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EPA/63 5/R-09/006D
www.epa.gov/iris
oEPA
TOXICOLOGICAL REVIEW
OF
cis-l,2-DICHLOROETHYLENE
and
trans-1,2-DICHLOROETHYLENE
(CAS Nos. cis: 156-59-2; trans: 156-60-5; mixture: 540-59-0)
In Support of Summary Information on the
Integrated Risk Information System (IRIS)
August 2010
This document is an Interagency Science Discussion and Final Agency Review draft. This
information is distributed solely for the purpose of pre-dissemination peer review under
applicable information quality guidelines. It has not been formally disseminated by EPA. It
does not represent and should not be construed to represent any Agency determination or policy.
It is being circulated for review of its technical accuracy and science policy implications.
U.S. Environmental Protection Agency
Washington, DC
-------
DISCLAIMER
This document is a preliminary review draft for review purposes only. This information is
distributed solely for the purpose of pre-dissemination peer review under applicable information
quality guidelines. It has not been formally disseminated by EPA. It does not represent and
should not be construed to represent any Agency determination or policy. Mention of trade
names or commercial products does not constitute endorsement or recommendation for use.
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CONTENTS—TOXICOLOGICAL REVIEW OF cis-/trans-l,2-DICHLOROETHYLENE
(CAS Nos. cis: 156-59-2; trans: 156-60-5; mixture: 540-59-0)
LIST OF TABLES v
LIST OF FIGURES vii
LIST OF ACRONYMS viii
FOREWORD x
AUTHORS, CONTRIBUTORS, AND REVIEWERS xi
1. INTRODUCTION 1
2. CHEMICAL AND PHYSICAL INFORMATION 3
3. TOXICOKINETICS 5
3.1. ABSORPTION 5
3.1.1. Oral 5
3.1.2. Inhalation 5
3.1.3. Dermal 6
3.2. DISTRIBUTION 7
3.3. METABOLISM 8
3.3.1. Metabolism in Animals 9
3.3.2. Metabolism in Human Preparations In Vitro 13
3.3.3. CYP2E1 Inactivation by 1,2-DCE 13
3.4. ELIMINATION 14
3.5. PHYSIOLOGICALLY BASED PHARMACOKINETIC MODELS 14
4. HAZARD IDENTIFICATION 18
4.1. STUDIES IN HUMANS—EPIDEMIOLOGY, CASE REPORTS, CLINICAL
TRIALS 18
4.2. SHORT-TERM, SUBCHRONIC, AND CHRONIC STUDIES AND CANCER
BIO AS SAYS IN ANIMALS—ORAL AND INHALATION 18
4.2.1. Oral Exposure 18
4.2.1.1. Short-term Studies 18
4.2.1.2. Subchronic Studies 20
4.2.1.3. Chronic Studies 31
4.2.2. Inhalation Exposure 31
4.2.2.1. Short-term Studies 31
4.2.2.2. Subchronic Studies 31
4.2.2.3. Chronic Studies 36
4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND
INHALATION 36
4.3.1. Oral Exposure 36
4.3.1.1. cis-1,2-DCE 36
4.3.1.2. trans-1,2-DCE 36
4.3.1.3. Mixtures of cis- and trans-1,2-DCE 37
4.3.2. Inhalation Exposure 38
4.3.2.1. cis-1,2-DCE 38
4.3.2.2. trans-1,2-DCE 38
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4.3.2.3. Mixtures of cis- and trans-1,2-DCE 39
4.4. OTHER DURATION- OR ENDPOINT-SPECIFIC STUDIES 39
4.4.1. Acute Studies 39
4.4.1.1. Oral Exposure 39
4.4.1.2. Inhalation Exposure 41
4.4.2. In Vivo Neurological Behavioral Studies 43
4.4.3. Immunological Studies 44
4.4.3.1. cis-1,2-DCE 44
4.4.3.2. trans-1,2-DCE 45
4.4.3.3. Mixtures of cis- and trans-1,2-DCE 47
4.4.4. Toxicity Studies by Other Routes 47
4.4.4.1. Intraperitoneal Injection 47
4.4.4.2. Dermal Application 49
4.4.4.3. Eye Irritation 49
4.4.4.4. Skin Irritation 50
4.5. MECHANISTIC DATA AND OTHER STUDIES IN SUPPORT OF THE MODE
OF ACTION 50
4.5.1. Hepatotoxicity Studies 50
4.5.2. Nephrotoxicity Studies 52
4.5.3. Studies with Cell Cultures 53
4.5.4. Genotoxicity 54
4.5.4.1. In Vitro Studies 54
4.5.4.2. In Vivo Studies 58
4.5.5. Quantitative Structure-Activity Relationship Studies 59
4.6. SYNTHESIS OF MAJOR NONCANCER EFFECTS 62
4.6.1. Oral 65
4.6.1.1. cis-1,2-DCE 65
4.6.1.2. trans-1,2-DCE 67
4.6.1.3. Mixtures of cis- and trans-1,2-DCE 73
4.6.2. Inhalation 73
4.6.2.1. cis-1,2-DCE 73
4.6.2.2. trans-1,2-DCE 74
4.6.2.3. Mixtures of cis- and trans-1,2-DCE 76
4.7. EVALUATION OF CARCINOGENICITY 78
4.7.1. Summary of Overall Weight of Evidence 78
4.7.2. Synthesis of Human, Animal, and Other Supporting Evidence 78
4.8. SUSCEPTIBLE POPULATIONS AND LIFE STAGES 79
4.8.1. Possible Childhood Susceptibility 79
4.8.2. Possible Gender Differences 80
4.8.3. Other—Genetic Polymorphisms 80
4.8.3.1. Cytochrome P450 2E1 80
4.8.3.2. Glutathione S-Transferase 81
5. DOSE-RESPONSE ASSESSMENT 82
5.1. ORAL REFERENCE DOSE 82
5.1.1. cis-1,2-DCE 82
5.1.1.1. Choice of Principal Study and Critical Effect—with Rationale and
Justification 82
5.1.1.2. Methods of Analysis, Including Models 83
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5.1.1.3. RfD Derivation—Including Application of Uncertainty Factors 85
5.1.1.4. Previous Oral Assessment 86
5.1.2. trans-1,2-DCE 86
5.1.2.1. Choice of Principal Studies and Critical Effects—with Rationale and
Justification 86
5.1.2.2. Methods of Analysis, Including Models 88
5.1.2.3. RfD Derivation—Including Application of Uncertainty Factors 92
5.1.2.4. Previous Oral Assessment 93
5.2. INHALATION REFERENCE CONCENTRATION 94
5.2.1. cis-1,2-DCE 94
5.2.2. trans-1,2-DCE 94
5.3. UNCERTAINTIES IN THE ORAL REFERENCE DOSE (RfD) 95
5.4. CANCER ASSESSMENT 98
6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HAZARD AND
DOSE RESPONSE 99
6.1. HUMAN HAZARD POTENTIAL 99
6.2. DOSE RESPONSE 101
6.2.1. Noncancer - Oral Exposure 101
6.2.1.1. cis-1,2- DCE 101
6.2.1.2. trans-1,2- DCE 102
6.2.2. Noncancer - Inhalation Exposure 103
6.2.2.1. cis-1,2-DCE 103
6.2.2.2. trans-1,2-DCE 103
6.2.3. Cancer 103
7. REFERENCES 105
APPENDIX A: SUMMARY OF EXTERNAL PEER REVIEW, PUBLIC COMMENTS,
AND DISPOSITION A-l
APPENDIX B: BENCHMARK DOSE MODELING RESULTS AND OUTPUTS B-l
B. 1. RfD for cis-1,2-DCE B-l
B. 1.1. Relative Liver Weight B-14
B.1.2. Relative Kidney Weight B-8
B.2. RfD for trans-1,2-DCE B-ll
B.2.1. Decreased Antibody Directed Against sRBC (Shopp et al., 1985) B-14
B.2.2. Absolute Thymus Weight (Barnes et al., 1985) B-147
B.2.3. Relative Liver Weight (NTP, 2002) B-20
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LIST OF TABLES
2-1. Properties of the 1,2-dichloroethylene isomers and their mixture 4
3-1. Tissue: air partition coefficients of the 1,2-dichloroethylene isomers in the rat 8
4-1. Body weight and relative organ weights of rats exposed to cis-1,2-DCE by gavage for 90
days 21
4-2. Absolute kidney weights in rats treated with trans- 1,2-DCE via drinking water
for 90 days 23
4-3. Final body weights in rats exposed to trans-1,2-DCE in the feed for 14 weeks 25
4-4. Red blood cell counts in rats exposed to trans-1,2-DCE in the feed
for 14 weeks 26
4-5. Relative liver weights in mice and rats exposed to trans-1,2-DCE in
the feed for 14 weeks 27
4-6. Results of 90-day study in CD-I mice exposed to trans-1,2-DCE in
the drinking water 29
4-7. Histopathological changes in subchronic inhalation study of trans-1,2-DCE 32
4-8. Selected hematology findings in rats exposed to trans-1,2-DCE by inhalation for 90
days 34
4-9. Humoral immune response to sRBCs in CD-I mice exposed to trans-1,2-DCE in
drinking water for 90 days (day 4) 46
4-10. Effect of 1,2-DCE isomers on urinary protein and glucose 24 hours after intraperitoneal
treatment of male Swiss mice 53
4-11. In vitro genotoxicity studies using cis- and trans-1,2-dichloroethylene 56
4-12. In vivo genotoxicity studies using cis- and trans-1,2-dichloroethylene 59
4-13. Summary of major noncancer subchronic studies for oral exposure to 1,2-DCE 62
4-14. Summary of major noncancer subchronic studies for inhalation exposure to 1,2-DCE .... 62
5-1. Relative liver and kidney weights of rats exposed to cis-1,2-DCE by gavage for 90 days.. 83
5-2. Humoral immune response to sRBCs in CD-I mice exposed to trans-1,2-DCE in
drinking water for 90 days (Day 4) 89
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5-3. Relative and absolute thymus weights in female mice exposed to trans-1,2-DCE in the
drinking water for 90 days 90
5-4. Relative liver weights in male and female mice and rats exposed to trans-1,2-DCE in
the feed for 14 weeks 91
B-1. BMDS modeling summary of relative liver weights in female rats exposed to
cis-1,2-DCE by gavage for 90 days B-l
B-2. BMDS modeling summary of relative liver weights in male rats exposed to
cis-1,2-DCE by gavage for 90 days B-5
B-3. BMDS modeling summary of relative kidney weight in male rats exposed to
cis-1,2-DCE by gavage for 90 days B-8
B-4. BMDS modeling summary of relative kidney weight in female rats exposed to
cis-1,2-DCE by gavage for 90 days B-l 1
B-5. BMDS modeling summary of decreased antibody directed against sheep RBC in
