985, 1990), repair of damage through cell
proliferation (Forkert et al., 1985), depletion of GSH, and covalent binding of 1,1-DCE mediated
through the formation of DCE epoxide by CYP2E1 (Dowsley et al., 1996; Forkert and Mousa,
1991; Forkert, 1999b; Lee and Forkert, 1994; Moussa and Forkert, 1992). No studies are
available showing similar effects in the lungs of rats.
Forkert and Reynolds (1982) investigated the ability of 1,1-DCE administered orally to
induce pulmonary injury. Male C57BL6 mice (three to five per group) were administered a
single dose of 1,1-DCE by gavage in mineral oil at 0, 100, or 200 mg/kg. At 100 mg/kg, Clara
cells showed extensive dilatation of cisternae and degeneration of the endoplasmic reticulum.
The bronchiolar epithelium showed a few vacuolated cells 12 hours after exposure. By 24 hours
the Clara cells showed prominent cytoplasmic vacuoles, but ciliated cells were not affected. By
48 hours, complete recovery had occurred. At 200 mg/kg, both ciliated and Clara cells showed
necrosis of the bronchiolar epithelium. By 24 hours, the lesion had increased in severity and
areas of bronchioles were denuded of epithelium. Peribronchial and perivascular edema,
hemorrhage, and focal atelectasis were also present. Complete recovery occurred by 7 days.
A subsequent study (Forkert et al., 1985) examined regeneration of the damaged
epithelium by cellular proliferation. Male C57BL6 mice were administered a single dose of 1,1-
DCE by gavage in mineral oil at 200 mg/kg followed by a single pulse of 3H-thymidine.
Changes in cellular proliferation were calculated from measurement of radioactivity incorporated
into total pulmonary DNA. Incorporation of radioactivity was significantly inhibited 1 day after
treatment and thereafter increased. The peak incorporation of radioactivity occurred between 3
and 5 days after treatment and returned to baseline by day 7. The majority of the radioactivity
was taken up by the nonciliated bronchiolar epithelial cells.
4.2.2. Longer-Term Exposure
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4.2.2.1. Oral
4.2.2.1.1. Rats. NTP (1982) conducted a 14-day study of 1,1-DCE in male and female F344 rats
(five animals of each sex, 9 weeks old) by gavage in corn oil at 0, 10, 50, 100, 500, or 1,000
mg/kg. Survival was 10/10, 10/10, 10/10, 10/10, 7/10, and 3/10 mg/kg, respectively. Mean
body weight was significantly depressed at >500 mg/kg. Hemorrhagic necrosis in the liver was
observed in all of the rats that died at 500 and 1,000 mg/kg.
In the same study, male and female F344 rats (10 of each sex, 9 weeks old) were
administered 1,1-DCE by gavage in corn oil at 0, 5, 15, 40, 100, or 250 mg/kg five times per
week for 13 weeks. Representative tissues from rats receiving 250 mg/kg and from control rats
were examined microscopically. Livers from all groups were examined. Three female rats
receiving 250 mg/kg died during the first week of the study. No other rats died. Mean body
weight was depressed 13% for male rats receiving 250 mg/kg as compared with controls. Mean
body weight in other groups was comparable. Only the liver showed effects attributed to 1,1-
DCE. At 250 mg/kg, the three female rats that died showed severe centrilobular necrosis.
Minimal to moderate hepatocytomegaly was seen in the rest of the rats at 250 mg/kg. Minimal
to mild hepatocytomegaly was seen in 6/10 male rats and 3/10 female rats that received 100
mg/kg. No biologically significant changes were observed in rats that received 40 mg/kg or less.
The no-observed-adverse-effect level (NOAEL) in this study is 40 mg/kg (equivalent to 28.5
mg/kg-day); the lowest-observed-adverse-effect level (LOAEL) is 100 mg/kg (equivalent to 71.4
mg/kg-day).
4.2.2.1.2. Mice. NTP (1982) conducted a 14-day study in male and female E6C3Fl mice (five
of each sex, 9 weeks old) administered 1,1-DCE by gavage in corn oil at 0, 10, 50, 100, 500, or
1,000 mg/kg. Survival was 10/10 in all groups except the 1000 mg/kg group, where survival was
0/10. Hemorrhagic necrosis in the liver was observed in all mice at 1,000 mg/kg.
In the same study, male and female B6C3FJ mice (10 of each sex, 9 weeks old) were
administered 1,1-DCE by gavage in corn oil at 0, 5, 15, 40, 100, or 250 mg/kg five times per
week for 13 weeks. Representative tissues from mice receiving 100 and 250 mg/kg and from
control mice were examined microscopically. Livers from all groups were also examined.
Survival was 20/20, 19/20, 19/20, 19/20, 15/20, and 1/20 at 0, 5, 15, 40, 100, and 250 mg/kg,
respectively. At 100 mg/kg, there was a decrease in mean body weight in males (14%) but not in
females. No change in mean body weight was observed at lower exposures. Only the liver
showed effects attributed to 1,1-DCE. Centrilobular necrosis of the liver was observed in 5/10
males and 5/10 females that received 250 mg/kg and 2/10 males and 2/10 females that received
100 mg/kg. No biologically significant changes in the liver occurred in mice receiving 40 mg/kg
or below. The NOAEL in this study is 40 mg/kg (adjusted to a continuous daily exposure of
28.6 mg/kg-day); the LOAEL is 100 mg/kg (adjusted to a continuous daily exposure of 71.4
mg/kg-day).
4.2.2.1.3. Dogs. Quasi et al. (1983) conducted a study in beagle dogs (four per group, 8 months
old) administered 1,1-DCE by gavage in peanut oil at 0, 6.25, 12.5, or 25 mg/kg-day for 97 days.
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There were no significant differences among the groups in appearance and demeanor, mortality,
body weight, food consumption, hematology, urinalysis, clinical chemistry determinations, organ
weights, and organ-to-body weight ratios. No exposure-related gross or histopathological
changes were present in tissues. There was no depletion of the nonprotein sulfhydryl levels in
the liver or kidneys. The NOAEL in this study is 25 mg/kg-day (the highest exposure tested).
4.2.2.2. Inhalation
Gage (1970) exposed four male and four female Alderly Park rats (body weight 200 g) to
200 ppm or 500 ppm 1,1-DCE 6 hrs/day for 20 days. At 200 ppm there was slight nasal
irritation (not further described). At necropsy all organs appeared normal. At 500 ppm there
was nasal irritation (not further specified), retarded weight gain (data not reported), and liver cell
degeneration (not further defined).
Plummer et al. (1990) exposed black hooded Wistar rats to 50 ppm 1,1-DCE (18 males
and 18 females, age not specified) continuously for 4 weeks (except for two 1.5-hour periods per
week) or to 250 ppm (six males and six females, age not specified) for 6 hrs/day, 5 days/wk for 4
weeks. The total exposure (concentration x time) was the same for the two profiles
(33,533 ppm/hr for the continuous exposure and 32,200 ppm/hr for the intermittent exposure).
Rats in the intermittent exposure group showed signs of early coagulative necrosis in the liver
(incidence not reported). Eleven of the 12 rats in the continuous-exposure group showed less
severe injury, including fatty changes in variable numbers of hepatocytes and only very
occasional focal liver cell necrosis. The LOAEL in this study is 50 ppm.
Prendergast et al. (1967) evaluated the toxicity of 1,1-DCE in Long-Evans and Sprague-
Dawley rats, Hartley guinea pigs, beagle dogs, New Zealand albino rabbits, and squirrel
monkeys. One set of test animals (15 rats/group, 15 guinea pigs/group, 3 rabbits/group,
2 dogs/group, or 3 monkeys/group) was exposed to 1,1-DCE vapors for 8 hrs/day, 5 days/wk, for
a total of 30 exposures at 395 ± 32 mg/m3. The age of the animals was not specified. The
exposed animals were evaluated for visible signs of toxicity, mortality, and hematologic,
biochemical, pathologic, and body weight changes. In this study there were no deaths, no visible
signs of toxicity, and no histopathological changes. The NOAEL in this study is 395 mg/m3 (the
highest exposure tested), equivalent to an adjusted NOAEL based on continuous exposure of 94
mg/m3.
Another set of test animals (15 rats/group, 15 guinea pigs/group, 3 rabbits/group,
2 dogs/group, or 3 or 9 monkeys/group) was exposed continuously for 90 days to 1,1-DCE
vapors at 189 ± 6.2, 101 ± 4.4, 61 ± 5.7, or 20 ± 2.1 mg/m3. The concurrent controls included
304 rats, 314 guinea pigs, 48 rabbits, 34 dogs, and 57 monkeys. The age of the animals was not
specified. The exposed animals were evaluated for visible signs of toxicity, mortality, and
hematologic, biochemical, pathologic, and body weight changes. There was apparent exposure-
related mortality in guinea pigs and monkeys. In the 0, 20, 61, 101, or 189 mg/m3 exposure
groups, guinea pig mortality was 2/314, 2/45, 3/15, 3/15, and 7/15, and monkey mortality was
1/57, 1/21, 0/9, 2/3, and 3/9, respectively. The guinea pigs died between days 3 to 9 of exposure;
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the monkeys died on days 26, 39, 47, 60, and 64 of exposure. There were no visible signs of
toxicity in any surviving animals. At the highest exposure in monkeys, but not in guinea pigs,
there was some histopathological evidence of liver damage (see below). In guinea pigs at the
highest exposure, there was an increase in serum glutamic-pyruvic transaminase and liver
alkaline transaminase (see below). Because visible signs of toxicity were not observed, and only
minor liver damage was apparent in this study, the mortality data in guinea pigs and monkeys are
given no weight.
Varying degrees of growth depression were found in all exposures, but were significant
in all species only at 189 mg/m3. The test animals exhibited no significant hematologic
alterations, and serum urea nitrogen levels were within control limits in all exposures in which
determinations were made. Significant elevations of serum glutamic-pyruvic transaminase and
liver alkaline phosphatase activities were found in rats (a threefold and 1.75-fold increase,
respectively) and guinea pigs (a sevenfold and 2.4-fold increase, respectively) exposed to 189
mg/m3 (other species not tested) but not at 20 mg/m3 (enzyme levels at intermediate exposures
not tested). Histopathological examination of liver from dogs, monkeys, and rats revealed
damage at 189 mg/m3 (other species not examined). The effects observed included fatty
metamorphosis, focal necrosis, hemosiderosis deposition, lymphocytic infiltration, bile duct
proliferation, and fibrosis. The changes were most severe in dogs. Sections of kidney from all
rats showed nuclear hypertrophy of the tubular epithelium. No detectable liver or kidney
damage was observed in any species exposed to 101 mg/m3 or less. The NOAEL in this study is
101 mg/m3 (equivalent to 25 ppm); the LOAEL is 189 mg/m3 (equivalent to 47 ppm).
4.2.3. Chronic Studies and Cancer Bioassays
4.2.3.1. Oral
4.2.3.1.1. Rats. Ponomarkov and Tomatis (1980) treated 24 female BD IV rats by gavage with
1,1-DCE dissolved in olive oil (150 mg/kg body weight) on gestation day (GD) 17. Their
offspring (81 males and 80 females) were treated weekly with 1,1-DCE at 50 mg/kg body weight
by gavage in olive oil from the time of weaning for 120 weeks or until the animal was moribund.
A control group of offspring (49 males and 47 females) received only olive oil. Liver and
meningeal tumors were more frequently observed in treated than in untreated animals, but the
difference was not statistically significant. The total number of tumor-bearing animals was not
statistically different between the treated and untreated groups.
NTP (1982) conducted chronic toxicity and carcinogenicity studies for 104 weeks in
male and female F344 rats (50 of each sex in each group, 9 weeks old) by gavage in corn oil at 0,
1, or 5 mg 1,1-DCE/kg-day. There were no significant differences in survival, clinical signs, or
body weight between test animals and controls for any group, suggesting that the maximum
tolerated dose was not achieved. The results of histopathological examination indicated chronic
renal inflammation in male rats (26/50, 24/48, 43/48) and female rats (3/49, 6/49, 9/44). The
increase was statistically significant only in males at the highest exposure. As this lesion
commonly occurs in aged male albino rats (Kluwe et al., 1984, Kluwe, 1990), it is not
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considered to be biologically significant in this study. All of the increased tumor incidences that
were statistically significant by the Fisher exact test or by the Cochran-Armitage linear trend test
(adrenal pheochromocytoma, pancreatic islet cell adenoma or carcinoma, and subcutaneous
fibroma in males and pituitary adenoma in females) were not significant when life-table analyses
were used. This difference occurs because life table analyses adjust for intercurrent mortality
and thus minimize the impact of animals dying before the onset of late-appearing tumor. This
adjustment was particularly critical for the analyses of tumor incidences in male rats because 12
controls and 10 low-dose animals were accidentally killed during week 82 of the study.
Accordingly, NTP concluded that no increased incidence of tumors was found at any site in
these bioassays. Under the conditions of this bioassay, 1,1-DCE administered by gavage was not
carcinogenic for F344 rats. The NOAEL in this study is 5 mg/kg-day (the highest exposure
tested).
Quast et al. (1983) conducted a 2-year chronic toxicity and carcinogenicity study of 1,1-
DCE in Sprague-Dawley rats (6-7 weeks old). There were 80 rats of each sex in the control
group and 48 rats of each sex in each exposed group. The 1,1-DCE was incorporated in the
drinking water of the rats at nominal concentrations of 0, 50, 100, or 200 ppm. The time-
weighted average exposure over the 2-year period was 7, 10, or 20 mg/kg-day for males and 9,
14, or 30 mg/kg-day for females. Rampy et al. (1977) also reported some of the data; Humiston
et al. (1978) reported more detailed data. No significant differences were found between the
groups in appearance and demeanor, mortality, body weight, food consumption, water
consumption, hematology, urinalysis, clinical chemistry determinations, organ weights, or organ-
to-body weight ratios. After 1 year of study, no depletion of the nonprotein sulfhydryl levels in
the liver or the kidneys was observed (Rampy et al., 1977).