male mice exposed to trans-1,2-DCE in drinking water for 90 days B-l4
B-6. BMDS modeling summary of decreased absolute thymus weight in female mice
exposed to trans-1,2-DCE for 90 days B-17
B-7. BMDS modeling summary of relative liver weight in male mice exposed to
trans-1,2-DCE in the feed for 14 weeks B-20
B-8 BMDS modeling summary of relative liver weight in female mice exposed to
trans-1,2-DCE in the feed for 14 weeks B-24
B-9. BMDS modeling summary of relative liver weight in female rats exposed to
trans-1,2-DCE in the feed for 14 weeks B-25
LIST OF FIGURES
2-1. Chemical structures of cis- and trans-1,2-dichloroethylene 3
3-1. Proposed metabolic scheme for cis- and trans-1,2-dichloroethylene 9
3-2. PBPK model for cis- and trans-1,2-dichloroethylene in rats 16
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LIST OF ACRONYMS
AAP 4-aminoantipyrine
ACGIH American Conference of Governmental Industrial Hygienists
ADH alcohol dehydrogenase
AFC antibody-forming cell
AH aniline hydroxylation
AIC Akaike Information Criteria
ALP alkaline phosphatase
ALT alanine aminotransferase
AST aspartate aminotransferase
ATSDR Agency for Toxic Substance and Disease Registry
BMD benchmark dose
BMR benchmark response
BUN blood urea nitrogen
CAS Chemical Abstracts Service
CASRN Chemical Abstracts Service Registry Number
CI confidence interval
CNS central nervous system
Con A concanavalin A
CYP450 cytochrome P450
DAF dosimetric adjustment factor
DCA dichloroacetic acid
DCE dichloroethylene
DTH delayed-type hypersensitivity
ECio concentration causing 10% change in effect
ED50 median effective dose
EN-D ethylmorphine N-demethylation
EPA Environmental Protection Agency
G-6-Pase glucose-6-phosphatase
GC gas chromatography
GC/MS gas chromatography/mass spectrometry
GD gestation day
GSH reduced glutathione
GST glutathione S-transferase
GSTZ glutathione S-transferase zeta
HEC human equivalent concentration
ID50 concentration to achieve 50% decrease in immobility or 50% inhibitory dose to
growth of cells
i.p. intraperitoneal or intraperitoneal^
IRIS Integrated Risk Information System
Km Michaelis constant
LC50 median lethal concentration
LD50 median lethal dose
LDH lactate dehydrogenase
LOAEL lowest-observed-adverse-effect level
LPS lipopolysaccharide
MTD maximum tolerated dose
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NOAEL
no-observed-adverse-effect level
NLM
National Library of Medicine
POD
point of departure
PBPK
physiologically based pharmacokinetic
PSP
phenolsulfonephthalein
QSAR
quantitative structure-activity relationship
RAM
rate of metabolism
RBC
red blood cell
RtC
reference concentration
Rfl)
reference dose
RVMT
rate of change of inhibitable metabolism
S9
supernatant fraction
SDH
sorbitol dehydrogenase
SGOT
glutamate oxaloacetate transaminase (now called AST)
SGPT
glutamate pyruvate transaminase (now called ALT)
sRBC
sheep red blood cell
TEARS
thiobarbituric acid-reactive substances
TLV
threshold limit value
UF
uncertainty factor
U.S. EPA
U.S. Environmental Protection Agency
Vmax
maximum substrate turnover velocity
voc
volatile organic compound
WBC
white blood cell
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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 cis- and
trans-1,2-dichloroethylene. It is not intended to be a comprehensive treatise on the chemical or
toxicological nature of cis- and trans-1,2-dichloroethylene.
The intent of Section 6, Major Conclusions in the Characterization of Hazard and Dose
Response, is to present the major conclusions reached in the derivation of the reference dose,
reference concentration and cancer assessment, where applicable, and to characterize the overall
confidence in the quantitative and qualitative aspects of hazard and dose response by addressing,
the quality of data and related uncertainties. The discussion is intended to convey 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 (202) 566-1676 (phone), (202) 566-1749 (fax), or
hotline.iris@epa.gov (email address).
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AUTHORS, CONTRIBUTORS, AND REVIEWERS
CHEMICAL MANAGER
Audrey Galizia, Dr. PH
Office of Research and Development
U.S. Environmental Protection Agency
Edison, NJ
AUTHORS
Audrey Galizia, Dr. PH
Office of Research and Development
U.S. Environmental Protection Agency
Edison, NJ
D. Charles Thompson, R.Ph., Ph.D., DABT
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
CONTRIBUTORS
Ted Berner, M.S.
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
Christine Cai, M.S.
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
CONTRACTOR SUPPORT
C. Clifford Conaway, Ph.D., DABT
Consulting Toxicologist
Mahopac, NY
Janusz Z. Byczkowski, Ph.D., DABT
Toxicology Consultant
Fairborn, OH
Susan Goldhaber, M.S.
Toxicology Consultant
Raleigh, NC
George Holdsworth, Ph.D.
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Oak Ridge Institute for Science and Education
Oak Ridge, TN
REVIEWERS
This document has been provided for review to EPA scientists, interagency reviewers
from other federal agencies and White House offices, and the public, and peer reviewed by
independent scientists external to EPA. A summary and EPA's disposition of the comments
received from the independent external peer reviewers and from the public is included in
Appendix A.
INTERNAL EPA REVIEWERS
Andrew Rooney, Ph.D.
Office of Research and Development
U.S. Environmental Protection Agency
Research Triangle Park, NC
Channa Keshava, Ph.D.
Office of Research and Development
U.S. Environmental Protection Agency
Research Triangle Park, NC
Allan Marcus, Ph.D.
Office of Research and Development
U.S. Environmental Protection Agency
Research Triangle Park, NC
Karen Hogan, M.S.
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
Lynn Flowers, Ph.D., DABT
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
EXTERNAL PEER REVIEWERS
James V. Bruckner, Ph.D.
University of Georgia
Robert A. Howd, Ph.D.
Office of Environmental Health Hazard Assessment (OEHHA)
California Environmental Protection Agency
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Ralph L. Kodell, Ph.D.
University of Arkansas for Medical Sciences
Janice Longstreth, Ph.D., DABT
The Institute for Global Risk Research, LLC
Michael I. Luster, Ph.D.
M. I. Luster and Associates, LLC
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1. INTRODUCTION
This document presents background information and justification for the Integrated Risk
Information System (IRIS) Summaries of the hazard and dose-response assessments of cis- and
trans-1,2-dichloroethylene (cis- and trans-1,2-DCE). Toxicological assessment of mixtures of
cis- and trans- 1,2-DCE is beyond the scope of this document. IRIS Summaries may include oral
reference dose (RfD) and inhalation reference concentration (RfC) values for chronic and other
exposure durations, and a carcinogenicity assessment.
The RfD and RfC, if derived, provide quantitative information for use in risk assessments
for health effects known or assumed to be produced through a nonlinear (presumed threshold)
mode of action. The RfD (expressed in units of mg/kg-day) is defined as an estimate (with
uncertainty spanning perhaps an order of magnitude) of a daily exposure to the human
population (including sensitive subgroups) that is likely to be without an appreciable risk of
deleterious effects during a lifetime. The inhalation RfC (expressed in units of mg/m3) 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). Reference
values are generally derived for chronic exposures (up to a lifetime), but may also be derived for
acute (<24 hours), short-term (>24 hours up to 30 days), and subchronic (>30 days up to 10% of
lifetime) exposure durations, all of which are derived based on an assumption of continuous
exposure throughout the duration specified. Unless specified otherwise, the RfD and RfC are
derived for chronic exposure duration.
The carcinogenicity assessment provides information on the carcinogenic hazard
potential of the substance in question and quantitative estimates of risk from oral and inhalation
exposure may be derived. 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 may be derived from the application of a
low-dose extrapolation procedure. If derived, the oral slope factor is a plausible upper bound on
the estimate of risk per mg/kg-day of oral exposure. Similarly, a plausible inhalation unit risk is
an upper bound on the estimate of risk per (ig/m3 air breathed.
Development of these hazard identification and dose-response assessments for cis- and
trans- 1,2-DCE has followed the general guidelines for risk assessment as set forth by the
National Research Council (1983). U.S. Environmental Protection Agency (U.S. EPA)
Guidelines and Risk Assessment Forum Technical Panel Reports that may have been used in the
development of this assessment include the following: Guidelines for the Health Risk Assessment
ofChemical Mixtures (U.S. EPA, 1986a), Guidelines for Mutagenicity Risk Assessment (U.S.
EPA, 1986b), Recommendations for and Documentation of Biological Values for Use in Risk
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Assessment (U.S. EPA, 1988), Guidelines for Developmental Toxicity Risk Assessment (U .S.
EPA, 1991), 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 ofInhalation Dosimetry (U .S. EPA, 1994b), Use of the Benchmark Dose Approach
in Health Risk Assessment (U .S. EPA, 1995), Guidelines for Reproductive Toxicity Risk
Assessment (U.S. EPA, 1996), Guidelines for Neurotoxicity Risk Assessment (U.S. EPA, 1998a),
Science Policy Council Handbook: Risk Characterization (U.S. EPA, 2000a), Benchmark Dose
Technical Guidance Document (U.S. EPA, 2000b), Supplementary Guidance for Conducting
Health Risk Assessment of Chemical Mixtures (U.S. EPA, 2000c), A Review of the Reference
Dose and Reference Concentration Processes (U.S. EPA, 2002b), Guidelines for Carcinogen
Risk Assessment (U.S. EPA, 2005a), Supplemental Guidance for Assessing Susceptibility from
Early-Life Exposure to Carcinogens (U.S. EPA, 2005b), Science Policy Council Handbook:
Peer Review (U.S. EPA, 2006a), and A Framework for Assessing Health Risks of Environmental
Exposures to Children (U.S. EPA, 2006b).
The literature search strategy employed for these compounds was based on the Chemical
Abstracts Service Registry Number (CASRN) and at least one common name. Any pertinent
scientific information submitted by the public to the IRIS Submission Desk was also considered
in the development of this document. The relevant literature was reviewed through May 2010.
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2. CHEMICAL AND PHYSICAL INFORMATION
There are two isomers of 1,2-DCE, the cis-isomer and the trans-isomer. The cis-isomer
is configured with the chlorine atoms on the same side of the C=C double bond, while in the
trans-isomer they are on opposite sides, resulting in different physical, chemical, and biological
properties (Figure 2-1). In general, 1,2-DCE has historically been used as a solvent for waxes,
resins, and acetylcellulose, in the extraction of rubber, and as a coolant in refrigeration plants
(National Library of Medicine [NLM], 2006). Currently, the trans-isomer is the only isomer
commercially available in the United States (American Conference of Governmental Industrial
Hygienists [ACGIH], 2001). Current uses for trans-1,2-DCE include its use as a degreasing
agent and as one component of formulated products used for precision cleaning of electronic
components. A small amount is used as a blowing agent for specialty foam.
\ /H \ . -Cl
C = C C = C
/ \ /
CI CI CI H
cis trans
Figure 2-1. Chemical structures of cis- and trans-l,2-dichloroethylene.
Some relevant chemical and physical properties of cis-1,2-DCE, trans-1,2-DCE, and a
mixture of both isomers are listed in Table 2-1 (NLM, 2006; Agency for Toxic Substance and
Disease Registry [ATSDR], 1996). Exposure to the 1,2-DCEs may occur after the chemicals are
released to the environment from industrial emissions, leaching from landfills, or evaporation
from wastewater streams. The estimated half-lives of cis- and trans-1,2-DCE in air are 12 and
5 days, respectively. Volatilization is the major fate process when the chemicals are released to
surface water, with an estimated half-life of about 3-6 hours. In soil, 1,2-DCE may leach
through the subsurface and contaminate groundwater. The chemical may also be found in
groundwater due to anaerobic degradation of more highly chlorinated chemicals, such as
trichloroethylene and tetrachloroethylene (ATSDR, 1996). Although no degradation occurs in
sterile microcosms, anaerobic biodegradation of the cis-isomer to chloroethane and vinyl
chloride and biodegradation of the trans-isomer to vinyl chloride alone have been reported
(Barrio-Lage et al., 1986, as reported in ATSDR, 1996). The cis-isomer is degraded more
readily than the trans-isomer (Barrio-Lage et al., 1986, as reported in ATSDR, 1996). The rates
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of degradation of both isomers are dependent on the availability of an electron donor and the
presence of active anaerobes.
Table 2-1. Properties of the 1,2-dichloroethylene isomers and their mixture
Descriptor
cis-Isomer
trans-Isomer
Mixture
CAS name
cis-1,2-Dichloroethy lene
trans-1,2-Dichloroethy lene
1,2-Dichloroethylene
CAS number
156-59-2
156-60-5
540-59-0
Primary synonyms
cis-l,2-Dichloroethene,
1,2-cis-dichloroethylene, cis-
acetylene dichloride,
cis-l,2-DCE
trans-1,2-Dichloroethene,
1,2-trans-dichloroethylene,
trans-acetylene dichloride,
trans-l,2-DCE
1,2-Dichloroethene,
acetylene dichloride, 1,2-DCE
Chemical formula
C2H2C12
Molecular weight
96.95
Boiling point
60.1°C at 760 mm Hg
48.7°C at 760 mm Hg
Approximately 55°C
Melting point
-80°C
-49.8°C
-50°C
Specific gravity
1.2837 @20°C/4°C
1.2565 @20°C/4°C
Approximately 1.28
Vapor pressure
2.00 1 02 mm Hg @ 25°C
3.31 102mmHg@25°C
2.01 102 mm Hg @ 25°C
Solubility
Miscible with alcohol, ether,
acetone, benzene,
chloroform; solubility in
water = 6.41 g/L @ 25°C
Miscible with alcohol, ether,
acetone, benzene,
chloroform; solubility in
water = 4.52 g/L @ 25°C
Miscible with alcohol, ether,
acetone, benzene, chloroform;
solubility in water = 3.5 g/L
@25°C
Odor
Ethereal, slightly acrid, sweet, pleasant
Odor threshold (air)
NAa
0.085 ppm
NAa
Partition coefficients:
Log Kow
Log Koc
1.86
1.69 (estimated)
2.06
1.56 (estimated)
2.00
NA
Henry's law constant
4.08 10"3 atm-m3/mol
@24.8°C
9.28 10"3 atm-m3/mol
@24.8°C
Approximately 4.08 10'3
atm-m3/mol @ 24.8°C
Flash point
2-4°C
2°C
2°C
Conversion factor
1 mg/m3 = 0.252 ppm; 1 ppm = 3.97 mg/m3
"NA = not available.
Sources: NLM (2006); ATSDR (1996).
The 1,2-DCEs are highly flammable; the vapors may explode when heated or exposed to
an open flame. Combustion by-products of 1,2-DCE include hydrogen chloride and phosgene
(NLM, 2006).
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3. TOXICOKINETICS
3.1. ABSORPTION
3.1.1. Oral
No studies were identified that examined oral absorption of either cis- or trans- 1,2-DCE.
3.1.2. Inhalation
Considerable work has been done in this area to determine parameters applicable to
physiologically based pharmacokinetic (PBPK) modeling. Filser and Bolt (1979) studied the
uptake of cis- and trans-1,2-DCE in male Wistar rats (250 g). The animals were exposed to
various initial concentrations of the substances in a closed chamber, and the decline of the
substance with time was monitored by gas-liquid chromatography. The authors did not report
initial airborne concentrations, but, judging from time zero in their graphs (Filser and Bolt,
1979), they ranged from about 20 to 1,000 ppm. Plots by the authors of chamber concentration
vs. time displayed two or three phases, depending on the initial concentration. A first phase of
rapid decline of gas concentration in the chamber represented the initial uptake of gas and its
equilibration with the chamber atmosphere that lasted about 2 hours for cis-1,2-DCE and
1.5 hours for trans-1,2-DCE. The second phase was typical of a first-order metabolic
disappearance of the substance when initial gas concentrations were sufficiently low but took on
the characteristics of zero-order elimination with high gas concentrations, saturating the
metabolic capacity of the animals in the chamber. A third phase was seen in cases with high
initial gas concentration, where, with time, the concentration in the chamber fell low enough to
no longer saturate the metabolic enzymes, displaying first-order disappearance characteristics
from there on.
Filser and Bolt (1979) analyzed the disappearance curves mathematically and established
saturation points for both cis- and trans-1,2-DCE (i.e., gas concentrations at which the metabolic
capacities of the experimental animals became saturated and gas disappearance from the
chamber changed from first to zero order). These values were given as 20 ppm for cis-1,2-DCE
and 15 ppm for trans-1,2-DCE. The shorter equilibration time and lower saturation point
concentration for trans- 1,2-DCE were interpreted by the authors to indicate slower metabolic
removal of trans-1,2-DCE, as compared with cis-l,2-DCE.