The only treatment-related effect observed was minimal hepatocellular midzonal fatty
change and hepatocellular swelling. At the termination of the study, male rats showed an
increased incidence of minimal heptocellular fatty change (control, 14/80; 50 ppm, 5/48; 100
ppm, 13/48; 200 ppm, 19/47) and minimal hepatocellular swelling (control, 0/80; 50 ppm, 1/48;
100 ppm, 2/48; 200 ppm, 3/47). The changes were statistically significant (p<0.05) only in the
200 ppm group. Female rats also showed an increased incidence of minimal hepatocellular fatty
change (control, 10/80; 50 ppm, 12/48; 100 ppm, 14/48; 200 ppm, 22/48; statistically significant
[p<0.05] at 100 and 200 ppm) and minimal hepatocellular swelling (control, 3/80; 50 ppm, 7/48;
100 ppm, 11/48; 200 ppm, 20/48; statistically significant [p<0.05] in all groups). No exposure-
related neoplastic changes occurred at any exposure. No hepatocellular necrosis was evident at
any exposure.
On the basis of the minimal nature of the hepatocellular swelling reported by the authors
and no change in liver weight, no change in clinical chemistry measurements diagnostic for liver
damage, and no other indication of abnormal liver function, the hepatocellular swelling is not
considered to be biologically significant or an adverse effect in this study. The statistically
significant hepatocellular midzonal fatty change, however, is considered a minimal adverse
effect in this study. Accordingly, the NOAEL in male rats is 10 mg/kg-day and the LOAEL is
20 mg/kg-day; the NOAEL in female rats is 9 mg/kg-day and the LOAEL is 14 mg/kg-day. A
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benchmark dose (BMD) analysis was conducted for the results in female rats (Appendix B). In
female rats, the BMD10 (the dose that gives a 10% response) is 6.6 mg/kg-day and the BMDL10
(the lower 95% confidence limit on the BMD10) is 4.6 mg/kg-day.
Maltoni et al. (1985) conducted a carcinogenicity and toxicity study of 1,1-DCE in
Sprague-Dawley rats. Animals (9 or 10 weeks old) were exposed by gavage in olive oil to 0,
0.5, 5, 10, or 20 mg/kg, 4-5 days/wk for 52 weeks. There were two control groups, one with 150
animals (75 of each sex) and the other with 200 animals (100 of each sex). The exposed groups
had 100 animals (50 of each sex). Following the 52-week exposure, animals were observed until
spontaneous death (total duration 147 weeks). Body weight was measured every 2 weeks during
the 52-week exposure and every 8 weeks thereafter. Full necropsy and histopathological
examination were performed. No biologically significant changes were observed in mortality or
body weight, and no biologically significant noncancer or cancer effects were found in any
organ.
4.2.3.1.2. Mice. NTP (1982) conducted chronic toxicity and carcinogenicity studies for 104
weeks of 1,1-DCE in male and female B6C3Fj mice (50 of each sex in each group, 9 weeks old)
by gavage in corn oil at 0, 2, or 10 mg/kg. No significant differences were observed in survival,
clinical signs, or body weight in any group, and there was evidence of only slight toxicity in the
liver, suggesting that the maximum tolerated dose was not achieved. The only noncancer effect
observed by histopathological examination was necrosis of the liver (male: 1/46, 3/46, 7/49;
female: 0/47, 4/49, 1/49). The effect was not statistically significant at either exposure (p = 0.6
and 0.06 at the mid- and high-exposure levels in males using a two-tailed test, respectively). The
only observed significant increase (p<0.05) in tumor incidence occurred in low-dose females for
lymphoma (2/48, 9/49, 6/50) and for lymphoma or leukemia (7/48, 15/49, 7/50). These increases
were not considered to be related to 1,1-DCE administration because similar effects were not
found in the high-dose females or in males. Under the conditions of this bioassay, 1,1-DCE
administered by gavage was not carcinogenic for B6C3Fj mice. In male and female mice the
NOAEL is 10 mg/kg-day (the highest exposure tested). The BMD10 is 7.8 mg/kg-day and the
BMDL10is4.1 mg/kg-day.
4.2.3.1.3. Trout. Hendricks et al. (1995) conducted an 18-month carcinogenicity study of 1,1-
DCE in rainbow trout (8 weeks old) at 4 mg/kg-day. Tissues examined for neoplasms included
liver, kidney, spleen, gill, gonads, thymus, thyroid, heart, stomach, pyloric ceca, duodenum,
rectum, pancreas, and swimbladder. 1,1-DCE produced no neoplasms at the exposure levels
used and no increase in liver weight. There was no evidence of any other chronic toxic effects.
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4.2.3.2. Inhalation
4.2.3.2.1. Rats. Lee et al. (1977, 1978) exposed 2-month-old Charles River CD rats (36 males
and 35 females) to 55 ppm 1,1 -DCE for 6 hrs/day, 5 days/wk for 12 months. No significant
changes were observed in survival, body weight, hematology, clinical blood chemistry,
pulmonary macrophage count, cytogenetic analysis of bone marrow, x-ray examination of
extremities, collagen contents in liver and lung, serum aminolevulinic acid (ALA) synthetase,
urinary ALA level, or serum alpha-fetoprotein. A mild to markedly severe focal, disseminated
vacuolization was observed in livers of most of the rats. No hemangiosarcomas were found in
the liver or lung. The incidence of hemangiosarcomas in mesenteric lymph node or
subcutaneous tissue was 2/36 in males and 0/35 in females.
Viola and Caputo (1977) exposed 2-month-old Sprague-Dawley rats (30 males and 30
females per group) to 0, 75, or 100 ppm 1,1-DCE for 22-24 months (hours of daily exposure not
reported). The incidence of tumors observed at necropsy (males and females combined) was
15/60, 10/36, and 20/60, respectively. The tumors observed were classified as subcutaneous
fibromas or abdominal lymphomas. The histopathological results from this study have not been
published. No other data were reported.
In the same study, 2-month-old albino Wistar rats (37 males and 37 females) were
exposed to 1,1-DCE for 4 hrs/day, 5 days/wk for 12 months. Exposures were 200 ppm for the
first 6 months and 100 ppm for the rest of the study. A control group of 30 males and 30 females
received air only. The incidence of tumors (described as reticulum cell sarcomas of a
nonsincytial type, primarily in the abdominal cavity) was 15/60 and 17/74 in control and
exposed rats, respectively. No other data were reported.
Hong et al. (1981) evaluated mortality and tumor incidence in groups of 2-month-old CD
rats of both sexes exposed to 0 or 55 ppm 1,1-DCE 6 hrs/day, 4 days/wk for 1 month (4 of each
sex), 3 months (4 of each sex), 6 months (4 of each sex), or 10 months (16 of each sex).
Following exposure, all animals were observed for an additional 12 months. In rats exposed for
10 months, there was an increase in mortality following the 12-month observation period (67%
in exposed; 41% in controls). There was no significant increase in tumors at any site for any
exposure period.
Maltoni et al. (1985) conducted a carcinogenicity and toxicity study of 1,1-DCE in
Sprague-Dawley rats. Animals (16 weeks old) were exposed by inhalation to 0, 10, 25, 50, 100,
or 150 ppm for 4 hrs/day, 4-5 days/wk for 52 weeks. The control group had 200 animals (100 of
each sex); the 10, 25, 50, and 100 ppm groups had 60 animals (30 of each sex), and the 150 ppm
group had 120 animals (60 of each sex). Following the 52-week exposure, animals were
observed until spontaneous death (total duration 137 weeks). Body weight was measured every
2 weeks during the 52-week exposure and every 8 weeks thereafter. Full necropsy and
histopathological examination were performed. No biologically significant changes in mortality
or body weight were observed, and there were no biologically significant noncancer effects in
any organ in either sex or an increase in tumors in males at any site. There was a statistically
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significant increase (p<0.05) in each treatment group as compared with the control group in the
number of females with mammary fibromas and fibroadenomas. The incidence was 44/56
(78.6%), 24/24 (100%), 20/20 (100%), 21/22 (95.4%), 21/23 (91.3%), and 38/43 (88.4%) in the
control, 10, 25, 50, 100, and 150 ppm groups, respectively. The latency time and the number of
tumors per tumor-bearing animal were similar among all groups. The incidence of mammary
carcinoma in the exposed groups was consistently less than that of controls—16/56 (28.6%),
5/24 (20.8%), 4/20 (20%), 1/21 (4.5%), 3/21 (13.0%), and 9/38 (20.9%) in the control, 10, 25,
50, 100, and 150 ppm groups, respectively. This study provides no evidence that 1,1-DCE is
carcinogenic in male and female Sprague-Dawley rats.
Quast et al. (1986) and Rampy et al. (1977) reported results from studies in which male
and female Sprague-Dawley rats (Spartan substrain, 86 animals/group) were exposed to 1,1-DCE
by inhalation 6 hrs/day, 5 days/wk for up to 18 months. Interim sacrifices occurred at 1, 6, and
12 months. Rats were exposed to 1,1-DCE concentrations of 10 ppm and 40 ppm for the first 5
weeks of the study. Because of the absence of observable treatment-related effects among rats
sacrificed after 1 month of exposure, the concentrations were increased to 25 and 75 ppm.
Exposures were continued at these concentrations through the 18th month of the study. The
surviving animals were then held without exposure to 1,1-DCE until 24 months. Cytogenetic
evaluations were performed on a separate group of animals (four/sex) exposed to 0, 25, or 75
ppm for 6 months.
A separate 90-day study using 20 rats/sex/treatment group was conducted at 0, 25, and 75
ppm, with an interim sacrifice of 8 rats/group at 30 days. No exposure-related changes in
mortality, appearance and demeanor, body weight, clinical chemistry determinations,
hematologic evaluations, urinalysis, or cytogenetic evaluation of bone marrow preparations were
observed. Minimal hepatocellular fatty change in the midzonal region of the hepatic lobule was
observed in both male and female rats in the 25 ppm and 75 ppm groups at the 6-month interim
sacrifice (male: control, 0/5; 25 ppm, 1/5; 75 ppm, 4/5; female: control, 0/5; 25 ppm, 2/5; 75
ppm, 4/5). The fatty change was also observed at the 12-month sacrifice, but there was no
indication of progression of severity (male: control, 0/5; 25 ppm, 3/5; 75 ppm, 5/5; female:
control, 0/5; 25 ppm, 5/5; 75 ppm, 5/5). At the 18-month sacrifice the incidence of this change
was no longer increased in male rats (control, 0/27; 25 ppm, 0/25; 75 ppm, 1/27). However, the
change persisted in female rats (control, 0/16; 25 ppm, 6/29; 75 ppm, 7/20). In female rats the
fatty change was statistically significant (p<0.05) only at the higher exposures. During the last 6
months of the study, after exposure had been discontinued, this effect was no longer discernible
(male: control, 0/46; 25 ppm, 1/47; 75 ppm, 0/51; female: control, 0/49; 25 ppm, 0/46; 75 ppm,
1/48).
Although the incidence of several tumors and/or tumor types was found to be statistically
increased or decreased as compared to controls, none of these differences were judged to be
attributable to 1,1-DCE. The tumor incidence data for both control and treated rats in this study
was comparable to historical control data for the Sprague-Dawley rats (Spartan substrain) used
by this laboratory for several studies of similar design and duration. Although the minimal
hepatocellular midzonal fatty change is reversible, did not result in altered organ weight, clinical
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chemistry changes diagnostic for liver damage, or any obvious decrement in liver function, the
fatty change in liver is considered a minimal adverse effect. Accordingly, the NOAEL in male
rats in this study is 75 ppm (the highest exposure tested). The NOAEL for female rats is 25
ppm; the LOAEL is 75 ppm. A BMD analysis was conducted (Appendix B). In female rats the
BMC10 (the concentration that gives a 10% response) is 15.1 ppm and the BMCL10 (the lower
95% confidence limit on BMC10) is 9.8 ppm, equivalent to 1.8 ppm adjusted for continuous
exposure (9.8 ppm x 6/24 x 5/7).
Cotti et al. (1988) exposed Sprague-Dawley rats to 1,1-DCE at 0 or 100 ppm for 4-7
hrs/day, 5 days/wk. The exposures were to 13-week-old females for 104 weeks (60 control
animals and 54 exposed animals) and to the offspring of pregnant rats exposed from GD 12 and
for 15 or 104 weeks after birth (158 males and 149 females as controls, 60 males and 60 females
exposed for 15 weeks, and 62 males and 61 females exposed for 104 weeks). Animals were
observed until spontaneous death. In males and females exposed for 104 weeks and in male
offspring exposed for 15 weeks, a slight decrease in body weight was observed (data not
reported). An increased percentage of rats bearing malignant tumors (30.9 vs. 17.3% in controls)
and an increased number of malignant tumors per 100 animals (34.1 vs. 17.9% in controls) were
observed in male and female offspring exposed for 104 weeks (statistical analysis not presented).
An increase in leukemia that appeared to be related to length of exposure was also observed in
offspring (4.2% for controls, and 8.3% and 11.4% for exposure of 15 and 104 weeks,
respectively). Tumors at other sites (total benign and malignant tumors, total benign and
malignant mammary tumors, malignant mammary tumors, pheochromocytomas) showed no
change or a decreased incidence. Data from this study are also reported in Maltoni et al. (1985).
4.2.3.2.2. Mice. Lee et al. (1977, 1978) exposed 2-month-old CD-I mice (18 males and 18
females) to 0 or 55 ppm 1,1-DCE for 6 hrs/day, 5 days/wk, for up to 12 months. No deaths
occurred in the control or exposed groups. Weight gain was comparable between groups. No
changes in hematology, clinical blood chemistry, cytogenetic analysis of bone marrow, x-ray
examination of extremities, or serum alpha-fetoprotein were observed. The livers showed no
increase in mitotic figures using 14C-thymidine incorporation. Animals exposed for 6 to 12
months had several changes in the liver, including enlarged and basophilic hepatocytes with
enlarged nuclei, mitotic figures or polyploidy, microfoci of mononuclear cells, focal
degeneration, and necrosis. The incidence and severity of these lesions progressed with length
of exposure (data not reported). The incidence of bronchioalveolar adenoma (males and females
combined) for 1-3 months, 4-6 months, 7-9 months, and 10-12 months of exposure was 0/24,
1/8, 2/10, and 3/28, respectively. The incidence of hemangiosarcomas in liver (males and
females combined) for 6 months, 7-9 months, and 10-12 months of exposure was 0/16, 1/10,
and 2/28, respectively. No hemangiosarcomas were found in other tissues.