In vitro gas/blood distribution data indicate that trans-1,2-DCE is less soluble in blood
than cis-l,2-DCE, which would suggest that inhalation uptake of trans-l,2-DCE is less than that
of cis-1,2-DCE. Eger et al. (2001) conducted experiments with male Sprague-Dawley rats and
found that the alveolar concentration of trans- 1,2-DCE required to induce anesthesia in 50% of
the animals was about twice as high as that of cis-l,2-DCE. Gargas et al. (1989, 1988) published
blood:air partition coefficients of 21.6 and 9.58 in the rat and 9.85 and 6.04 in humans, for cis-
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and trans-1,2-DCE, respectively. Sato and Nakajima (1987) also reported values of 9.2 and 5.8
for cis- and trans- 1,2-DCE, respectively, but the species in which these values were obtained was
not specified. A comparison of values given by Gargas et al. (1989) for humans, rats, and, to a
lesser extent, mice indicated that, for most chlorinated aliphatics examined, human blood had
only about one-half the affinity of that measured in rat and mouse blood. Equilibrium constants
for inhalation uptake over exhalation elimination calculated by Filser and Bolt (1979) showed
the same approximate 2:1 ratio (i.e., 20 and 11.5 for cis- and trans-1,2-DCE, respectively).
Therefore, several studies support the conclusion that cis- and trans-DCE are absorbed relatively
quickly by the lungs and in a ratio of 2:1.
Andersen et al. (1980) used male F344 rats (180-280 g) to conduct inhalation
experiments with trans-1,2-DCE, using a closed chamber system with gas recirculation. The
results were similar to those obtained by Filser and Bolt (1979) in that the uptake of
trans-1,2-DCE leveled off after about 2 hours, with about 40-60% of the gas remaining in the
chamber at exposure concentrations of 10,000, 1,000, and 30 ppm. Using a model developed
earlier (Ramsey and Andersen, 1984), Gargas et al. (1988) calculated the maximum substrate
turnover velocity (Vmax) values for pulmonary uptake of both cis- and trans-1,2-DCE of
30.9 jamol/kg-hour (3 mg/kg-hour) for rats (likely this value is true only for trans-1,2-DCE, as
estimated by Andersen et al. [1980]). Both Andersen et al. (1980) and Filser and Bolt (1979)
pointed out that results obtained in a given rat strain could not be extrapolated to another strain.
In addition, Gargas et al. (1990) pointed out that the uptake of gaseous cis- or trans-1,2-DCE
could be approximated only by using a model that corrected for suicide inhibition of the
cytochrome P450 (CYP450) enzymes that metabolize these agents.
In an experiment using isolated perfused liver from female Wistar rats and exposing the
perfusate to cis- or trans-1,2-DCE in the gas phase, Bonse et al. (1975) found that, at a given
concentration in the gas phase, trans-1,2-DCE attained less than one-half the concentration of
cis-1,2-DCE in liver, which was attributed by the study authors in part to inhibition of CYP450
by the trans-isomer.
3.1.3. Dermal
No studies were located that investigated the dermal uptake of cis- or trans-1,2-DCE
either as a liquid or from the vapor phase. Pleil and Lindstrom (1997) conducted experiments
with human volunteers who were exposed to cis-1,2-DCE via showering with contaminated
water (informed consent from the volunteers and institutional approval were obtained).
Appearance of the substance in exhaled air, collected as single breaths of 1 L volume, was
monitored by gas chromatography/mass spectrometry (GC/MS). Samples of
microenvironmental air from the exposure area and control samples of inspired air after the
exposure were also collected and analyzed using the same equipment. In two separate
experiments, two volunteers were exposed under a shower for 10 minutes each to an
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environment with 125 and 83.9 jag/m3 cis-l,2-DCE in the air and 28.4 and 20.4 ju.g/L in the
water, respectively. Samples of exhaled air and blood were collected for 30 minutes after
exposure. Samples of air from the exposure area and control samples of inspired air after the
exposure were also collected and analyzed. The authors calculated total exposure doses of 1.19
and 2.34 jag, while the corresponding maximum blood concentrations were 0.25 and 0.18 ju.g/L,
respectively. The authors considered these values as indicative of efficient absorption of
cis-l,2-DCE with mixed inhalation and dermal exposure.
The interim report Dermal Exposure Assessment (U.S. EPA, 1992) gives a dermal
2
permeability coefficient, Kp, of 1.0 10 cm/hour for human skin. This value references uptake
from aqueous solution, but there is uncertainty in whether it refers to cis-l,2-DCE,
trans-1,2-DCE, or mixed isomers. By using a formula for dermal absorption of liquids proposed
by Potts and Guy (1992):
log Kp = -2.7 + 0.71 log Kqw - (0.0061 molecular weight)
_2 _2
Kp values of 1.07 10 and 1.55 10 cm/hour are obtained for cis-and trans-1,2-DCE,
respectively. Such values indicate efficient dermal absorption, comparable to that of lipophilic
aromatics, such as cresols, chlorophenols, orhexanol (U.S. EPA, 1992). These values also
suggest that the higher lipophilicity of trans-1,2-DCE may increase its dermal absorption.
3.2. DISTRIBUTION
No in vivo studies pertaining to organ and/or tissue distribution of cis- or trans-1,2-DCE
have been reported in the literature. However, Bonse et al. (1975) reported that in an experiment
using isolated perfused liver from female Wistar rats, at equimolar concentrations in the
perfusate (with chlorinated ethylenes added as vapors at constant rates that allowed for steady
state conditions of substrate uptake and conversion), uptake for cis-1,2-DCE was about 3 times
faster than for trans-1,2-DCE. Gargas et al. (1988) reported tissue:air partition coefficients (at
37°C) for rat (species not specified) tissues in vitro that are compiled in Table 3-1. These data
provide further support, albeit indirectly, that the trans-isomer is likely to be taken up less
efficiently by mammalian tissues than the cis-isomer. Furthermore, if the previously discussed
relationship between rat and human blood:air partition coefficients is assumed to be predictive,
then the extent of uptake of the two isomers into human tissues would be roughly half that of the
corresponding rat tissues shown below (Table 3-1).
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Table 3-1. Tissue:air partition coefficients of the 1,2-dichloroethylene
isomers in the rat (in vitro)
Tissue
Partition coefficient
cis-l,2-Dichloroethylene
trans-l,2-Dichloroethylene
Blood
21.6
9.58
Liver
15.3
8.96
Muscle
6.09
3.52
Fat
227
148
Source: Gargas et al. (1988).
3.3. METABOLISM
Henschler and Bonse (1977) proposed a metabolic scheme for cis- and trans-1,2-DCE,
shown in Figure 3-1. Metabolism of 1,2-DCE is initially catalyzed by hepatic CYP450, and
limited experimental evidence suggests that CYP2E1 may be the primary pathway for
metabolism of cis-or trans-1,2-DCE in the rat (Costa and Ivanetich, 1984, 1982). Studies
suggest that the metabolism of 1,2-DCE involves epoxidation of the ethylene double-bond,
forming dichlorinated epoxides, which can undergo a nonenzymatic rearrangement (Costa and
Ivanetich, 1984, 1982) and produce several metabolites. Following rearrangement of the
intermediate epoxide, reduction of the resulting dichloroacetaldehyde to dichloroethanol may be
catalyzed by alcohol dehydrogenase (ADH). Studies by Costa and Ivanetich (1984, 1982)
provide evidence that dichloroacetaldehyde is the predominant metabolite of CYP450, which is
extensively converted to dichloroethanol and dichloroacetate by dehydrogenases present in
hepatocytes. Oxidative dechlorination of the minor metabolite, dichloroacetic acid (DCA) to
glyoxylate is catalyzed by glutathione S-transferase zeta (GSTZ) (Costa and Ivanetich, 1982).
The enzymes involved in further biotransformation of 1,2-DCE metabolites have not been
characterized.
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epoxide
(oxirane)
CYP2E1
CYP3A4
trans
1,2-dichloroethylene
CIS
spontaneous?
o®
H
C—CI
H—C
H—C—C—CI
OH H
2,2-dichloroethanol
alcohol
dehydrogenase
c—c—CI
dichloroacetic acid
di chl oroacetal dehy de
c—c—CI
H
HO
glycolic acid C C OH
. HO J,
glutathione
S-transferase
NADH
/?
\
H
glyoxylic acid
alanine
c—
c—c—NH
HO
HO
FAD
glycine
c—c
HO OH
oxalic acid
Sources: Adapted from U.S. EPA, 2003; Henschler and Bonse, 1977.
Figure 3-1. Proposed metabolic scheme for cis- and trans-
1,2-dichloroethylene.
3.3.1. Metabolism in Animals
Bonse et al. (1975) studied the metabolism of several chlorinated C2-compounds, among
them cis- and trans-1,2-DCE, in isolated perfused livers from female Wistar rats (170-230 g).
The perfusate was supplemented with various concentrations of the compounds in the gas phase.
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Concentrations of 1,2-DCE and metabolites in liver tissue and perfusate were monitored with gas
chromatography (GC). Bonse et al. (1975) expected to find epoxide isomers (oxiranes) as the
primary metabolites of 1,2-DCE. Both cis- and trans-epoxides are unstable, however, and were
shown to rearrange spontaneously to form dichloroacetaldehyde, which was then readily
converted to DCA and 2,2-dichloroethanol. Levels of enzymes from liver cells in the perfusate
(i.e., lactate dehydrogenase [LDH], aspartate aminotransferase [AST] (glutamate oxaloacetate
transaminase [SGOT], or alanine aminotransferase [ALT] (glutamate pyruvate transaminase
[SGPT]) increased with time. The authors interpreted these findings to be indicative of liver
damage. Higher activity levels of these enzymes were detected in the cis-1,2-DCE perfusate
compared with corresponding activities in the trans-l,2-DCE perfusate.