Hong et al. (1981) evaluated mortality and tumor incidence rates in mice exposed to 1,1-
DCE. Groups of 2-month-old albino CD-I mice of both sexes were exposed to 0 or 55 ppm for 6
hrs/day, 4 days/wk for 1 month (8 of each sex), 3 months (8 of each sex), or 6 months (12 of
each sex). Following exposure, all animals were observed for an additional 12 months. In mice
exposed for 6 months there was a slight increase in mortality following the 12-month
19
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observation period (46% in exposed, 39% in controls). There was no significant increase in
tumors at any site for any exposure period.
Maltoni et al. (1985) conducted a carcinogenicity and toxicity study of 1,1-DCE in Swiss
mice. Animals (9 or 16 weeks old) were exposed by inhalation to 0, 10, or 25 ppm for 4 hrs/day,
4-5 days/wk, for 52 weeks. Groups of animals exposed to >50 ppm showed extreme toxicity
after only a few exposures, causing termination of this portion of the bioassay. There were two
control groups, one with 180 animals (90 of each sex) and the other with 200 animals (100 of
each sex). The 10 ppm group had 60 animals (30 of each sex). Two groups were exposed to
25 ppm: one group consisted of 60 animals (30 of each sex) and the other of 240 animals (120 of
each sex). Following the 52-week exposure, animals were observed until spontaneous death
(total duration 126 weeks). Body weight was measured every 2 weeks during the 52-week
exposure and every 8 weeks thereafter. Full necropsy and histopathological examination were
performed.
No biologically significant changes in body weight were seen. The exposed animals had
a somewhat higher survival than did controls. No biologically significant noncancer effects
where observed in any organ, except for a marginal increase in regressive changes in the kidney
(presumably necrosis and proliferation of the cortical tubules) and a marginal increase in kidney
abscesses and nephritis. In males the incidence of regressive changes was 103/190 (54%), 23/30
(77%), and 102/150 (68%), and the incidence of kidney abscesses and nephritis was 45/190
(24%), 13/30 (43%), and 58/150 (39%) in the control, 10 ppm, and 25 ppm exposure groups,
respectively. The results in male mice were statistically significant (p<0.05) for both effects at
both exposures. In females the incidence of regressive changes was 93/190 (49%), 19/30 (63%),
and 97/150 (65%), and the incidence of kidney abscesses and nephritis was 52/190 (27%), 8/30
(27%), and 50/150 (33%) in the control, 10 ppm, and 25 ppm exposure groups, respectively.
The results in female mice were statistically significant (p<0.05) only for regressive changes at
the higher exposure. There was a statistically significant increase (p<0.01) over controls in
kidney adenocarcinomas in male mice at 25 ppm, but not in male mice at 10 ppm or in female
mice at either exposure. The incidence was 0/126 (0%), 0/25 (0%), and 28/119 (23.5%) in male
mice in the combined control, 10 ppm, and combined 25 ppm groups, respectively.
A statistically significant increase (p<0.01) over controls was seen in mammary
carcinomas in female mice at both exposures, but there was no clear exposure-response
relationship. The incidence was 3/185 (1.6%), 6/30 (20%), and 16/148 (11%) in females in the
combined control, 10 ppm, and combined 25 ppm groups, respectively. There was also a
statistically significant increase (p<0.01) over controls in pulmonary adenomas in both exposed
groups, but there was no clear exposure-response relationship. The incidence was 12/331
(3.6%), 14/58 (24.1%), and 41/288 (14.2%) in male and female mice combined in the combined
control, 10 ppm, and combined 25 ppm groups, respectively. No pulmonary carcinomas were
observed in any mice. The incidence data are reported as the number of tumor-bearing animals
compared to the number of animals alive when the first tumor was observed in that organ
(kidney adenocarcinoma, 55 weeks; mammary tumor, 27 weeks; pulmonary adenoma, 36
weeks).
20
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4.2.3.2.3. Hamsters. Maltoni et al. (1985) conducted a carcinogenicity and toxicity study of
1,1-DCE in Chinese hamsters. Animals (28 weeks old) were exposed by inhalation to 0 or 25
ppm for 4 hrs/day, 4-5 days/wk for 52 weeks. The control group had 35 animals (18 male and
17 female); the 25-ppm group had 60 animals (30 of each sex). Following the 52-week
exposure, animals were observed until spontaneous death (total duration 157 weeks). Body
weight was measured every 2 weeks during the 52-week exposure and every 8 weeks thereafter.
Full necropsy and histopathological examination were performed. No biologically significant
changes were seen in mortality or body weight, and there were no biologically significant
noncancer or tumor effects in any organ.
4.2.3.3. Dermal
Van Duuren et al. (1979) evaluated the carcinogenicity of 1,1-DCE in male and female
non-inbred Ha:ICR Swiss mice. Carcinogenicity was assessed in three types of tests: a dermal
initiation-promotion assay, a repeated dermal application assay, and a subcutaneous injection
assay. Vehicle, no-treatment, and positive control groups were included in the tests. In the
initiation-promotion assay, 1,1-DCE was tested as a tumor-initiating agent with phorbol
myristate acetate as the promoter. Thirty female mice were treated with 121 mg 1,1-DCE. A
significant increase (p<0.005) was observed in skin papillomas (nine in eight mice). In the
repeated dermal application assay, exposures of 40 and 121 mg/mouse were used. 1,1-DCE was
applied to the back of the shaved animals (30 females/dose). No sarcomas were observed at the
site of treatment. No statistically significant increase in tumors was observed at any site remote
from the site of treatment. In the subcutaneous injection assay, the test animals were given
weekly injections of 2 mg of 1,1-DCE. After 548 days on test, none of the animals had
developed sarcomas at the injection site. 1,1-DCE showed initiating activity in the two-stage
carcinogenesis experiments but was inactive as a whole-mouse dermal carcinogen and after
subcutaneous injection.
4.3. REPRODUCTIVE AND DEVELOPMENTAL STUDIES—ORAL AND
INHALATION
4.3.1. Direct Infusion
Dawson et al. (1990) conducted studies in Sprague-Dawley rats using direct infusion of a
solution of 1.5 or 150 ppm 1,1-DCE to the gravid uterus during the period of organ
differentiation and development. The delivery rate of the test solution was 0.5 |j,L/hour
beginning at GD 7 and continuing for 2 weeks. On GD 22 the pregnant rats were killed and the
gravid uterus was removed for examination. The only effect noted was an increase in a variety
of congenital heart changes (atrial septal, pulmonary valve, aortic valve, and membranous
ventricular septal changes). The incidence of total cardiac changes was 3% in the control group
and 12.5% and 21% in the 1.5 and 150 ppm groups, respectively. The increase was statistically
significant (p<0.05) at both exposures; however, the statistical analysis was based on total
occurrence, not on numbers of litters affected or fetuses per litter affected.
21
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Goldberg et al. (1992) conducted studies on chick embryos to determine whether 1,1-
DCE was a cardiac teratogen. On day 3 of incubation, fertilized White Leghorn chick eggs (N =
418) were inoculated just above the embryo with 30 |j,L of a test solution of 1,1-DCE in mineral
oil at 5 |iM (N = 76), 20 |j,M (N = 62), or 25 |j,M (N = 76). Two control groups were also tested
using normal saline (N = 96) or mineral oil (N = 108). Chicks were terminated on day 18 of
incubation. No change was seen in mortality among groups. Cardiac changes included atrial
and ventricular septal changes, malformations of all valves, and great vessel changes. Cardiac
and great vessel changes occurred in 4% of each of the two control groups and in 17, 19, and 2%
of the low-, mid-, and high-dose groups, respectively.
4.3.2. Oral
Nitschke et al. (1983) evaluated the reproductive and developmental toxicity of 1,1-DCE
in Sprague-Dawley rats. Three generations of the test animals were exposed to drinking water
containing nominal 1,1-DCE concentrations of 0 (initially 15 males and 30 females), 50, 100, or
200 ppm (initially 10 males and 20 females at each exposure). The authors provided no
information on water consumption. This study was a companion study to Quast et al. (1983) and
used the same concentrations of 1,1-DCE in drinking water. In the Quast et al. study the average
exposure to females was 9, 14, or 30 mg/kg-day. After 100 days of exposure, the rats were
mated. In the Nitschke et al. three-generation study, there were no biologically significant
changes in fertility index, in average number of pups per litter, in average body weight of pups,
or in pup survival at any exposure. Neonatal survival was decreased from concurrent control
values in the f2 and f3a litters of dams ingesting 1,1-DCE from drinking water. The survival
indices, however, were within the range of control values for this strain of rats in this laboratory.
The authors attributed the decreased survival index in f2 to increased litter size at birth in dams
exposed to 1,1-DCE. The apparent effect seen in the f3a litters was not repeated in subsequent
matings of the same adults to produce either the f3b or the f3c litters. The authors attributed the
decreased survival in the f3a litters as being due to chance.
Histopathological examination of tissues of rats exposed to 1,1-DCE in the drinking
water in utero, during lactation, and postweaning revealed slight hepatocellular fatty change and
an accentuated hepatic lobular pattern of a reversible nature in the adult rats (data not reported,
but the observation is consistent with that reported by Quast et al. [1983] in a chronic bioassay).
These effects were observed in the 100 and 200 ppm groups in the Fx generation and in all
groups of the F2 generation. The authors did not present incidence data and did not report
statistical analysis. Exposure to 1,1-DCE in drinking water at concentrations causing mild, dose-
related changes in the liver did not affect the reproductive capacity of rats through three
generations that produced six sets of litters. The NOAEL for reproductive and developmental
toxicity in this study is 200 ppm for exposure to 1,1-DCE in drinking water (the highest
exposure tested and about 30 mg/kg-day).
Murray et al. (1979) evaluated the developmental toxicity of 1,1-DCE administered in
drinking water at 0 (27 animals) or 200 ppm (26 animals) to pregnant Sprague-Dawley rats
(body weight 250 g). Rats were exposed on GDs 6-15 at 40 mg/kg-day. Using standard
22
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techniques for soft and hard tissue examination, no teratogenic effects were seen in the embryos,
and there was no evidence of toxicity to the dams or their offspring. The NOAEL for
developmental toxicity in this study is 40 mg/kg-day (the highest exposure tested).
Dawson et al. (1993) evaluated the ability of 1,1-DCE administered in drinking water at
110 ppm or 0.15 ppm to female Sprague-Dawley rats (body weight 250 g) to induce fetal cardiac
changes. Rats were administered 110 ppm 1,1-DCE for 61 days before mating or for 48 days
before mating and for 20 days during gestation. Other rats were administered 0.15 ppm 1,1-DCE
for 82 days before mating or for 56 days before mating and for 20 days during gestation. The
dams were killed on GD 22 and the gravid uterus was removed and examined. No effect was
seen on maternal weight gain, average resorption sites (sites where development began but
resorption later occurred), or average implantation sites (sites that did not appear to develop
beyond implantation and contained a metrial gland only). There was no increase in the incidence
of cardiac changes when dams were exposed only before mating. There was, however, a
statistically significant increase (p<0.01) in the percent of fetuses with cardiac changes (atrial
septal, mitral valve, and aortic valve changes) when the dams were exposed before mating and
during gestation. The incidence was control, 7/232 (3%); 0.15 ppm, 14/121 (12%); and 110
ppm, 24/184 (13%). This statistical analysis was based on total occurrence of affected fetuses.
Because the exposure was to the dam and not to individual fetuses, a nested statistical analysis is
preferred. Such an analysis takes into account the correlation among fetuses within a litter and
the possible nesting of effects within litters. This analysis has not been conducted because all
the necessary data are not available.
The author provided additional data (letter from B. Dawson, University of Auckland,
New Zealand, to R. Benson, U.S. EPA, January 24, 2001) to resolve typographical errors in the
exposure information for each group and to clarify the number of affected litters and number of
fetuses per litter affected. The exposure to dams before and during pregnancy was 0, 0.02, or 18
mg/kg-day in the control, 0.15 ppm, and 110 ppm groups, respectively. The number of affected
litters was 5/21 (24%), 8/11 (73%), and 13/17 (76%). The mean number of affected fetuses per
litter for affected litters only was 1.40 (13% of the fetuses in the litter), 1.75 (16% of the fetuses
in the litter), and 1.85 (17% of the fetuses in the litter). The mean number of affected fetuses per
litter all litters was 0.33 (3% of the fetuses in the litter), 1.27 (12% of the fetuses in the litter),
and 1.41 (13% of the fetuses in the litter).
These investigators did a much more thorough evaluation of alterations in cardiac
development than is done in standard developmental toxicity testing protocols. There is no
experience with the background rates or the functional significance of such alterations from
other studies or laboratories. The incidence of alterations in control fetuses (3% of all fetuses,
24% of all litters, and 1.40 affected fetuses per affected litter) suggests a high background
incidence. The authors report that examinations were done blind to the treatment group, so the
data are presumed to be unaffected by observer bias.
No demonstrated exposure-response relationship was found in the Dawson et al. (1993)
study. A 900-fold increase in exposure did not produce a significant increase in response in any
23
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measure of effect. The observed cardiac changes are of questionable biological significance, as
there were no biologically significant effects reported on growth and survival in the three-
generation study (Nitschke et al., 1983). No cardiac effects were reported in a prenatal
developmental study (Murray et al., 1979); however, in this study exposure to 1,1-DCE did not
occur throughout pregnancy. The pharmacokinetics of 1,1-DCE make it biologically implausible
that the observed cardiac changes were causally associated with exposure to 1,1-DCE. The
exposures used in Dawson et al. (1993) were below the level of saturation of CYP2E1 in the rat
liver. Essentially all of the 1,1-DCE administered to the dams would have been metabolized in
the liver and would have reacted with GSH or macromolecules in the liver. See the discussion
and references in section 3. Therefore, it is extremely unlikely that any significant amount of
1,1-DCE or any toxic metabolite would have been present in the fetal compartment. CYP2E1 is
not expressed in fetal liver but begins to be expressed shortly after birth (Cresteil, 1998). EPA is
not aware of any information on the expression of CYP2E1 in fetal cardiac tissue. Cardiac
tissue, however, is not generally considered to be a tissue with significant potential for
metabolism of xenobiotics. For these reasons EPA cannot conclude that the observed cardiac
changes were caused by exposure to 1,1-DCE.
4.3.3. Inhalation
Short et al. (1977b) evaluated developmental toxicity of 1,1-DCE administered by
inhalation to pregnant CD-I rats (Charles River). Animals were exposed to 0 (58 animals), 15
ppm (18 animals), 57 ppm (20 animals), 300 ppm (18 animals), or 449 ppm (18 animals) for
22-23 hrs/day on GDs 6-16. Dams were sacrificed on GD 20. Maternal toxicity was indicated
by severe maternal weight loss (> 28 g/dam) at >15 ppm and by maternal mortality at >57 ppm.
There was a statistically significant increase in the mean number of fetuses per litter, with
hydrocephalus at 15 and 57 ppm, malaligned sternebrae at 15 ppm, and unossified sternebrae at
57 ppm. Because of the severe maternal toxicity at > 15 ppm (> 60 mg/m3), this study is not
useful for evaluating developmental toxicity.
In the same study, pregnant CD-I mice (Charles River) were exposed by inhalation to
1,1-DCE at 0 (65 animals), 15 ppm (23 animals), 30 ppm (19 animals), 57 ppm (21 animals), 144
ppm (18 animals), or 300 ppm (15 animals) for 22-23 hrs/day on GDs 6-16. Dams were
sacrificed on GD 17. Maternal toxicity occurred at > 30 ppm, as shown by statistically
significant decreases in maternal weight gain. At 144 and 300 ppm there was an increase in
maternal mortality. At 30 ppm and higher there was severe fetal toxicity with complete early
resorption of the litters. At 15 ppm there was no evidence of maternal toxicity, no decrease in
fetal body weight, and no decrease in the percentage of viable fetuses. At 15 ppm, there was an
increase in the mean number of fetuses per litter with hydrocephalus, occluded nasal passages,
microphthalmia, cleft palate, small liver, and hydronephrosis. None of these changes, however,
were statistically significant when compared to controls. Also at 15 ppm there was a statistically
significant increase in the mean number of fetuses with an unossified incus and with
incompletely ossified sternebrae. This study provides evidence of fetal toxicity at 15 ppm, the
only exposure without significant maternal toxicity. In this study the LOAEL for developmental
toxicity is 15 ppm (60 mg/m3), the lowest exposure tested.
24
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Murray et al. (1979) evaluated developmental toxicity of 1,1-DCE administered by
inhalation to pregnant Sprague-Dawley rats (body weight 250 g). Animals were exposed to 0
(20 or 47 animals), 20 ppm (44 animals), 80 ppm (30 animals), or 160 ppm (30 animals) for 7
hours/day on GDs 6-15. At 20 ppm there was no maternal toxicity and no effect on embryonal
or fetal development. At 80 and 160 ppm, there was toxicity to the dams (statistically significant
depression in weight gain at GDs 6-9, more severe at 160 ppm). At 80 and 160 ppm, there was
also a statistically significant increased incidence of wavy ribs and delayed ossification of the
skull, which was regarded as an embryotoxic effect. Both effects were more severe at 160 ppm.
No teratogenic effects were seen at any exposure. The NOAEL for developmental toxicity in
this study is 20 ppm (80 mg/m3); the LOAEL is 80 ppm (320 mg/m3). Under the guidelines for
developmental toxicity (U.S. EPA, 1991), these values are not adjusted to continuous exposure.
Murray et al. (1979) evaluated the developmental toxicity of 1,1-DCE administered by
inhalation to New Zealand white rabbits (body weight 3.4-4.7 kg). Animals were exposed to 0
(16 animals), 80 ppm (22 animals), or 160 ppm (18 animals) for 7 hrs/day on GDs 6-18. No
maternal toxicity or effect on embryonal or fetal development was observed at 80 ppm. Toxicity
to both the dams and their developing embryos was observed at 160 ppm, as indicated by a
marked increase in the incidence of resorptions per litter (0.3 ± 0.6 vs. 2.7 ± 3.9) and a
significant change in the incidence of several minor skeletal variations in their offspring,
including an increase in the occurrence of 13 pairs of ribs and an increased incidence of delayed
ossification of the fifth sternebra (data not reported). No teratogenic effects were seen at any
exposure. The NOAEL for developmental toxicity in this study is 80 ppm (320 mg/m3); the
LOAEL is 160 ppm (640 mg/m3). Under the guidelines for developmental toxicity (U.S. EPA,
1991), these values are not adjusted to continuous exposure.
4.4. OTHER STUDIES
4.4.1. Developmental Neurotoxicity
Short et al. (1977b) evaluated developmental neurotoxicity of 1,1-DCE administered by
inhalation to CD-I rats (Charles River). Pregnant rats were exposed to 0 (24 animals), 56 ppm
(20 animals), or 283 ppm (19 animals) for 22-23 hrs/day on GDs 8-20. Maternal toxicity was
seen at both exposures, as shown by weight loss of 7 g per dam at 56 ppm and 15 g per dam at
283 ppm. There was complete resorption of three litters at 283 ppm. A statistically significant
decrease in average pup weight as compared to controls was noted at both exposures on post-
natal day 1. The difference in pup weight between control and exposed groups decreased with
time and disappeared by postnatal day 21. There was no evidence of developmental
neurotoxicity at either exposure in pups tested at various times from postnatal day 1 to day 21 in
a battery of behavioral tasks, including surface righting, pivoting, auditory startle, bar holding,
righting in air, visual placing, swimming ability, physical maturation, and activity. This study
showed evidence of maternal and fetal toxicity at both exposures, but no evidence of
developmental neurotoxicity at either exposure. Accordingly, the NOAEL for developmental
neurotoxicity in this study is 283 ppm (1124 mg/m3), the highest exposure tested.
25
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4.4.2. Cardiac Sensitization
Siletchnik and Carlson (1974) investigated the effects of epinephrine on cardiac
sensitization by exposure to 1,1-DCE in male albino rats. The test animals (body weight,
250-400 g) were exposed to 1,1-DCE at 0 or 25,600 ± 600 ppm and the dose of epinephrine was
titrated to determine the minimum concentration needed to produce arrhythmias. A dose of
4 |ag/kg of epinephrine failed to induce cardiac arrhythmias in air-exposed animals. However,
the dose necessary to produce life-threatening arrhythmias was 2.0 |ig/kg following 58 to 61
minutes of exposure to 1,1-DCE, 1.0 |ig/kg following 64 minutes of exposure to 1,1-DCE, and
0.5 |ag/kg following 67 to 80 minutes of exposure. The cardiac sensitization was found to be
completely reversible upon discontinuance of exposure.
4.4.3. Species Specificity
Speerschneider and Dekant (1995) investigated the metabolic basis for the species- and
sex-specific nephrotoxicity and tumorigenicity of 1,1-DCE. In kidney microsomes from Swiss-
Webster male mice, the rate of oxidation of 1,1-DCE depended on the hormonal status of the
animals. Oxidation of 1,1-DCE was decreased by castration and restored when the castrate was
supplemented with exogenous testosterone. In kidney microsomes from naive female mice, the
rate of oxidation of 1,1-DCE was significantly lower than in males but could be increased by
administration of exogenous testosterone. Using an antibody to rat liver CYP2E1, the
researchers showed expression of a cross-reacting protein in male mouse kidney microsomes that
was regulated by testosterone and correlated with the ability to oxidize 1,1-DCE and other
substrates for CYP2E1 (e.g., p-nitrophenol and chlorozoxazone). The researchers also showed
that different strains of mice express different levels of CYP2E1. The strains most sensitive to
the effects of 1,1-DCE expressed greater levels of CYP2E1. Nephrotoxicity in Swiss-Webster
mice after inhalation of 1,1-DCE was observed in males and in females treated with exogenous
testosterone, but not in naive females. In kidney microsomes obtained from both sexes of rats
and in six samples of human kidney from male donors, no p-nitrophenol oxidase activity was
detected. Other research groups have also reported the absence of detectable CYP2E1 in human
kidney tissue (Amet et al., 1997; Cummings et al., 2000).
4.4.4. Genetic Toxicity
Reitz et al. (1980) investigated the ability of 1,1-DCE to cause DNA alkylation, DNA
repair, DNA replication, and tissue damage in liver and kidney of rats and mice. Male Sprague-
Dawley rats (body weight, 200-250 g) and male CD-I mice (body weight, 18-20 g) were
exposed by inhalation for 6 hours. Rats were exposed to 0 or 10 ppm; mice were exposed to 0,
10, or 50 ppm. In rats at 10 ppm, there was only a minimal increase in DNA alkylation and a
small increase in DNA replication (twofold increase in 3H-thymidine incorporation) in the
kidney but no increase in liver. In mice at 10 and 50 ppm, there was only a minimal increase in
DNA alkylation. In mice DNA repair was not increased in liver or kidney at 10 ppm or in liver at
26
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50 ppm, but was increased in kidney at 50 ppm. In kidney of mice, there was an eightfold
increase in DNA replication at 10 ppm and a 25-fold increase at 50 ppm, as measured by 3H-
thymidine incorporation. There was a corresponding increase in mitotic figures. No
histopathological damage or increased DNA replication in the liver of mice was observed at 10
or 50 ppm. In mice at 10 ppm, there was slight dilation and swelling and variable amounts of
nephrosis in the kidney, but no effect in the liver. At 50 ppm, mice showed toxic nephrosis in
the kidney and slight centrilobular swelling in the liver.
1,1-DCE induced mutations in Salmonella typhimurium and Escherichia coli in the
presence of an exogenous metabolic system. In Saccharomyces cerevisiae, 1,1-DCE induced
reverse mutation and mitotic gene conversion in vitro and in a host-mediated assay in mice. In a
single study in Saccharomyces cerevisiae, it induced aneuploidy in the presence and absence of
metabolic activation. In vitro, gene mutations were increased in mouse lymphoma cells but not
in Chinese hamster lung cells, with or without an exogenous metabolic system. In a single
study, 1,1-DCE induced sister chromatid exchanges in Chinese hamster lung cells in the
presence of an exogenous metabolic system but not in its absence. In single studies in vivo, 1,1-
DCE did not induce micronuclei or chromosomal aberrations in bone marrow or in fetal
erythrocytes of mice or dominant lethal mutations in mice or rats.
1,1-DCE causes gene mutations in microorganisms in the presence of an exogenous
activation system. Although most tests with mammalian cells show no evidence of genetic
toxicity, the test battery is incomplete, as it lacks an in vivo assessment of chromosomal damage
in the mouse lymphoma assay, a test EPA considers an important component of a genotoxicity
battery. Data on the genetic and related effects of 1,1-DCE are summarized in Table 3.
4.5. SYNTHESIS AND EVALUATION OF MAJOR NONCANCER EFFECTS AND
MODE OF ACTION
There are no useful epidemiological studies or case reports in humans characterizing the
noncancer health effects of 1,1-DCE.
Table 3. Genetic and related effects of 1,1-DCE
Test system
S. typhimurium
BA13/BAL13, forward
mutation
Result3
Without
-
With
+
Dose"
(LED/HID)
500
Reference
Roldan-Arjona et al.,
1991
27
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Test system
S. typhimurium TA 100,
reverse mutation
S. typhimurium TA 104,
reverse mutation
S. typhimurium TA 1535,
reverse mutation
S. typhimurium TA 1537,
reverse mutation
S. typhimurium TA 98,
reverse mutation
S. typhimurium TA 92,
reverse mutation
S. typhimurium TA 97,
reverse mutation
E. coli K12, forward or
reverse mutation
Result3
Without
NT
NT
-
+
NT
-
(+)
-
-
NT
-
-
-
-
-
-
-
With
+
+
+
+
+
+
+
-
+
+
+
(+)
+
(+)
(+)
+
(+)
Dose"
(LED/HID)
2% in air
5% in air
5% in air
5% in air
2% in air
375 ppm in air
125
500
3% in air
5% in air
375 ppm in air
375 ppm in air
375 ppm in air
125
375 ppm in air
5
242
Reference
Malaveille et al.,
1997
Jones and Hathway,
1978c
Simmon and Tardiff,
1978
Waskell, 1978
Bartsch et al., 1979
Oeschetal., 1983
Strobel and Grummt,
1987
Strobel and Grummt,
1987
Baden etal., 1977
Jones and Hathway,
1978c
Oeschetal., 1983
Oeschetal., 1983
Oeschetal., 1983
Strobel and Grummt,
1987
Oeschetal., 1983
Strobel and Grummt,
1987
Oeschetal., 1983
28
-------
Table 3. Genetic and related effects of 1,1-DCE (continued)
Test system
E. coli WP2 uvrA, reverse
mutation
S. cerevisiae D7, gene
conversion
S. cerevisiae D7, mitotic
gene conversion
S. cerevisiae D7, reverse
mutation
S. cerevisiae D61.M,
aneuploidy
Gene mutation, Chinese
hamster lung V79 cells, hprt
locus in vitro
Gene mutation, Chinese
hamster lung V79 cells,
ouabain resistance in vitro
Gene mutation, mouse
lymphoma L5178Y cells, tk
locus in vitro
Sister chromatid exchange,
Chinese hamster lung in
vitro
Chromosomal aberrations,
Chinese hamster DON-6
cells in vitro
Chromosomal aberrations,
Chinese hamster fibroblast
CHL cells in vitro
Result3
Without
-
-
+c
-
+c
+
—
—
?