For cis-l,2-DCE, Bonse et al. (1975) identified 2,2-dichloroethanol as the major
metabolite and DCA as a minor metabolite; for trans-1,2-DCE, only small amounts of these two
metabolites could be identified. Uptake of cis- 1,2-DCE in liver tissue was demonstrated to be at
least 2 times faster than the uptake of trans-1,2-DCE, which may partially account for the lower
concentrations of metabolites of trans-1,2-DCE in liver tissue. The authors also observed that
the amount of metabolites in liver tissue did not correlate with tissue uptake for these two
substances, thus confirming differing rates of metabolism as well.
Leibman and Ortiz (1977) studied the metabolism of the 1,2-DCE isomers in rat liver
homogenate supernatants (9000 g supernatant fraction [S9]) and suggested metabolic schemes
for the chlorinated ethylenes. For 1,2-DCE (isomer not specified), they proposed the same
sequence of events that Bonse et al. (1975) and Henschler and Bonse (1977) had proposed,
although Leibman and Ortiz (1977) were not able to experimentally identify DCA as a
metabolite of 1,2-DCE. They were, however, able to mechanistically describe the chemical
rearrangement via an epoxide intermediate that explains the formation of asymmetrically
substituted DCA from both symmetrically substituted 1,2-DCE isomers.
Costa and Ivanetich (1982) investigated the metabolism of the chlorinated ethylenes ex
vivo, using the S9 fraction from the livers of male Long-Evans rats. Some of the rats were
pretreated with enzyme inducers, such as p-naphthoflavone or phenobarbital, prior to sacrifice
and microsome preparation. The DCEs were added as ethanolic solutions to the microsomal
preparations. Metabolite identification was performed by gas/liquid chromatography. Following
treatment with both cis- and trans-1,2-DCE, measurable amounts of 2,2-dichloroethanol and
dichloroacetaldehyde were detected, with trans-l,2-DCE yielding about 25% the amount of
2,2-dichloroethanol that cis-1,2-DCE yielded. DCA was also formed from both substances,
although the amount was about 6 times less from trans-1,2-DCE than from cis-1,2-DCE. The
authors could not identify any of the dechlorination metabolites of cis- or trans-1,2-DCE shown
in the metabolic scheme in Figure 3-1, possibly because the S9 mix used did not contain
considerable glutathione S-transferase (GST) activity. Overall, the authors estimated the in vitro
CYP450-mediated metabolism of cis-1,2-DCE to be 4 times that of trans-1,2-DCE. Suicide
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inhibition of the CYP450 activity via covalent binding of a reactive intermediate to the heme
moiety was also observed; this propensity to bind heme was independent of the enzymatic
degradation of any DCE substrate. However, it was noted that substances whose epoxides
rearranged to an aldehyde (e.g., cis- or trans-1,2-DCE) bound to heme, while those that formed
acylchlorides (e.g., 1,1-DCE) did not bind to heme.
Costa and Ivanetich (1982) also found that metabolite binding to hepatic microsomes
induced a spectral shift indicative of binding to the active center of CYP450. Hanes plots of
substance concentration vs. spectral shift revealed two binding constants, suggesting that more
than one CYP450 isoform in the microsomes was involved. Pretreatment of the animals with
phenobarbital increased the affinity of the substrate for the low affinity binding site but did not
affect that of the high affinity binding site. Biphasic Hanes plots were observed with
cis-l,2-DCE when either phenobarbital-noninduced or -induced liver microsomes were used;
with trans-1,2-DCE, the plots were monophasic unless phenobarbital-induced microsomes were
used. Treatment with the nonspecific inhibitors carbon monoxide and SKF-525A suppressed the
formation of dichloroacetaldehyde or 2,2-dichloroethanol from both cis- or trans-1,2-DCE.
However, while metyrapone, a specific CYP3A4 inhibitor, was minimally effective in inhibiting
metabolism of cis-1,2-DCE, it was most effective in suppressing trans-1,2-DCE metabolism.
Accordingly, pretreatment with phenobarbital, which induces CYP3A4, among others, increased
the metabolism of trans-l,2-DCE more than cis-l,2-DCE. Therefore, CYP3A4 may play a role
in the metabolism of 1,2-DCE, but the exact nature and extent of this role need to be further
characterized. These researchers also conducted experiments that suggested that the formation
of 2,2-dichloroethanol from dichloroacetaldehyde was catalyzed by an NADPH-dependent ADH
that contaminated their microsomal preparations. Filser and Bolt (1980) also reported that
disulfiram, an ADH inhibitor, caused changes in the response of rats to inhaled trans-1,2-DCE
that were suggestive of ADH involvement in its metabolism.
In a subsequent publication, Costa and Ivanetich (1984) used hepatocytes from male
Long-Evans rat livers to study the metabolism of cis- and trans- 1,2-DCE. After incubation, cells
were destroyed with sulfuric acid and sodium tungstate and the supernatants extracted for
gas/liquid chromatography. Isolated rat hepatocytes metabolized cis-1,2-DCE primarily to
2,2-dichloroethanol (2.4 nmol/106 cells/10 minutes) with the formation of smaller amounts of
DCA (0.3 nmol/106 cells/10 minutes) and dichloroacetaldehyde (0.04 nmol/106 cells/
10 minutes). No other chlorinated metabolites were produced from cis-l,2-DCE in measurable
amounts. The metabolism of trans-1,2-DCE in isolated rat hepatocytes gave rise to DCA
(0.05 nmol/106 cells/10 minutes), traces of dichloroacetaldehyde (0.008 nmol/106 cells/
10 minutes), and 2,2-dichloroethanol (0.01 nmol/106 cells/10 minutes). This study by Costa and
Ivanetich (1984) showed that the total amount of trans-l,2-DCE metabolized was 8-25 times less
than that of cis-1,2-DCE, yielding only small amounts of DCA and trace amounts of
2,2-dichloroethanol and dichloroacetaldehyde.
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Costa and Ivanetich (1984) estimated Michaelis constant (Km) values of 0.67 mM for the
formation of DC A from cis-l,2-DCE (the metabolic yield with trans-1,2-DCE was too small for
rate estimation), 2.15 mM for the formation of dichloroacetaldehyde, and 2.55 mM for the
formation of 2,2-dichloroethanol when phenobarbital-induced hepatocytes were used. These
researchers also incubated the known metabolites of cis- and trans-1,2-DCE with cultured
hepatocytes. They observed that DCA and dichloroacetaldehyde were largely (—90%)
metabolized within 60 minutes in phenobarbital-induced hepatocyte culture. Degradation of
dichloroacetaldehyde yielded primarily DCA, with the formation of a small amount of
2,2-dichloroethanol.
The question of further metabolism of 2,2-dichloroethanol or DCA has not been
investigated in the context of cis- and/or trans-1,2-DCE metabolism. Barton et al. (1995), in an
attempt to model the toxicokinetics of chloroethylene mixtures, found that exposing male
Sprague-Dawley rats to 40 ppm trans-1,2-DCE for 4.5 hours did not affect nonprotein sulfhydryl
content (essentially reduced glutathione [GSH]) in the livers. This could be seen as an indication
that metabolites of trans-l,2-DCE do not undergo GSH conjugation to any major extent.
Similarly, Dowsley et al. (1999) were not able to detect any GSH conjugates of 1,1-DCE in
experiments with microsomal preparations from female CD-I mice, although 1,1-DCE forms the
same metabolite, dichloroacetaldehyde, as the 1,2-isomers. According to a metabolic scheme
provided in that paper, formation of acetyl chloride or its derivative would be a prerequisite for
GSH conjugation. McMillan (1986) found slight yet statistically significant reductions in
hepatic GSH concentrations following high single doses of cis- or trans-l,2-DCE (10% reduction
following 4.4 g/kg trans-1,2-DCE orally, 22% reduction following 1.9 g/kg trans-1,2-DCE
intraperitoneally [i.p.], and 17% reduction following 2 g/kg cis-l,2-DCE i.p.).
DCA is metabolized via oxidative dechlorination, a cytosolic process that does not
involve CYP450 but instead GSH, NADPH, and GSTZ (U.S. EPA, 2003). The resulting
metabolite is glyoxylate, which can undergo further oxidation to oxalate, reduction to glycolic
acid, and transamination to glycine with subsequent formyl group transfer to form serine. This
pathway is also presented in Figure 3-1. DCA stimulates peripheral glucose utilization and
therefore has been proposed as an agent for treatment of several metabolic disorders, including
diabetes and myocardial ischemia (Stacpoole, 1989). Oxalate can form insoluble crystals of
calcium oxalate that can cause kidney damage. The ultimate products of glyoxylate
biotransformation, glycine and serine, are utilized in protein synthesis.
Nakajima (1997) presented some evidence that both cis- and trans-l,2-DCE are
metabolized by CYP2E1. By using microsomal preparations from untreated, fasted, or ethanol-
pretreated rats, they found a twofold increase in the rate of metabolism (RAM) of cis-1,2-DCE in
microsomes from fasting rats and a threefold increase in its metabolism in microsomes from
ethanol-treated rats. Fasting and dietary ethanol are widely known to induce the activity of
CYP2E1 (Cederbaum, 2006; Wan et al., 2006). A comparatively low RAM of trans-l,2-DCE by
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microsomes from ethanol-treated rats was reported, which was not measurable using microsomes
from untreated or fasted rats. The results obtained with ethanol-induced liver microsomes
provide inferential evidence that CYP2E1 is involved in the metabolism of 1,2-DCE.