—
—
—
With
+
+
-
+
+
+
—
—
+
+
NT
NT
Dose"
(LED/HID)
375 ppm in air
2910
7300
2910
4876
2435
10% in air
10% in air
0.1 6% in air
75
2910
2000
Reference
Oeschetal., 1983
Bronzetti etal., 1983
Kochetal., 1988
Bronzetti et al., 1983
Kochetal., 1988
Kochetal., 1988
Drevon and Kuroki,
1979
Drevon and Kuroki,
1979
McGregor et al.,
1991
Sawada et al., 1987
Sasaki etal., 1980
Ishidate (ed.), 1983
29
-------
Table 3. Genetic and related effects of 1,1-DCE (continued)
Test system
Chromosomal aberrations,
Chinese hamster lung cells in
vitro
Host-mediated assay, S.
cerevisiae D7 in CD mouse
hosts
Micronucleus test, mouse
bone marrow in vivo
Micronucleus test, mouse
fetal erythrocytes in vivo
Chromosomal aberrations,
Sprague-Dawley rat bone
marrow in vivo
Dominant lethal test, male
CD-I mice
Dominant lethal test, CD rats
Result3
Without
-
+
+
-
-
—
—
With
+
NT
NT
-
-
—
—
Dose"
(LED/HID)
250
100 po x 23
400 po x 1
200 po x 1
100 po x 1
6 hrs/day,
3 days/wk,
2yrs
50 ppm inh,
6 hrs/day,
5 days
55 ppm inh
6 hrs/day,
5 days/wk,
11 wks
Reference
Sawadaetal., 1987
Bronzetti etal., 1981
Bronzetti et al., 1981
Sawadaetal., 1987
Sawadaetal., 1987
Rampy etal., 1977
Anderson et al.,
1977
Short etal., 1977c
a +, positive; (+), weak positive; -, negative; NT, not tested; ?, inconclusive.
b LED, lowest effective dose; HID, highest ineffective dose. In vitro tests, i-ig/ml; in vivo tests, mg/kg body
weight; po, orally; inh, inhalation
0 Positive in cells grown in logarithmic phase
In laboratory animals 1,1-DCE is rapidly absorbed following oral and inhalation
exposure. Most of the free 1,1-DCE, its metabolites, and covalently bound derivatives are found
in the liver and kidney. 1,1-DCE is rapidly oxidized by CYP2E1 to 1,1-DCE epoxide, which can
be transformed to 2-chloroacetyl-chloride and 2,2-dichloroacetaldehyde (Figure 1). It is not
known whether the metabolism of 1,1-DCE is the same in humans, although in vitro microsomal
preparations from human liver and lung form the same initial products.
30
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Following acute exposure by the oral or the inhalation route, the target organs are the
liver, the kidney, and the Clara cells of the lung.
Following longer term and chronic exposure at less than an acutely toxic exposure, the
liver is the major target in rats following oral or inhalation exposure. The minimal fatty change
observed in the liver of rats following long-term exposure—the critical effect—occurs primarily
in mid-zonal hepatocytes, but the change is not restricted to the centrilobular region. The
minimal fatty change in the liver also occurs in the absence of significant depletion of cellular
GSH. It is not known whether this reversible effect is the consequence of covalent binding of
1,1-DCE derivatives formed in situ by CYP2E1 or of disruption of phospholipid synthesis in the
cells. Although the minimal fatty change might not be considered adverse—as there is no
evidence of a functional change in the liver in rats exposed at this level, and GSH levels are not
reduced—it is defined as the critical effect from both oral and inhalation exposure because
limiting exposure to this level will protect the liver from more serious damage (for example,
fatty liver or necrosis) that could compromise liver function.
The kidney is the major target organ in mice following inhalation exposure. The effects
in the kidney appear to be related to a gender-specific expression of CYP2E1 in male mice, the
presence of higher amount of p-lyase in kidney tissue of mice relative to other species, and the
general pharmacokinetic principle that more 1,1-DCE will be delivered to the kidneys following
inhalation exposure relative to oral exposure.
There is no evidence that toxicity occurs in the respiratory tract following exposure to
1,1-DCE at levels that cause minimal toxicity in the liver of rats and in the kidney of mice.
However, regional responses in olfactory epithelium or bronchiolar changes in Clara cells might
have been missed by the methods used in the toxicological studies to evaluate these regions.
As shown in a three-generation study, there is no evidence that reproductive toxicity is a
critical effect for 1,1-DCE. No reproductive or developmental toxicity was observed at an
exposure that caused minimal toxicity in the liver of the dams. There is also no evidence that
teratogenicity is a critical effect. Some evidence was found of developmental variations in the
heart following direct infusion of 1,1-DCE into the uterus of pregnant rats and fertilized chicken
eggs and ingestion of 1,1-DCE by pregnant rats from drinking water, but it is not clear whether
these effects were directly caused by exposure to 1,1-DCE. The biological significance of these
cardiac structural variations is unclear. There is no indication that the structural variations have
functional consequences in the animals. However, animals known to have the structural
variations have not been tested under conditions of stress.
There are no focused studies on neurotoxicity, but no indication from chronic,
reproductive, and developmental bioassays in rats and mice by oral or inhalation exposure that
neurotoxicity is an important toxic endpoint. No long-term studies have evaluated
immunotoxicity in laboratory animals by any route of exposure; however, the existing bioassays
provide no suggestion that immunotoxicity is a critical effect.
-------
These various observations on toxicity and metabolism of 1,1 -DCE indicate that
cytotoxicity is associated with cytochrome P450-catalyzed metabolic activation of 1,1 -DCE to
reactive intermediates that bind covalently to cellular macromolecules. The extent of binding is
inversely related to loss of GSH, so that severities of tissue damage parallel the decline in GSH
(Forkert and Moussa, 1991; Moussa and Forkert, 1992). Hepatotoxicity is also exacerbated by
treatments that diminish GSH (McKenna et al., 1978b; Andersen et al., 1980; Jaeger et al., 1973,
1974). Thus, the responses to 1,1-DCE at low doses, which cause little depletion of GSH, are
expected to be very different from the responses at high doses, which cause substantial GSH
depletion. The targets of toxicity are centrilobular hepatocytes and bronchiolar Clara cells
(Forkert et al., 1986), cell types that are rich in CYP2E1 (Forkert et al., 1991; Forkert, 1995).
Immunohistochemical studies showed formation of DCE epoxide-cysteine protein adducts
within the centrilobular hepatocytes and Clara cells (Forkert et al., 1999a, b). In combination,
these findings indicate that DCE-induced toxicity is associated with formation and reactivity of
the DCE epoxide within the target centrilobular hepatocytes and Clara cells.
In the absence of specific information on the toxicity of 1,1-DCE in humans, the most
scientifically appropriate way of conducting a risk assessment would be to use a PBPK model to
calculate the concentration of the toxic metabolite in the target tissue. The model would
incorporate the appropriate physiological variables for laboratory animals and humans and what
is known about the mode of action for 1,1-DCE. As discussed above, the toxicity of 1,1-DCE is
attributed to its metabolites, not to the parent compound. Intracellular GSH provides a
mechanism for detoxification of the metabolites. Toxicity is attributed to the amount of
metabolite that escapes conjugation in the liver. The model would thus also incorporate
information on the rate of metabolism of 1,1-DCE in the liver, the initial amount of GSH in the
liver, the rate of conjugation of GSH with the reactive metabolites, and the rate of regeneration
of GSH.
Such a model is available for vinyl chloride (Clewell et al., 1999a, b). In fact, the vinyl
chloride model was developed using a simpler model developed by D'Souza and Andersen
(1988) for 1,1-DCE. The vinyl chloride model has been validated for humans by successfully
predicting the concentration of vinyl chloride in volunteers. Application of the model for oral
exposure to vinyl chloride shows that the human equivalent dose is equal to the dose to rats
divided by approximately 1.4; similarly, for inhalation exposure the human equivalent
concentration is equal to the inhalation exposure to rats divided by approximately 5. These
factors are not significantly different from those determined for 1,1-DCE in the original model
developed by D'Souza and Andersen (1988).
EPA does not believe that it is appropriate to apply the vinyl chloride model to this
assessment for 1,1-DCE at this time. For dose estimates in liver, the original 1,1-DCE model
needs to be updated to include the more current understanding of the metabolism of 1,1-DCE. In
addition, it also appears necessary for the model to estimate dose in other target tissues, namely,
the lung and kidney. For EPA to provide such analysis is beyond the scope of effort for an
assessment for the IRIS program. EPA also does not believe that there is adequate information
to apply the vinyl chloride model to 1,1-DCE using a simpler parallelogram approach (Jarabek et
32
-------
al., 1994; Williams et al., 1996). EPA may, however, modify this assessment when a more
complete PBPK model is available.
In the absence of a suitable PBPK model, EPA used its default procedure to determine
the RfD and the default procedure for a category 3 gas to determine the RfC. EPA recognizes
the scientific limitations of this approach for determining the RfC. The default procedure for a
category 3 gas was developed with the assumption that the parent gas, not a metabolite, is the
toxic substance. As discussed above, that is not the case for 1,1-DCE. At an exposure much less
than the point of saturation of the oxidative pathway for 1,1-DCE (approximately 200 ppm),
essentially all of the absorbed 1,1-DCE is metabolized in the liver. Under these conditions there
will be a constant ratio between the concentration of 1,1-DCE in the ambient air and the
concentration of the toxic metabolite in the liver. Therefore, the category 3 gas default
procedure will provide a reasonable approximation of the exposure-response relationship.
4.6. WEIGHT-OF-EVIDENCE EVALUATION AND CANCER
CHARACTERIZATION
Under the 1986 cancer guidelines (U.S. EPA, 1986a), 1,1-DCE is assigned to Group C,
possible human carcinogen.
Under the draft revised guidelines for carcinogen risk assessment (U.S. EPA, 1999), EPA
concludes 1,1-DCE exhibits suggestive evidence of carcinogenicity but not sufficient evidence to
assess human carcinogenic potential following inhalation exposure in studies in rodents. Male
mice developed kidney tumors at one exposure in a lifetime bioassay, a finding tempered by the
absence of similar results in female mice or male or female rats and by the enzymatic differences
(i.e., CYP2E1) between male mice and female mice, male and female rats, and human kidney
cells. Limited evidence of genotoxicity has been reported in bacterial systems with metabolic
activation. The data for 1,1-DCE are inadequate for an assessment of human carcinogenic
potential by the oral route, based on the absence of statistically or biologically significant tumors
in limited bioassays in rats and mice balanced against the suggestive evidence in male mice in a
single bioassay by inhalation and the limited evidence of genotoxicity. The human
epidemiological results on the carcinogenicity of 1,1-DCE are too limited to draw useful
conclusions. EPA concludes that the results of kidney tumors in one sex and one exposure in a
single species of rodents are too limited to support an exposure-response assessment.
Bioassays for cancer by the oral route of exposure have been conducted in rats (Maltoni
et al., 1985; NTP, 1982; Ponomarkov and Tomatis, 1980; Quast et al., 1983), mice (NTP, 1982),
and trout (Hendricks et al., 1995). Some of these bioassays were conducted at an exposure
below the maximum tolerated dose. The bioassay conducted by Maltoni et al. (1985) exposed
the animals for only 1 year. The bioassay conducted in rats by Quast et al. (1983) and the
bioassay conducted in mice by NTP (1982) were well conducted, and both showed some toxicity
in the liver at the highest exposure. Neither of these bioassays provides any significant evidence
that 1,1-DCE is a carcinogen by the oral route of exposure. The genotoxicity studies are
incomplete, but most studies in mammalian cells indicate a lack of genotoxicity.
33
-------
This assessment of carcinogen!city by the oral route of exposure under the draft revised
guidelines for carcinogen risk assessment (U.S. EPA, 1999) differs from the previous EPA
evaluation (U.S. EPA, 1985a, b, 1987). The previous evaluation derived an oral slope factor
from the highest of four slope factors calculated from two studies (NTP, 1982; Quast et al.,
1983) that did not show statistically significant increases in tumor incidence attributable to oral
exposure. The highest slope factor was based on the adrenal pheochromocytomas in male rats
(NTP, 1982). Following the 1999 draft revised guidelines for carcinogen risk assessment, EPA
emphasizes the importance of using data that show a statistically significant increase in tumor
incidence for calculating a slope factor. As there is no statistically or biologically significant
increase in tumor incidence at any site in the relevant oral bioassays, the present evaluation
characterizes the weight of evidence as inadequate and, accordingly, does not derive an oral
slope factor. This conclusion is consistent with the evaluation by the International Agency for
Research on Cancer (IARC, 1999).
Bioassays for cancer by the inhalation route of exposure have been conducted in rats (Lee
et al., 1977, 1978; Viola and Caputo, 1977; Hong et al., 1981; Maltoni et al., 1985; Quast et al.,
1986; Cotti et al., 1988), mice (Lee et al., 1977, 1978; Hong et al., 1981; Maltoni et al., 1985),
and hamsters (Maltoni et al., 1985). None of these bioassays was conducted by a protocol that
meets current standards. The major defects in most of these bioassays include exposure of the
animals for 1 year and exposure at less than the maximum tolerated dose. The only bioassay
showing some evidence of carcinogenicity was the study in Swiss-Webster mice (Maltoni et al.,
1985). This study was conducted at or near the maximum tolerated dose, as animals exposed at
50 ppm died after a few exposures. Although the animals were exposed for only 1 year and then
observed until natural death, this study showed an increased incidence of kidney
adenocarcinomas in male mice at 25 ppm but not at 10 ppm. The incidence of mammary
carcinomas in female mice and pulmonary adenomas in male and female mice did not increase
with increased exposure. The responses were actually lower at 25 ppm than at 10 ppm, but
survival and other toxicities were comparable.
There is evidence that the induction of kidney adenocarcinomas is a sex- and species-
specific response related to the expression of CYP2E1 in the kidney of male mice
(Speerschneider and Dekant, 1995; Amet et al., 1997; Cummings et al., 2000). However, the
data presented by these researchers are not sufficient to justify a conclusion that the kidney
tumors in male mice have no relevance for a human health risk assessment. This conclusion is
made with the knowledge that compounds similar in structure to 1,1-DCE (e.g.,
tetrachloroethylene, trichloroethylene, and 1,2-dichloroethylene) produce varying degrees of
kidney tumors in animal bioassays. The genotoxicity studies are incomplete, but most studies in
mammalian cells indicate a lack of genotoxicity. Accordingly, EPA concludes that the data on
the increased incidence of kidney adenocarcinomas in male mice (Maltoni et al., 1985) provide
suggestive evidence of carcinogenicity by the inhalation route of exposure. EPA also concludes,
considering the evidence of a potential sex- and species-specific response, that the results of this
bioassay showing an increase in tumors in one sex and one exposure in a single species of
rodents are too limited to support an exposure-response assessment.
34
-------
This assessment of carcinogen!city by the inhalation route of exposure under the draft
revised guidelines for carcinogen risk assessment (U.S. EPA, 1999) differs from the previous
EPA evaluation (U.S. EPA, 1985a, b, 1987). EPA's previous evaluation considered the
incidence of kidney adenocarcinomas (Maltoni et al., 1985) as providing sufficient evidence of
carcinogenicity to justify deriving an inhalation unit risk for quantifying the potential human
cancer risk. As noted in the paragraph above and in Section 4.4.3, the new data suggesting that
the kidney adenocarcinomas could be a sex- and species-specific response reduce the weight of
evidence for carcinogenicity by the inhalation route of exposure. Accordingly, the present
evaluation does not derive an inhalation unit risk. This conclusion is consistent with the
evaluation by I ARC (1999).