In summary, the metabolism of 1,2-DCE is initially catalyzed by hepatic CYP450. The
available evidence suggests that CYP2E1 may be the primary pathway for metabolism of cis- or
trans-l,2-DCE in the rat (Nakajima, 1997; Costa and Ivenetich, 1984, 1982). The metabolism of
cis- or trans-1,2-DCE is thought to involve spontaneous formation of epoxides, which can
rearrange to produce several metabolites (Costa and Ivenetich, 1984, 1982). These epoxides are
unstable and have been shown by Bonse et al. (1975) to rearrange spontaneously to form
dichloroacetaldehyde, which is then readily converted to DCA and 2,2-dichloroethanol. For cis-
1,2-DCE, Bonse et al. (1975) identified 2,2-dichloroethanol as the major metabolite and DCA as
a minor metabolite; for trans-1,2-DCE, only small amounts of these two metabolites could be
identified. There is also some evidence that both cis- and trans-1,2-DCE are metabolized by
CYP2E1 (Nakajima, 1997, Cederbaum, 2006; Wan et al., 2006).
The further metabolism of 2,2-dichloroethanol or DCA has not been investigated in the
context of cis- and/or trans-1,2-DCE metabolism. However, a study by Barton et al. (1995)
indicates that metabolites of trans-l,2-DCE do not undergo GSH conjugation to any major
extent.
3.3.2. Metabolism in Human Preparations In Vitro
Doherty et al. (1996) investigated the potential clastogenic activity of several chlorinated
hydrocarbons, among them 1,2-DCE (likely a mixture of both isomers), using several human cell
lines with variable CYP450 enzyme expression profiles. Their findings suggest that a direct-
acting genotoxic effect without the need for metabolic activation is possible and the production
of a metabolite that was less genotoxic than the parent compounds is also possible.
3.3.3. CYP2E1 Inactivation by 1,2-DCE
Cis- and trans-1,2-DCE are metabolized by microsomal oxidation (Filser and Bolt, 1979).
In vitro studies indicate that cis- and trans-1,2-DCE cause a loss of hepatic microsomal
cytochrome P450 and heme, thus suggesting cytochrome P450 inactivation by reaction products
(Costa and Ivanetich, 1982). Lilly et al. (1998) reported that cis- and trans-l,2-DCE inactivated
CYP2E1 in rats. As inhibitors of CYP2E1, an isoform of P450 that plays a role in the
bioactivation of a number of volatile organic compounds (VOCs) and other chemicals (Brady et
al, 1991; Guengerich et al., 1991; Nakajima et al., 1990, 1991; Seaton et al., 1994), sufficiently
high exposures to 1,2-DCE are expected to inhibit the CYP-mediated activation of a wide variety
of VOCs. For example, Barton et al. (1995) discovered that pre-exposure of rats to 40 ppm
trans-l,2-DCE for 1.5 hours resulted in marked inhibition of trichloroethylene and vinyl chloride
metabolism by competitive inhibition.
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3.4. ELIMINATION
Information on the elimination of cis- or trans-1,2-DCE or its metabolites is limited.
However, Pleil and Lindstrom (1997) have estimated elimination rate constants for the
disappearance from human blood of certain halogenated volatile organic compounds (VOCs),
including cis-l,2-DCE. Estimates were based on decay of exhaled breath concentrations
following a 10-minute shower exposure to contaminated water and published blood/air partition
coefficients for the VOC in question. Two volunteers (recruited with informed consent in
compliance with local institutional standards) were exposed in separate showering episodes, in
which estimated total absorbed doses of cis-l,2-DCE were 1.19 and 2.34 jag, respectively. The
kinetics of elimination of parent compound from breath suggested the existence of two biological
distribution compartments, which were presumed to represent the blood and "highly perfused
tissues" (e.g., liver). In the first fast-elimination compartment (presumed to represent
disappearance of cis-l,2-DCE from the blood), elimination half-lives of 0.82 and 2.37 minutes
were estimated in the two subjects; corresponding half-lives in the slower, highly perfused tissue
compartment were 8.96 and 29.33 minutes, respectively. These limited data suggest the
potential for variability in the elimination of cis-l,2-DCE in humans.
Considering the metabolic fates of the various possible metabolites of 1,2-DCE, it may be
assumed that whatever portion of dichloroethanol is not transformed to DCA will be ultimately
exhaled. For dichloroacetaldehyde and DCA, the IRIS Toxicological Review for Dichloroacetic
Acid (U.S. EPA, 2003) provides some useful information. Accordingly, glyoxylate formed via
GSTZ is ultimately broken down to carbon dioxide or oxidized to oxalate, which is excreted in
the urine. Dechlorination products, such as monochloroacetic acid, also are said to exist, but, for
the case of cis- or trans-1,2-DCE, this is at odds with the findings of Costa and Ivanetich (1984,
1982) who could not detect dechlorination products of 1,2-DCE in vitro with rat microsomes or
hepatocytes. A possible explanation is that, given the comparatively poor uptake and slow
metabolism of cis- and trans-1,2-DCE, tissue levels of DCA never become high enough to allow
for any measurable dechlorination reaction to occur.
3.5. PHYSIOLOGICALLY BASED PHARMACOKINETIC MODELS
A toxicokinetic description of distribution and elimination of inhaled cis- and
trans-l,2-DCE in rats was reported by Filser and Bolt (1979), who analyzed experimental data by
using a simplified compartmental model. However, their interpretation of the metabolic
clearance of 1,2-DCE failed to address the inactivation of CYP2E1 that had been observed both
in vivo and in vitro by Freundt and Macholz (1978) and later quantified in hepatic microsomal
preparation in vitro by Costa and Ivanetich (1982). This metabolic inactivation phenomenon
also complicated the fitting of experimental data, which had been obtained with different
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concentrations of cis- and trans-1,2-DCE vapors in a closed gas chamber, to atypical PBPK
model for VOCs, using only the metabolic constants, Vmax and Km-
Gargas et al. (1990) updated the PBPK model for rats with an algorithm that described
CYP2E1 suicide inhibition-resynthesis. In this algorithm (Clewell and Andersen, 1987), the rate
of enzyme inactivation was proportional to a second order rate constant (kd), multiplied by the
square of the initial RAM, thereby representing the reaction of free metabolite(s) with the
enzyme-substrate complex. The algorithm also included a term for the zero-order rate of enzyme
resynthesis (ks) during exposure. Subsequently, the PBPK model for cis- and trans-l,2-DCE in
rats was extended by Lilly et al. (1998) to describe quantitatively the mechanisms of both
suicidal inhibition of CYP2E1 by metabolic intermediate(s) and CYP2E1 resynthesis. This
algorithm ("Model 1" in Lilly et al., 1998) required four parameters, or kinetic constants: Vmaxc
(maximum rate of metabolism), Km (pseudo-Michaelis constant), kd (inhibition constant), and
Kde (enzyme degradation constant). The model-estimated kinetic constants VmaxC and Km were
4.53 mg/hour/kg and 0.19 mg/L for cis-l,2-DCE and 4.27 mg/hour/kg and 0.08 mg/L for
trans-1,2-DCE, respectively, with cis-l,2-DCE metabolite(s) being less potent inhibitor(s) of
CYP2E1 (kd = 2.07 [mg/hour] [hour]"1) than the metabolite(s) of trans-l,2-DCE
(kd = 496 [mg/hour] [hour]"1), under a similar enzyme degradation constant (Kde about
0.025 [hour]"1) (Lilly et al., 1998).
The PBPK model structure (Figure 3-2) consists of five dynamic tissue compartments
representing the lungs, fat, rapidly perfused tissues, slowly perfused tissues, and liver. All
perfusion-limited tissue compartments are linked through blood flow, following an anatomically
accurate, typical, physiologically based description (Lilly et al., 1998).
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Cv
QP.Ca
Cvf
Qf.Pf
Qr.Pr
Cvr
Qs.Ps
Cvs
Cvl
RAM
Chemical Metabolism
Slowly Perfused
Tissues
Rapidly Perfused
Tissues
Alveolar
Space
Lung
Blood
and Enzyme Inhibition
Source: Lilly et al. (1998) (reproduced with permission of Springer Verlag,
Heidelberg/New York).
Figure 3-2. PBPK model for cis- and tr an s-1,2- dichlo roethylene in rats.
Briefly, because cis- and trans-1,2-DCE are retained by the tissue(s) in each compartment
according to their tissue/blood partition coefficients (measured in vitro by Gargas et al., 1988),
the concentrations of both chemicals in venous blood (leaving the tissue) are lower than those in
arterial blood during the equilibration phase. Therefore, the rate of change in the amount of
either chemical in each tissue compartment (,) is given by the difference between concentration
in blood entering (Ca) and exiting (CV1) the tissue, multiplied by the blood flow (Q,). The
differential equations for each tissue compartment (except lungs) are integrated over time, giving
the amounts of cis- or trans- 1,2-DCE present in the tissue. Because the partition coefficient (P,)
and the actual volume of each tissue are known from the literature (Ramsey and Andersen,
1984), concentrations of cis- or trans-1,2-DCE in each tissue can be calculated over time.
For the lung compartment with two mass inputs (mixed venous blood and inhaled air)
and two outputs (arterial blood and exhaled air), at steady state the amount of either chemical in
alveolar air is in equilibrium with the amount in lung blood, and, thus, concentrations of cis- or
trans-1,2-DCE in arterial blood can be calculated from the simple mass balance equations, taking
into account the alveolar ventilation rate and the rate of blood flow through the lung (equal to
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cardiac output), both known from the literature (Ramsey and Andersen, 1984). For the liver
compartment, with mass input from blood and two outputs (venous blood and metabolism;
biliary excretion was not considered), the chemical mass transfer is given by the difference
between concentrations in portal (Ca) and venous (Cvi) blood multiplied by hepatic blood flow
(Qi) and corrected for metabolic clearance of cis- or trans-1,2-DCE.
The rates of metabolism (RAMs) (see Figure 3-2) of cis- and trans-1,2-DCE are
calculated from the Michaelis-Menten equation (using "metabolic capacity remaining" instead of
initial velocity Vmaxo) and subtracted from the rate of change in chemical mass in the liver. Rates
of change of inhibitable metabolism (RVMT) (see Figure 3-2), under the assumption that a
reactive metabolite reacts with enzyme-substrate complex ("Model 1" in Lilly et al., 1998), can
be calculated also from the Michaelis-Menten equation with a negative inhibition term (rate
constant -kd multiplied by RAM), whereas rates of change of metabolism due to enzyme
resynthesis can be calculated by a zero order term, multiplying Vmax by Kde (Bae et al., 2005;
Lilly et al., 1998).
A simplified scheme of the mass flow in the PBPK model for cis- and trans-1,2-DCE is
shown in Figure 3-2, according to Lilly et al. (1998). This model was calibrated with data
obtained in closed-chamber gas uptake studies with rats, as reported by Gargas et al. (1990).
From four different algorithms tested by Lilly et al. (1998), "Model 1," which assumes that
reactive metabolite(s) of cis- and trans- 1,2-DCE inactivate the CYP2E1 enzyme-substrate
complex, gave the best approximation of experimentally obtained data (Bae et al., 2005). One
could extrapolate the model to humans by allometrically scaling Vmax in the absence of exposure
and the resynthesis rate for CYP2E1, while assuming that the molecular rate of suicide inhibition
is the same for human and rat CYP2E1. However, in the absence of human data with which to
validate or calibrate this model, such an extrapolation would involve considerable uncertainty,
much greater than cases without suicide inhibition. (The data on human exhalation subsequent
to exposure in a shower is likely most sensitive to the parameters describing respiration, cardiac
output, and the blood: air partition coefficient, and these data are expected to provide little
information on metabolic rates.) Therefore, such extrapolation of the model is not attempted in
this assessment.
Since this PBPK model was not calibrated with human data, it cannot be scaled
allometrically to humans, whose liver CYP2E1 activity, resynthesis rate, and sensitivity to
inhibition differ from those in rats. Given the current state of knowledge, this PBPK model is
not useful for estimating the human equivalent dose from the available animal data for cis- or
trans-1,2-DCE. No other valid PBPK models of cis- or trans-1,2-DCE were identified.
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4. HAZARD IDENTIFICATION
4.1. STUDIES IN HUMANS—EPIDEMIOLOGY, CASE REPORTS, CLINICAL
TRIALS
There are limited data available from studies of effects of 1,2-DCE in humans. In an
early study (Lehmann and Schmidt-Kehl, 1936), the threshold for odor detection of
trans-1,2-DCE by two human subjects was reported to be 280 ppm (1,100 mg/m3). Slight eye
irritation occurred after 30 minutes inhalation exposure to 830 ppm (3,300 mg/m3), while at
exposure concentrations of 1,200 ppm (4,800 mg/m3) to 2,200 ppm (8,800 mg/m3) for 5—
10 minutes, both subjects reported symptoms of nausea, drowsiness, fatigue, vertigo, and a
feeling of intracranial pressure. Hamilton (1934, as cited in Dow Chemical Co. [Dow, 1962])
reported that a worker who entered a vat containing rubber dissolved in 1,2-DCE of unknown
isomeric composition was found dead the following morning. The exposure concentration and
duration of exposure were unknown. A human threshold limit value (TLV) of 200 ppm for cis-
or trans-1,2-DCE and mixtures of the two isomers has been established by ACGIH (2001).
4.2. SHORT-TERM, SUBCHRONIC, AND CHRONIC STUDIES AND CANCER
BIOASSAYS IN ANIMALS—ORAL AND INHALATION
A number of studies in animals have investigated the short-term and subchronic toxicity
of cis- or trans-1,2-DCE by either the oral or inhalation route. Presented below are summaries of
these investigations. No chronic studies for cis- or trans-1,2-DCE or their mixtures were
identified. No cancer studies for cis- or trans-1,2-DCE or their mixtures were identified.
4.2.1. Oral Exposure
4.2.1.1. Short-term Studies
4.2.1.1.1. cis-l,2-DCE. McCauley et al. (1990, unpublished) conducted a 14-day gavage study
of cis-l,2-DCE in male and female Sprague-Dawley rats. The study was subsequently published
(McCauley et al., 1995). Upon review and comparison of the unpublished McCauley report
(McCauley et al., 1990) and the published study (McCauley et al., 1995), errors in the
documentation of doses and other minor inconsistencies were noted. These errors were not
considered to compromise the reliability of the findings. Cis-1,2-DCE was administered by
gavage in corn oil vehicle to approximately 10-week-old Sprague-Dawley rats (10/sex/dose) at
doses of 0, 1, 3, 10, and 20 mmol/kg-day (equivalent to 0, 97, 291, 970, and 1,940 mg/kg-day,
respectively).1 At the end of the exposure period, animals were sacrificed and the brain, gonads,
^oses in the 1995 study were incorrectly converted from mmol/kg-day to mg/kg-day. The doses presented here are
the correctly converted doses. In addition, the doses for the acute and subchronic study as presented in the 1995
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heart, kidneys, adrenals, liver, spleen, and thymus were weighed and examined for gross
pathology. Blood samples were collected for hematological and clinical chemistry examination.
Tissues from controls and the high-dose group animals were examined for histopathologic
changes.2
During the study, male and female rats in the 1,940 mg/kg-day groups released excessive
clear secretions around the nose and/or mouth and appeared agitated, followed by lethargy and
ataxia. These symptoms were most common immediately after dosing. Gavage-related deaths
were reported in the 1,940 mg/kg-day group (2/10 males and 3/10 females) and 970-mg/kg-day
group (1/10 males and 1/10 females). Increases in water consumption were also seen in both
male and female rats in the 1,940 mg/kg-day groups.
With the exception of a 10% decrease in male rat body weights in the 1,940 mg/kg-day
dose group, there were no significant changes in the final mean body weights. Significant dose-
related increases in relative liver weight were reported in both males (16-38%) and females (15—
39%) at all dose levels. Statistically significant increases were observed for relative kidney
weights in females in the 970 and 1,940 mg/kg-day groups (14 and 12%, respectively) and for
relative testes weights in males in the 1,940 mg/kg-day group (23%). Serum phosphorus levels
were significantly elevated in females in all experimental groups, and serum cholesterol was
increased in the 1,940 mg/kg-day group. Serum calcium was statistically significantly increased
in male groups dosed with 970 and 1,940 mg/kg-day. Decreases in blood urea nitrogen (BUN)
occurred in females at doses of 291(14%), 970 (28%), and 1,940 mg/kg-day (17%). (Increases
in BUN are generally indicative of an effect on kidney function.) Hematocrit values for females
were decreased by 8 to 11% at the 291, 970, and 1,940 mg/kg-day dose groups; similar effects
did not occur in males. The authors considered most of the clinical chemistry and hematology
effects to be marginal and not biologically meaningful or dose related. No compound-related
histopathological changes were found at sacrifice. The authors noted that cis-l,2-DCE affected
organ-to-body-weight ratios at relatively low exposure levels, but in light of negative
histopathology, these data were difficult to interpret.
4.2.1.1.2. trans-l,2-DCE. Barnes et al. (1985) conducted a 14-day gavage study in male and
female CD-I mice. Concentrations of trans-l,2-DCE were prepared so that each mouse received
approximately 1/100 and 1/10 of the lethal dose (LD50) (21 and 210 mg/kg) daily. No significant
differences in weight gain were observed among the treated groups. Weights of the brain, liver,
spleen, lungs, thymus, kidney, and testes were not altered by DCE exposure. All organ weights
were within the limits of historical controls, and there was no treatment effect when the weights
were expressed as absolute weight, percent of body weight, or organ-to-brain ratios. There were
published paper were reversed, i.e., the doses listed for the 14-day study are really for the 90-day study and vice-
versa.
2According to the unpublished report (McCauley et al., 1990), only half the controls, rather than all controls as
reported in McCauley et al. (1995), were examined for histopathologic changes.
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no changes seen in hematocrit or hemoglobin values. Fibrinogen levels were decreased by 12%
in the 210 mg/kg treatment group, and prothrombin activity increased slightly as manifested by a
7% decrease in prothrombin time. There were no significant differences in SGPT (ALT) activity
or BUN levels; however, a statistically significant decrease (29%) in the LDH levels of the 210
mg/kg trans-1,2-DCE group was observed. In a study by the same laboratory, discussed below
in Section 4.4.3.2, values for leukocyte counts, hematocrit, hemoglobin, fibrinogen, and
prothrombin time did not differ significantly from control values when identical experiments
were conducted in male CD-I mice (Munson et al., 1982).