1,1-DCE causes gene mutations in microorganisms in the presence of an exogenous
activation system. Although most tests with mammalian cells have shown no evidence of
genetic toxicity, the test battery is incomplete because it lacks an in vivo test for chromosomal
damage in the mouse lymphoma system.
A number of uncertainties exist in the assessment of the carcinogenicity of 1,1-DCE. As
noted above, many of the bioassays by the inhalation route of exposure were not conducted at
the maximum tolerated dose or for the full lifetime of the animals. EPA has acknowledged this
uncertainty in the weight of evidence classification. In addition, our knowledge of the metabolic
pathways for 1,1-DCE in the human is incomplete. Although it is likely that the initial oxidation
of 1,1-DCE in humans occurs via CYP2E1, there could be other CYP isoforms that could
activate 1,1-DCE. Thus, there is some potential for a species-specific carcinogenic response in
humans similar to the apparent sex- and species-specific response observed by Maltoni et al.
(1985) in the kidney of male mice.
4.7. SUSCEPTIBLE POPULATIONS
There are no adequate epidemiological studies or case reports in humans directly
demonstrating a susceptible human population. However, because of the role of CYP2E1 and
GSH in the expression of toxicity of 1,1-DCE, individuals with high levels of CYP2E1 (e.g.,
abusers of ethanol and individuals routinely exposed to ketones and heterocyclic compounds and
other inducers of CYP2E1) could be more sensitive to the adverse effects of 1,1-DCE. There is
some evidence, however, that the rate of hepatic blood flow is an important limiting factor in the
metabolism of 1,1-DCE (Kedderis, 1997). This effect would reduce the importance of the
variability among individuals in concentration of CYP2E1 in the liver as a determinant of
susceptibility to the adverse effects of 1,1-DCE. Individuals at risk from exposure to 1,1-DCE
would also include those who have an extremely low level of GSH, for example, individuals who
are malnourished or fasting or who are poisoned from acetaminophen (Wright and Moore, 1991).
4.7.1. Possible Childhood Susceptibility
Although there are many drugs that exhibit a higher systemic clearance in children than
in adults, no studies in laboratory animals or epidemiological studies or case reports in humans
35
-------
have demonstrated increased susceptibility (i.e., greater response at the same exposure) of
children to 1,1-DCE. The major determinants of the liver toxicity of 1,1-DCE are the CYP2E1
and the GSH content of the liver, cardiac output, and liver volume. CYP2E1 was not detectable
in fetal liver samples from humans, but it increased dramatically within hours after birth and
nearly reached levels found in adults by 1 year of age (Cresteil, 1998). Fetal CYP2E1 may be
inducible by exposure to CYP2E1 substrates such as ethanol (Carpenter et al., 1996). No
significant difference between children and adults was found in the activity of CYP2E1 using a
nonspecific substrate, ethoxycoumarin (Blanco et al., 1999). On the basis of these observations,
it does not seem likely that children will exhibit increased susceptibility to the adverse liver
effects of 1,1-DCE. The variability between children and adults in the GSH content of the liver,
cardiac output, and liver volume are likely to be within the intraspecies uncertainty factor (UF)
of 10 used to derive the RfD and the RfC.
As noted in section 4.3.2, some data in laboratory animals studies suggest an increased
incidence of cardiac changes following exposure to 1,1-DCE (Dawson et al., 1990; Goldberg et
al., 1992; Dawson et al., 1993). It would be helpful if more definitive studies with a greater
range of exposures were conducted to determine the cause and biological significance of the
cardiac changes apparently associated with exposure to 1,1-DCE during the period of cardiac
organogenesis.
4.7.2. Possible Gender Differences
Some data suggest that nephrotoxicity might be a specific response in male mice, and
there is some indication that female rats might be more sensitive than male rats to hepatotoxicity,
as the fatty change—the critical effect—appeared at a slightly lower exposure in female rats as
compared to male rats. There are no epidemiological studies or case reports in humans
suggesting gender specificity for any target tissue.
5. DOSE-RESPONSE ASSESSMENTS
5.1. ORAL REFERENCE DOSE (RfD)
5.1.1. Choice of Principal Study and Critical Effect
The candidate oral chronic or long-term studies for deriving the RfD included the studies
by Maltoni et al. (1985) and Quast et al. (1983) in Sprague-Dawley rats, the three-generation
reproduction study by Nitschke et al. (1983) in Sprague-Dawley rats, the study by NTP (1982) in
F344 rats, and the study by NTP (1982) in C57B16 mice. The Maltoni et al. (1985) study was
rejected because the animals were exposed for only 1 year and there was no evaluation of
endpoints at the termination of exposure. The Nitschke et al. (1983) study provides evidence of
minimal liver toxicity at exposures comparable to those reported in Quast et al. (1983); however,
Nitschke et al. did not provide information on the actual exposure of the animals the number of
animals responding at each exposure or a statistical analysis of the results. The NTP (1982)
36
-------
study in rats did not show any toxicity at the highest exposure tested (5 mg/kg-day) but did show
a nonstatistically significant increase in liver necrosis in male mice at 10 mg/kg-day following
gavage dosing. This study was not used because the gavage route of exposure affects the
pharmacokinetics of 1,1-DCE and the exposure-response relationship.
The Quast et al. (1983) study exposed animals for 2 years to 1,1-DCE in drinking water
and provided exposure-response data for minimal toxicity in the liver (hepatocellular midzonal
fatty change and hepatocellular swelling). These data were used to derive the RfD. The
incidence of hepatocellular swelling was statistically significant at all exposures in female rats.
On the basis of the minimal nature of this effect, as reported by the authors, this response is not
considered to be biologically significant in this study. Nonetheless, BMD modeling was
conducted on the exposure-response data for this effect and revealed a BMD10 of 7.7 mg/kg-day
and a BMDL10 of 4.7 mg/kg-day (analysis not presented). The critical effect is hepatocellular
midzonal fatty change in female rats. The NOAEL for this effect is 9 mg/kg-day, the LOAEL is
14 mg/kg-day, the BMD10 is 6.6 mg/kg-day, and the BMDL10 is 4.6 mg/kg-day (Appendix B).
Although this minimal effect might not be considered adverse—as there is no evidence of a
functional change in the liver in rats exposed at this level, and GSH levels are not reduced—the
BMDL10 is used to derive the RfD, as limiting exposure to the BMDL10 will protect the liver
from more serious damage (for example, fatty liver or necrosis) that could compromise liver
function.
5.1.2. Methods of Analysis
BMD modeling using data from Quast et al. (1983) for the critical effect was used to
determine the BMD10 of 6.6 mg/kg-day and the BMDL10 of 4.6 mg/kg-day.
As discussed in section 4.5, no validated pharmacokinetic model was available for this
assessment. Accordingly, EPA used its default procedure for determining the RfD.
5.1.3. RfD Derivation
The RfD of 0.05 mg/kg-day was calculated from the BMDL10 of 4.6 mg/kg-day and a
total UF of 100 and a modifying factor (MF) of 1 (4.6 mg/kg-day x 1/100 = 0.046, rounded to
0.05 mg/kg-day). Individual UFs of 10 each were used for interspecies extrapolation and
intraspecies variability because there were no applicable data to justify departing from the
default values. Derivation of the RfD from the BMDL10 for the minimal fatty change in the liver
does not require an effect-level extrapolation. This conclusion is based on the minimal nature of
the fatty change and its questionable biological significance because of the absence of any
observable functional deficit in the liver. A subchronic-to-chronic extrapolation factor was not
applied because the Quast et al. (1983) study exposed the animals for 2 years. A database UF
was not applied because the database is considered complete.
A number of long-term bioassays in rodents exposed by the oral or inhalation route show
that liver toxicity is the critical effect. There is no chronic bioassay in a nonrodent mammal;
37
-------
however, 90-day bioassays in several species (rats, mice, dogs, guinea pigs, rabbits, and
monkeys) suggest similar exposure-response relationships across species. Therefore, the lack of
a chronic bioassay in a nonrodent mammal is not considered a data gap. No focused studies of
90 days or longer exist for evaluating neurotoxicity or immunotoxicity. EPA does not consider
these data gaps compelling enough to require application of a database UF.
This RfD differs from the previous EPA value of 0.009 mg/kg-day. The previous EPA
evaluation used the same study but considered the lowest exposure of 9 mg/kg-day in female rats
as a LOAEL for minimal hepatocellular fatty change and minimal hepatocellular swelling and
applied a total UF of 1000 (10 for LOAEL-to-NOAEL extrapolation, 10 for interspecies
extrapolation, and 10 for human variability). As noted above, EPA no longer considers
hepatocellular swelling in the absence of other effects, such as increased liver enzymes in the
serum, as biologically significant in this bioassay. The increased incidence of midzonal fatty
change at 9 mg/kg-day in female rats is not statistically significant. The NOAEL in this bioassay
in 9 mg/kg-day. In addition, the present evaluation uses BMD methodology and calculates
BMDL10 for midzonal fatty change in female rats.
5.2. INHALATION REFERENCE CONCENTRATION (RfC)
5.2.1. Choice of Principal Study and Critical Effect
The candidate studies for deriving the RfC included the studies by Maltoni et al. (1985)
in Sprague-Dawley rats and Swiss-Webster mice and by Quast et al. (1986) in Sprague-Dawley
rats. The Maltoni et al. (1985) study was rejected because the animals were exposed for only 1
year, and there was no evaluation of endpoints at the termination of exposure. Thus the true
incidence of the effect due to exposure to 1,1-DCE cannot be determined. The Quast et al.
(1986) study exposed the animals for 18 months and provided exposure-response information for
minimal toxicity in the liver. The critical effect is minimal hepatocellular midzonal fatty change
in female Sprague-Dawley rats. The NOAEL for this effect in female Sprague-Dawley rats is 25
ppm, the LOAEL is 75 ppm, the BMC10 is 15.1 ppm, and the BMCL10 is 9.8 ppm (Appendix B).
The BMCL10 adjusted to continuous exposure (BMCLADJ) is 1.75 ppm (6.9 mg/m3) (9.8 ppm x
6/24 x 5/7 = 1.75 ppm, 1.75 ppm x 3.97 mg/m3 per ppm = 6.9 mg/m3). Although this minimal
effect might not be considered adverse—as there is no evidence of a functional change in the
liver in rats exposed at this level and GSH levels are not reduced—the BMCL10 is used to derive
the RfC, as limiting exposure to the BMCL10 will protect the liver from more serious damage
(fatty liver or necrosis) that could compromise liver function.
5.2.2. Methods of Analysis
As discussed in section 4.5, no validated PBPK model is available for interspecies
extrapolation. Accordingly, EPA used its default procedure for a category 3 gas (a gas that is
relatively insoluble and unreactive in the extrathoracic and tracheobroncial liquid and tissue
[U.S. EPA, 1994b]) to determine the RfC. BMD analysis was used to determine a BMC10 of
15.1 ppm and aBMCL10 of 9.8 ppm. The BMCLADJ is 6.9 mg/m3. The human equivalent
38
-------
concentration for the BMCL10 (BMCLj^) is calculated using inhalation dosimetry for a category
3 gas:
BMCLmc = BMCLADJ x (Hb/g)A/(Hb/g)H
The blood:air partition coefficient in rats [Hb/g)A] is 5 (D'Souza and Andersen, 1988). No
published data are available to determine the blood:air partition coefficient in humans [(Hb/g)H)].
Unpublished data from a single measurement in one person (verbal statement by M. Andersen,
Colorado State University, to R. Benson, U.S. EPA, Aug. 7, 2001) suggest a value for the
blood:air partition coefficient of 1.75. EPA does not consider this observation sufficiently robust
for deriving the RfC. In addition, EPA has made a policy decision that a ratio for the blood:air
partition coefficient greater than 1 will not be used to derive the RfC (U.S. EPA, 1994b).
Therefore, the default value of 1 is used for (Hb/g)A/(Hb/g)H. The BMCLj^ is 6.9 mg/m3
3
5.2.3. RfC Derivation
The RfC of 0.2 mg/m3 is calculated from the BMCLj^c of 6.9 mg/m3 in a chronic
bioassay using a total UF of 30 and an MF of 1 (6.9 mg/m3 x 1/30 = 0.23, rounded to 0.2 mg/m3).
A UF of 3 is used for interspecies extrapolation because a dosimetric adjustment was used.
There is some suggestion that effects in the kidney of mice may occur at an exposure lower than
the level that causes effects in the liver of rats. Thus, there is some uncertainty as to whether the
most sensitive species has been used to derive the RfC. As noted above, however, the long-term
study in mice (Maltoni et al., 1985) is not suitable for deriving the RfC. A UF of 10 is used for
intraspecies variability because there were no applicable data to justify departure from the
default value. Derivation of the RfC from the BMCL10 for the minimal fatty change in the liver
does not require an effect-level extrapolation. This conclusion is based on the minimal nature of
the fatty change and its questionable biological significance because of the absence of any
observable functional deficit in the liver. Although the rats in Quast et al. (1986) were exposed
for 18 months rather than for their full lifetime, there was no indication that the fatty change was
progressing. In contrast, the evidence indicated that the fatty change was decreasing in
incidence with continued exposure. EPA, therefore, did not apply a subchronic-to-chronic
extrapolation factor. A database UF was not applied because the database is considered
complete.
A number of long-term bioassays in rodents exposed by the oral or inhalation route show
that liver toxicity is the critical effect. There is no chronic bioassay in a nonrodent mammal;
however, 90-day bioassays in several species (rats, mice, dogs, guinea pigs, rabbits, and
monkeys) suggest similar exposure-response relationships across species. Therefore, the lack of
a chronic bioassay in a nonrodent mammal is not considered a data gap. No studies of 90 days
or longer exist for evaluating neurotoxicity or immunotoxicity. EPA does not consider these
data gaps compelling enough to require application of a database UF.
The previous EPA evaluation did not derive an RfC.
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5.3. CANCER ASSESSMENT
None of the bioassays by the oral route of exposure provide sufficient evidence that 1,1-
DCE is a carcinogen. Accordingly, EPA did not derive an oral slope factor. This differs from
EPA's previous evaluation (U.S. EPA, 1987), which relied on studies that did not show a
statistically significant increase in tumor incidence attributable to oral exposure to 1,1-DCE.
One bioassay by the inhalation route of exposure showed suggestive evidence of
carcinogenicity for humans. There is evidence suggesting that the tumor response in male mice
is a sex- and species-specific response. While the previous EPA evaluation relied on these data,
EPA does not currently believe that the suggestive evidence of a tumor response provides
sufficient weight of evidence to justify deriving an inhalation unit risk.
6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION
OF HAZARD AND DOSE RESPONSE
6.1. HUMAN HAZARD POTENTIAL
1,1-DCE does not occur naturally. It is used mainly in the production of PVDC, which is
used principally in food packaging. 1,1-DCE can be found in the environment from release
during its manufacture and use, from the breakdown of products containing its polymers, and
from breakdown of other chlorinated ethanes and ethenes.
1,1-DCE is rapidly absorbed following oral and inhalation exposure. It is rapidly
oxidized by CYP2E1 to reactive intermediates that bind covalently with tissue macromolecules,
or it can be conjugated with tissue GSH. The GSH status of the exposed animal is a major
determinant in the expression of cellular toxicity. In addition, the presence of renal CYP2E1 and
renal p-lyase activity seem to be major determinants in the expression of nephrotoxicity in mice.
As there is evidence that human kidney does not contain CYP2E1, the kidney is unlikely to be a
target tissue in humans.
There are no useful epidemiological studies or case reports in humans showing adverse
health effects. The target organs for noncancer effects in laboratory animals are the liver, the
kidney, and the Clara cells of the lung. A number of bioassays show that 1,1-DCE is a not
carcinogen by the oral or dermal route of exposure. One bioassay in male mice shows
suggestive evidence that 1,1-DCE is a carcinogen by the inhalation route of exposure. However,
the weight of evidence is not sufficient to conclude that carcinogenesis is the critical effect by
the inhalation route of exposure. No useful epidemiological studies or case reports exist that
directly demonstrate a susceptible human population or increased susceptibility of children to the
adverse effects of 1,1-DCE. Some data demonstrate gender specificity in mice to the increased
incidence of renal adenocarcinomas, but no useful epidemiological studies or case reports in
humans suggest gender specificity for any target tissue.
40
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6.2. DOSE RESPONSE
The RfD of 0.05 mg/kg-day was calculated from the BMDL10 for liver toxicity of 4.6
mg/kg-day in a chronic bioassay in rats using a total UF of 100 and an MF of 1 . Individual UFs
of 10 each were used for interspecies extrapolation and intraspecies variability.
The RfC of 0.2 mg/m3 was calculated from the BMCLj^c for liver toxicity of 6.9 mg/m3
in a chronic bioassay in rats using a total UF of 30 and an MF of 1 . A UF of 3 was used for
interspecies extrapolation, as a dosimetric adjustment was used. A UF of 10 was used for
intraspecies variability.
Data showing equivocal carcinogen! city by the oral route of exposure are not sufficient to
justify calculating an oral slope factor under the draft revised cancer guidelines (U.S. EPA,
1999). The suggestive data showing carcinogen! city by the inhalation route of exposure are not
considered of sufficient weight to justify calculating an inhalation unit risk.
41
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Reynolds, ES; Moslen, MT; Boor, JP; et al. (1980) 1,1-dichloroethylene hepatoxicity. Time
course of GSH changes and biochemical aberrations. Am J Pathol 101:331-342.
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Reynolds, ES; Kanz, MF; Chieco, P; et al. (1984) 1,1-dichloroethylene: an apoptic hepatoxin?
Environ Health Perspect 57:313-320.
Roldan-Arjona, T; Garcia-Pedrajas, D; Luque-Romero, L; et al. (1991) An association between
mutagenicity of the Ara test of Salmonella typhimurium and carcinogenicity in rodents for 16
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Sasaki, M; Sugimura, K; Yoshida, MA; et al. (1980) Cytogenic effects of 60 chemicals on
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Sawada, M; Sofuni, T; Ishidate, M, Jr. (1987) Cytogenic studies on 1,1-dichloroethylene and its
two isomers in mammalian cells in vitro and in vivo. Mutat Res 187:157-163.
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APPENDIX A. SUMMARY OF EXTERNAL PEER REVIEW COMMENTS
AND DISPOSITION
At the Peer Review Workshop held on August 7, 2001, in Washington, DC, the Peer
Review Panel addressed each of the General Questions and Chemical-Specific Questions in its
charge. The questions and a summary of the Panel's responses follow. EPA also received
scientific comments from the public. These comments are included in a separate section. EPA
provides a response only if the recommendation differs significantly from what EPA included in
the final assessment or if additional explanation was necessary.
Scientific Comments from the Peer Review Panel
General Question 1: Are you aware of any other data/studies that are relevant (i.e., useful
for the hazard identification or dose-response assessment) for the assessment of the adverse
health effects, both cancer and noncancer, of this chemical?
The Panel was not aware of any other primary toxicity studies with 1,1-DCE that need to
be considered. One panelist provided copies of recent mechanistic studies completed by the
panelist's research group on the relationship between metabolism and toxicity in mice. A
continuing theme throughout the comments from several panelists was the strong
recommendation to emphasize the very-well-developed understanding of the mode of action of
1,1-DCE, including the mechanistic basis of tissue toxicity caused by 1,1-DCE metabolites in
lung, liver, and kidney in rodents.
General Question 2: (a) For the RfD and the RfC, has the most appropriate critical effect
been chosen (i.e., that adverse effect appearing first in a dose-response continuum)? (b) For
the cancer assessment, are the tumors observed biologically significant? Are the tumors
observed relevant to human health? Points relevant to this determination include whether
the choice follows from the dose-response assessment, whether the effect is considered
adverse, and whether the effect (including tumors observed in the cancer assessment) and
the species in which it is observed is a valid model for humans.
The Panel agreed that the liver fatty changes in Quast et al. (1983, 1986) were the
appropriate critical effects, although the Panel was divided on the question of whether these
minimal, reversible fatty changes were adverse. Two panelists called the response adverse.
Because these alterations appeared to have no impact on organ function or the health of the
animals, other panelists believed that they should be regarded as adaptive rather than adverse
changes.
In the 1986 IRIS documentation for 1,1-DCE, the exposure of 9 mg/kg-day from the oral
study (Quast et al., 1983) was considered the LOAEL and was used as the point of departure for
determining the RfD. The 2001 documents regard this same exposure from Quast et al. (1983)
as a NOAEL and calculate a BMD (BMDL10) as the point of departure for subsequent analysis.
EPA should include a justification explaining why 9 mg/kg-day was previously considered a
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LOAEL and is currently considered a NOAEL. At least one of the panelists was opposed to
using BMD methods when good quality data were available to estimate NOAELs.
The Panel regarded the kidney tumors in Maltoni et al. (1985) as biologically significant,
that is, they were directly related to the 1,1-DCE exposures and were increased in incidence
significantly as compared to controls at the 25 ppm exposure concentration. However, this
increased incidence was found only in a single study, it was found only at the highest
concentration, and it was species-, strain- and gender-specific. The tumors might have relevance
as indicators of potential carcinogenic responses for humans at very high exposures. However,
the mode of action, including metabolism by an enzyme (CYP2E1) present in the mouse kidney
at very much higher activity than in the human kidney, and the intrinsic nephrotoxicity of
1,1-DCE are not expected to lead to carcinogenic potential at much lower environmental
exposures in humans. The enzyme required to bioactivate the S-(2,2-dichloro-l-
hydroxy)ethylglutathione directly in kidney, that is, cysteine-beta-lyase, is also present at much
lower activities in human compared to mouse kidneys. Thus, the high-exposure carcinogenic
responses in mice were not considered a relevant model for human cancer risk at environmental
exposures. Again, one panelist indicated that these differences in bioactivating enzyme activities
between mice and humans were so large that 1,1-DCE should be not be regarded as a potential
human carcinogen at all.
EPA has added wording to indicate that the fatty change might not be considered
adverse, but it is being used to derive the RfD and RfC, as limiting exposure to this level will
protect the liver from more serious damage that might compromise liver function.
EPA believes that BMD analysis has several advantages because it uses more of the
experimental data from the study and allows comparison of different studies using the same
incidence for the effect. For example, in Quast et al. (1983) the response at the lowest exposure
is not statistically different from that of controls, but the response is still elevated (i.e., 25% vs.
13%) and is not much different from the response at the mid-exposure level (29%) that was
statistically different from the control. Similarly the response for necrosis in male mice (NTP,
1982) at the highest exposure was 14%, but this change was not statistically significant (p=0.06)
when a two-tailed test was used.
General Question 3: For the RfD and the RfC, have the appropriate studies been chosen as
"principal"? The principal study should present the critical effect in the clearest dose-
response relationship. If not, what other study (or studies) should be chosen and why?
The Panel unanimously agreed that Quast et al. (1983, 1986) were the appropriate studies
for the RfC and RfD evaluations. The Panel also discussed the Dawson et al. (1993)
developmental study, which suggested an increased incidence of cardiac malformations in
neonatal rats after exposure of dams to 1,1-DCE in drinking water before mating and throughout
gestation. This study was discussed both to assert why the Quast et al. (1983, 1986) studies were
used and why the panel did not recommend use of the Dawson et al. (1993) developmental study
as the principal study.
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Although their reasons differed, the panelists unanimously believed that the Dawson et
al. (1993) developmental toxicity study should not be considered as the principal study or
considered to represent a potential developmental hazard from 1,1-DCE exposure. The reasons
included concerns for the high positive responses on a litter basis in the controls, the lack of
increased response between the two exposures that varied by 900-fold, and quality control issues
identified in a 1996 Agency for Toxic Substances and Disease Registry review of other
developmental toxicity studies with trichloroethylene (TCE) conducted by these investigators.
Quality control issues, including lack of analytical confirmation of the concentrations in the
drinking water in the TCE studies, were brought to the attention of the Panel by one panelist on
the basis of his participation in an earlier review of these studies. Finally, other studies by Fisher
et al., 2001 were cited as failing to replicate developmental cardiac changes with TCE.
Before the discussion of the deficiencies in the developmental toxicity drinking water
studies, no panel member felt that Dawson et al. (1993) study should be used as the principal
study. Interestingly, the panelists were against using the Dawson et al. (1993) study because it
does not provide confidence that the effects were exposure-related and associated with DCE
exposures, not because the changes were variations in cardiac morphology.
General Question 4: Studies included in the RfD and RfC under the heading
"Supporting/Additional Studies" are meant to lend scientific justification for the
designation of critical effect by including any relevant pathogenesis in humans, any
applicable mechanistic information, and any evidence corroborative of the critical effect or
to establish the comprehensiveness of the database with respect to various endpoints (such
as reproductive/developmental toxicity studies). Should other studies be included under
the "Supporting/Additional" category? Should some studies be removed?
In terms of supporting/additional studies, the Panel once more stressed the need to
include (1) information on mode of action of 1,1-DCE regarding its metabolism to toxic
intermediates, (2) the role of GSH in limiting the reactivity of the metabolites with tissue targets,
and (3) the nephrotoxicity of 1,1-DCE as a precursor in development of the mouse tumors. The
toxicity of 1,1-DCE increases markedly in fasted rats and in rats treated to deplete GSH stores.
The protective role of GSH also means that responses at high exposures, where GSH is
significantly depleted, cannot be extrapolated directly to lower exposure levels without
considering the role of GSH in detoxifying metabolites before they react to initiate toxicity.
Some early quantitative attempts were made to assess the protection afforded by GSH against
liver toxicity in rats (Andersen et al., 1980). In general, the Panel noted that quantitative
information related to pharmacokinetics and metabolism of 1,1-DCE has not been optimally
utilized for evaluating the RfC and RfD for this compound.
One panelist suggested including in the IRIS Summary the Short et al. (1977c) paper on
dominant lethality in male rats after vinyl chloride or vinylidene chloride exposures. The
Toxicological Review includes this reference. Suggestions were also make to include specific
references on susceptible populations and children's health issues—two concerns that need to be
addressed in evaluating risks posed by 1,1-DCE to diverse human populations.
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General Question 5: (a) Are there other data that should be considered in developing the
UFs or the MF? (b) Do you consider that the data support use of different values than those
proposed?
The Panel agreed that the UFs applied to derive the RfD were acceptable. EPA based the
RfD on conventional risk assessment methods and, as a result, developed a conservative estimate
of toxicity. EPA should include text about mode of action data and the rationale for the use of a
BMD. The Panel agreed that the UFs applied to derive the RfC were acceptable. However, the
panelists suggested that EPA justify the use of the default values in deriving the RfC and expand
the discussions of mode-of-action data. One panelist also suggested that EPA include text
describing how the RfC would ideally be derived if appropriate mode of action data were
available. This panelist also suggested that EPA consider using the pharmacokinetic models
developed for vinyl chloride or chloroform in the 1,1-DCE assessment. Language clarifying the
text in Section 5.1.2 of the Toxicological Review was also necessary.
General Question 6: Do the confidence statements and weight-of-evidence statements
present a clear rationale and accurately reflect the utility of the principal study, the
relevancy of the critical effect to humans, and the comprehensiveness of the database? Do
these statements make sufficiently apparent all the underlying assumptions and limitations
of these assessments? If not, what needs to be added?
The Panel agreed that the document was well organized and clearly stated the reasons for
selecting the two major studies for the exposure-response analysis. The weight of evidence for
carcinogenic potential was also clearly articulated. The rationale for using the factor of 10 for
intra-individual differences, which would include both knowledge of metabolic parameters and
the role of these parameters together with blood flow in controlling amount metabolized, needs
to be strengthened. In addition, the decision for not using Dawson et al. (1993) as a critical
study needs to be strengthened in relation to the quality and reproducibility of the study rather
than questioning the nature of the changes as normal variation or potentially adverse alterations
in structure. In regard to the carcinogenic response in the high-concentration male mice in the
single positive study, both the mode-of-action discussion and the nephrotoxicity of 1,1-DCE
deserve to be more heavily emphasized.
Chemical-Specific Question 1: Do you agree that the minimal hepatocellular swelling
(Quast et al., 1983) is not an adverse response but that the minimal hepatocellular fatty
change in the midzonal region (Quast et al., 1983,1986) is adverse response?
The Panel unanimously agreed that the hepatocellular swelling is not an adverse
response. The Panel also unanimously agreed that the fatty changes in the oral and inhalation
studies should be used for the exposure-response assessment, although the Panel was not
unanimous in calling these changes adverse. Several panelists believed that these changes are
transient adaptive responses that clarify upon cessation of exposure. However, the Panel did not
believe that calling the responses adverse was imperative for using them for the exposure-
response analysis.
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EPA has added wording to indicate that the fatty change might not be considered
adverse, but it is being used to derive the RfD and RfC, as limiting exposure to this level will
protect the liver from more serious damage that might compromise liver function.
Chemical-Specific Question 2: Do you agree that the cardiac changes (Dawson et al., 1993)
are properly characterized as variations in cardiac morphology and should not be
considered adverse effects?
The Panel's determination that the Dawson et al. (1993) study was unusable was not
because the changes were acceptable variations in cardiac morphology. Instead, the unanimous
opinion of the Panel was that the study does not provide confidence that the effects were
exposure-related and causally associated with DCE exposures. The Panel did not formally
address the question of whether these valvular and septal changes should be regarded as simply
variations in cardiac morphology. However, several panelists stated that these changes would be
considered adverse and suitable for a exposure-response assessment if they were actually related
to 1,1-DCE exposures. Due to the concerns with the study noted in General Question 3, the
Panel unanimously believed that this study should not be used for exposure-response assessment
and believed that there was no convincing evidence that these changes were actually related to
1,1-DCE in the drinking water.
Chemical-Specific Question 3: Is the weight of evidence for cancer from both oral and
inhalation exposure assigned at the appropriate level?
The Panel agreed that the weight of evidence assigned for oral exposures is appropriate,
that is, the available data do not indicate any cancer risks via this route of administration. For
inhalation exposures, the weight of evidence under the new cancer guidelines (U.S. EPA, 1999)
should be used. The Panel felt that the renal tumors observed in the Maltoni et al. (1985) study
are biologically significant, that is, tumor incidence is causally related to exposure to 1,1-DCE.
However, the renal tumors were observed only at the highest exposure level of 25 ppm in male
mice. Such tumors were not observed in any other study regardless of the species, strain,
exposure, or exposure route. On the basis of published studies on the lack of activity of CYP2E1
in human kidneys and the much lower activity of beta-lyase in human kidney, the Panel felt that
the renal tumors observed in the mouse study are of questionable relevance to humans exposed at
environmental levels. Further, the Panel agreed that 1,1-DCE is likely to be carcinogenic only at
exposures at which GSH is depleted and cytotoxicity is expressed. One panelist believed that
1,1-DCE should be regarded as not likely to be carcinogenic to humans. Without dissent, the
Panel unanimously agreed that derivation of an inhalation unit risk (IUR) from the renal tumor
incidence in mice was inappropriate.
The Panel recommended that EPA include the following narrative (from the 1999 draft
EPA guidance document for cancer risk assessment) for 1,1-DCE cancer risk assessment:
Suggestive evidence of carcinogenicity but not sufficient to assess human
carcinogenic potential: This descriptor is applied when carcinogenicity data are
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suggestive but inconclusive. For example, this descriptor would be applicable in
situations where increased tumor incidence is marginal or is observed only in a
single study. According to EPA guidelines, a cancer dose-response assessment is
not indicated for chemicals with this descriptor.
On the basis of the above narrative, neither a quantitative cancer assessment nor an IUR
derivation is warranted for 1,1-DCE.
Chemical-Specific Question 4: Do you agree that an inhalation unit risk should not be
derived from the data on kidney adenocarcinomas in Swiss mice (Maltoni et al., 1985)?
The Panel agreed unanimously that it was inappropriate to use the Maltoni et al. (1985)
study with the high-exposure-level increased incidence of adenocarcinoma in male mice to
derive an IUR. However, differences of opinion existed about whether the cancer endpoint
should be evaluated using a margin-of-exposure approach from the proposed revisions to the
EPA carcinogen risk assessment guidelines.
The Panel noted in Chemical-Specific Question 3 that the 1999 EPA guidelines would
not pursue a cancer risk assessment for the descriptor: Suggestive evidence of carcinogenicity
but not sufficient to assess human carcinogenic potential.
However, additional discussions among the panelists led to the question of the benefit of
conducting an nonlinear cancer risk assessment for 1,1-DCE using a margin-of-exposure
approach with appropriate UFs. Three panelists felt that, based on the in-depth knowledge of the
mode of action, the weight of evidence for 1,1-DCE carcinogenicity, and the toxicity of 1,1-DCE
to mouse kidneys, the use of a margin-of-exposure approach for 1,1-DCE is not warranted at this
time. However, two panelists felt that the Swiss mice data on renal tumor incidence should be
analyzed according to the margin-of-exposure approach, as available quantitative data on 1,1-
DCE metabolism (the relative enzyme levels and GSH levels) in Swiss mice and human kidneys
are not sufficient to ignore the concern of possible renal tumor incidence in humans at high
exposures. One panelist felt that a margin-of-exposure cancer risk approach could be used,
although the UFs applied should be no larger than those recommended for the RfC derivation
with the critical effects in the liver.
EPA does not believe that a margin-of-exposure approach for a cancer risk assessment
for 1,1-DCE is warranted.
Additional Comments
The Panel also provided other comments to improve the scientific quality of the
document. The Panel emphasized adding materials on susceptibility, interactions with other
compounds, specific risks to children and neonates, and influences of lifestyle such as smoking
in altering susceptibility or risk. The Panel strongly suggested a revision of Figure 1 and careful
development of mode-of-action arguments in the text. The Panel provided a suggested revision
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to Figure 1 and a rewrite of the mode-of-action section based on current knowledge of the
toxicology of 1,1 -DCE.
Scientific Comments from the Public
One public commentor wanted EPA to include additional references to LC50 studies.
EPA has incorporated several references to LC50 studies. Several references to LC50
studies listed by the commentor were not included, as they are unpublished reports from BASF
and are not available to EPA.
One public commentor thought that the RfC should be lower than the value selected by
EPA because 1,1-DCE shows acute and developmental toxicity in the 10-25 ppm range, only
slightly higher than the BMCL10 calculated from the results of Quast et al. (1986). The
commentor was concerned that EPA had not used the most sensitive studies and endpoints.
EPA has reevaluated the results of the studies cited. The acute study by Reitz et al.
(1980) shows evidence of slight damage to the kidney of mice at 10 ppm following a single
6-hour exposure. The exposure is equivalent to a continuous exposure of 2.5 ppm. The
developmental study of Short et al. (1977b) in mice shows effects at 15 ppm, the only exposure
not showing severe maternal toxicity. Exposure in this study was for 22-23 hours/day on GDs 6
to 16. The effects in these studies occurred at an exposure higher than the 1.8 ppm calculated
continuous exposure in Quast et al. (1986) and are not used to derive the RfC. The long-term
study by Maltoni et al. (1985) in male mice showed effects in the kidney at 10 ppm. Exposure in
this study was for 4 hrs/day, 4.5 days/wk, for 52 weeks, which is equivalent to a continuous
exposure of 1.1 ppm. Animals were then held without exposure until spontaneous death (total
duration 126 weeks). Because there was no investigation of the effects at the termination of the
1-year exposure, this study cannot be used to derive the RfC. EPA has used the default
interspecies UF of 3 because there is some concern that the most sensitive species was not used
to derive the RfC.
One public commentor, although concurring with EPA's decision not to use the cardiac
changes observed by Dawson et al., (1993) to derive the RfD, wanted a more extensive
discussion of the uncertainty raised by this observation. The commentor suggested that this
uncertainty does not support EPA's decision to assign the database to the "medium" confidence
category in Section I.A.5 of the IRIS summary. The commentor was concerned about the reports
of cardiac abnormalities associated with exposure to chlorinated solvents from drinking water in
human epidemiological studies. The commentor was also concerned that, because there is no
test of cardiac function in stressed animals, the functional consequences of the morphological
changes might not be observable in the developmental studies that have been conducted. The
commentor cited the results of Siletchnik and Carlson (1974) on cardiac sensitization in support
of that latter concern.
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Public commenters were concerned about EPA's withdrawal of the IUR for 1,1 -DCE.
These commentors wanted a stronger justification for EPA's position that the IUR should not be
derived from Maltoni et al. (1985) and a more extensive discussion of the uncertainties and data
gaps. One commentor expressed the view that a full discussion of the uncertainties and data
gaps should lead EPA to a different conclusion on the need for providing an IUR in this
assessment. The commentors advanced a number of scientific reasons for their concern about
the lack of an IUR in this assessment. The most important reasons included (1) the poor quality
of the bioassays on 1,1 -DCE and the suspicious results in some of these bioassays; (2) the
structural correlation between 1,1-DCE and vinyl chloride, the fact that vinyl chloride seems to
express its full carcinogenic potential from short-term exposure to young animals, and the lack
of comparable testing on 1,1-DCE; (3) limitations in the range of metabolites investigated by
Speerschneider and Dekant (1995) that argue against using this study as the primary rationale for
not quantifying the cancer risk; (4) the lack of detailed knowledge about the metabolism of 1,1-
DCE by humans; (5) the possibility that CYP isoforms other than CYP2E1 could activate 1,1-
DCE in humans; and (6) EPA's unstated assumption of concordance in tumor site between mice
and humans.
EPA has added additional discussion of the uncertainties in the cancer assessment to
Section 4.6 of the Toxicological Review. However, EPA does not believe it is appropriate to
increase the weight of evidence in the cancer assessment based on uncertainty. With regard to
the testing of 1,1-DCE for carcinogenicity in immature animals, the only relevant bioassay was
reported by Cotti et al. (1988). This bioassay exposed pregnant Sprague-Dawley rats from GD
12 to parturition and the subsequent offspring for 13 or 104 weeks. The results of this bioassay
provide no convincing evidence that 1,1-DCE is carcinogenic.
Additional References
EPA added the following references to the toxicological review: Blanco et al. (1999);
Clewell et al. (1995a); Costa and Ivanetich (1984); Cresteil (1998); El-Masri et al. (1996a); El-
Masri et al. (1996b); Fisher et al. (2001); Forkert (1995); Forkert (1999a); Forkert et al. (1986);
Forkert et al. (1991); Jaeger et al. (1973); Jarabek et al. (1994); Liebler et al. (1985); Liebler et
al. (1988); Short et al. (1977a); Short et al. (1977c); Williams et al. (1996); and Wright and
Moore (1991).
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APPENDIX B. BENCHMARK DOSE (BMD) ANALYSIS
B.I. ORAL
Data on fatty change in the liver from Quast et al. (1983) were analyzed using EPA's
BMD software. Each of the seven models gave an adequate fit (p>0.2). The gamma, logistic,
multistage, quanta-linear, and Waybill models showed the best visual fit to the data points. The
gamma, multistage, quanta-linear, and Waybill models showed identical Acacia's Information
Criterion (AID) values and identical BMD and BMDLs. The results from the gamma model are
presented.
The form of the probability function is:
P[response] =background+(l-background)*CumGamma[slope*dose,power]
where CumGammaQ is the cumulative Gamma distribution function
Default initial (and specified) parameter values:
Background = 0.12963
Slope = 0.0192007
Power = 1.12817
Asymptotic correlation matrix of parameter estimates:
(*** The model parameter(s) -Power have been estimated at a boundary point or have
been specified by the user and do not appear in the correlation matrix)
Background
Background
Slope
1
-0.54
Slope
-0.54
1
Parameter estimates:
Variable Estimate SE
Background 0.125627 0.0350171
Slope 0.0158781 0.00405428
Power 1 NA
NA indicates that this parameter has hit a bound implied by some inequality constraint
and thus has no standard error.
62
-------
Analysis of deviance table:
Model
Full model
Fitted model
Reduced model
AIC = 242.458
Goodness of Fit:
Logflikelihood) Deviance Test
-119.212
-119.229 0.0326243
-128.113 17.8011
DF p-value
2
3
0.9838
0.0004834
Dose
0
9
14
30
Est. Prob.
0.1256
0.2421
0.2999
0.4570
Expected
10.050
11.619
14.396
21.935
Observed
10
12
14
22
Size
80
48
48
48
Scaled Residual
-0.01693
0.1284
-0.1246
0.01895
Chi-square = 0.03 DF = 2
BMD computation:
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 6.63557
BMDL = 4.61215
^-value = 0.9838
0.6
0.5
0.4
c
o
'•5 0.3
0.2
0.1
Gamma Multi-Hit Model with 0.95 Confidence Level
Gamma Multi-Hit
BMDL BMP
0 5
09:28 04/25 2001
10
15
dose
63
20
25
30
-------
B.2 INHALATION
Data on fatty change in the liver from Quast et al. (1986) were analyzed using EPA's
BMD software. The gamma, multistage, and quantal-linear models gave an adequate fit (p>0.2).
These models also gave an adequate visual fit to the data points. The quantal-linear model gave
the lowest AIC value. The results from this model are presented.
The form of the probability function is:
P[response] = background + (l-background)*[l-EXP(-slope*dose)]
Default initial (and specified) parameter values:
Background = 0.0294118
Slope = 0.00549306
Power = 1 Specified
Asymptotic correlation matrix of parameter estimates:
(*** The model parameter(s) -Background -Power have been estimated at a boundary
point or have been specified by the user and do not appear in the correlation matrix)
Slope = 1
Parameter estimates:
Variable Estimate SE
Background 0 NA
Slope 0.00697979 0.00194885
NA indicates that this parameter has hit a bound implied by some inequality constraint
and thus has no standard error.
Analysis of deviance table:
Model Log(likelihood) Deviance Test DF p-value
Full model -27.7336
Fitted model -28.0929 0.718624 2 0.6982
Reduced model -32.5262 9.58514 2 0.008291
AIC = 58.1858
64
-------
Goodness of Fit:
Dose
0
25
75
Est. Prob. Expected Observed Size Scaled Residual
0.0000 0.000 0 16 0
0.1601 4.643 6 29 0.6869
0.4075 8.151 7 20 -0.5237
Chi-square = 0.75 DF = 2
BMD computation:
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMC = 15.0951
BMCL = 9.84365
/7-value = 0.6886
Quantal Linear Model with 0.95 Confidence Level
0.6
0.5
c 0.3
g
'-4->
% 0.2
0.1
0
Quantal Linear
BMDLBMP
0
10
20
09:42 04/25 2001
30 40
dose
50
60
70
80
65
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