4.2.1.1.3. Mixtures of cis- andtrans-l,2-DCE. In a dissertation, McMillan (1986) reported a
statistically significant increase in kidney weight in male Sprague-Dawley-derived rats (6/group)
administered a dose of 5 mmol (485 mg/kg-day) of a 50:50 mixture of cis-l,2-DCE and
trans-l,2-DCE (in a sesame seed oil vehicle, 1 mL/kg) in a 14-day gavage study. Slight
reductions (statistically significant) in plasma creatinine and BUN levels, and an increase in
plasma calcium levels were also recorded at termination.
4.2.1.2. Subchronic Studies
4.2.1.2.1. cis-l,2-DCE. In a 90-day study, 10 Sprague-Dawley rats/sex/group, approximately
70 days old at study initiation, were administered 97% pure cis-l,2-DCE in corn oil by gavage
(3 mL/kg) at doses of 0, 32, 97, 291, and 872 mg/kg-day (McCauley et al., 1995, 1990).
Comparison of the unpublished McCauley report (McCauley et al., 1990) and the published
study (McCauley et al., 1995) revealed errors in the documentation of administered doses and
other minor inconsistencies.3 These errors and inconsistencies were not considered to
compromise the reliability of the 90-day study findings. At the end of the 90-day exposure
period, animals were sacrificed and the brain, gonads, heart, kidneys, adrenals, liver, spleen, and
thymus were weighed and examined for gross pathology. Blood samples were collected for
hematological and clinical chemistry examinations. Tissues from controls and the high-dose
group animals were examined for histopathologic changes.4
Clinical observations during the study were reported by the authors as minimal and not
compound-related. Gavage deaths were present in both the treated and control groups
3The administered doses in McCauley et al. (1995) were reported as 0, 0.33, 1, 3, and 9 mmol/kg-day, which when
converted to mg/kg-day, are 0, 32, 97, 291, and 872 mg/kg-day. McCauley et al. (1995), however, reported the
converted doses incorrectly as 0, 10, 32, 98, and 206 mg/kg/day. The doses presented here are the correctly
calculated doses of doses of 0, 32, 97, 291, and 872 mg/kg-day, as reported in McCauley et al. (1990). In addition,
the summary of clinical chemisty findings in McCauley et al. (1995) did not adjust for early gavage-related deaths in
the number of animals studied. In addition, the doses for the acute and subchronic study as presented in the 1995
published paper were reversed, i.e., the doses listed for the 14-day study are really for the 90-day study and vice-
versa.
4According to the unpublished report (McCauley et al., 1990), only half the controls, rather than all controls as
reported in McCauley et al. (1995), were examined for histopathologic changes.
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(1/10 female rats at 32 mg/kg-day; 1/10 female rats at 97 mg/kg-day; 1/10 male controls;
3/10 male rats at 291 mg/kg-day; 4/10 male rats at 872 mg/kg-day).
Terminal body weights in male rats at the two highest dose groups were lower than
controls by 10-11%, but were not considered by the author as statistically significant; no
treatment-related effects on body weight were reported in female rats.
Absolute liver weights were statistically significantly increased by 10, 15 and 24% in
female rats at doses of 97, 291 and 872 mg/kg-day, respectively. The increases in absolute liver
weight of 6, 13, 5 and 15% in male rats of the 32, 97, 291 and 872 mg/kg-day dose groups,
respectively, were not statistically significant nor dose related (McCauley et al., 1990). Relative
liver weights (expressed as a ratio of liver weight to body weight) were statistically significantly
increased in a dose-related manner in males and females (Table 4-1). The increases were 15, 17,
and 32% for males and 14, 19, and 30% for females at 97, 291, and 872 mg/kg-day, respectively.
Histopathological evaluation revealed no specific hepatic injury. The authors concluded that
there was a consistent, dose-related increase in relative liver weight in both sexes and that this
effect, in light of the negative histopathology findings, may reflect hypertrophy and hyperplasia.
Table 4-1. Body weights and relative organ weights of rats exposed to cis-
1,2-DCE by gavage for 90 days
Control
Dose (mg/kg-day)
32
97
291
872
Malesa
Mean final body
weight (g)
578 62.0
558 75.1
569 55.7
520 46.6
512 55.1
Kidney
0.70 0.06
0.80 0.06b
(14%)c
0.83 0.06b
(19%)c
0.83 0.10b
(19%)c
0.89 0.06b
(27%)c
Liver
2.85 0.26
3.15 0.27
(10%)c
3.28 0.18b
(15%)c
3.34 0.44b
(17%)c
3.75 0.20b
(32%)c
Females'1
Mean final body
weight (g)
315 23.4
316 26.7
305 38.2
303 24.8
301 40.7
Kidney
0.69 0.06
0.71 0.05
(3%)c
0.82 0.23
(19%)c
0.85 0.21
(23%)c
0.85 0.06
(23%)c
Liver
2.82 0.19
2.91 0.18
(3%)c
3.21 0.22b
(14%)c
3.36 0.18b
(19%)c
3.67 0.27b
(30%)c
Thymus
0.99 0.18
1.40 0.27
1.00 0.29
1.11 0.33
1.16 0.31b
"Values are mean ± standard deviation (SD).
bStatistically significantly different from control group; p 0.05 by Tukey's multiple comparison test.
'Values are percent increases from control group.
Source: McCauley et al. (1995, 1990).
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Absolute kidney weights in female rats were increased by 3, 16, 17, and 17% compared
to the control at doses of 32, 97, 291 and 872 mg/kg-day, respectively, but were not statistically
significant. In male rats increases in absolute kidney weight of 9, 17, 7 and 14% for the 32, 97,
291 and 872 mg/kg-day dose groups, respectively, were not statistically significantly elevated
compared to the control nor dose related (McCauley et al., 1990). Statistically significant
increases in relative kidney weights (as a ratio of kidney weight to body weight) were recorded
in male rats in all dose groups (14, 19, 19, and 27% at 32, 97, 291, and 872 mg/kg-day,
respectively) (Table 4-1). Female rats exhibited increased (although not statistically significant)
relative kidney weights in the three highest doses (19, 23, and 23% at 97, 291, and 872 mg/kg-
day, respectively). Relatively large variances in the female dose groups may explain why
relative kidney weight increases in females were not statistically significant. Histopathological
findings for kidney effects were negative, leading the authors to hypothesize that the increases in
relative kidney weight may be due at least in part to decreased body weight gain.
Sporadic changes (although noted as statistically significant) in some clinical chemistry
parameters were observed. BUN levels were significantly decreased (40%) at the highest dose in
males but not in females. Serum calcium levels were significantly elevated by 8 and 10% in
males at the 32 and 97 mg/kg-day doses, respectively, and serum phosphorus was significantly
decreased by 14% in males exposed to 32 mg/kg-day. In females, serum phosphorus was
significantly increased by 34 and 25% in the groups dosed with 97 and 291 mg/kg-day,
respectively. No significant changes were reported in AST activity. Hemoglobin and hematocrit
level, and red blood cell (RBC) count were significantly decreased in female rats dosed at
291 mg/kg-day, while only hematocrit was significantly decreased in females dosed with
872 mg/kg-day. In males, similar decreases (ranging from 6 to 10% compared with the control)
occurred in hemoglobin in the 291 and 872 mg/kg-day groups and in hematocrit in the 97, 291,
and 872 mg/kg-day groups. Overall the changes in clinical chemistry and hematology
parameters were considered by the authors to be marginal and of questionable biological
significance. No noteworthy compound-related histopathological changes were observed in any
dose group.
4.2.1.2.2. trans-l,2-DCE. There are three subchronic studies that evaluated oral exposure to
trans-l,2-DCE (NTP, 2002; Hayes et al., 1987; Barnes et al., 1985, with a companion
immunology study by Shopp et al. [1985], which is discussed in Section 4.4.3.2). In a 90-day
study by Hayes et al (1987), groups of 20 male and 20 female rats, approximately 29-37 days of
age, were administered 98% pure trans-l,2-DCE in drinking water containing 1% emulphorto
promote solubility. The experimental groups consisted of an untreated control group, a 1%
emulphor control group, and three test groups receiving drinking water containing trans-1,2-DCE
sufficient to provide approximate daily doses of 500, 1,500, and 3,000 mg/kg. Based on fluid
consumption measured twice weekly, actual mean doses were 0, 402, 1,314, and 3,114 mg/kg-
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day for males and 0, 353, 1,257, and 2,809 mg/kg-day for females. Effects on body weight,
organ weights, hematology, urine, and blood chemistries were examined. Gross pathological
examinations were performed following removal and weighing of selected organs.
A dose-related increase in fluid consumption was observed among the groups receiving
drinking water with emulphor, but the differences did not attain statistical significance. No
trans-1,2-DCE-related changes in behavior or interim deaths were observed. The mean body
weight of male rats increased from 100 to approximately 500 g for all dose groups during the
study. Female rats increased in body weight from 100 g at the beginning of the study to 250 g at
the end of the study. Although the male rats gained considerably more weight than the females
during the course of the study, at termination no statistically significant compound-related
differences in body weights or body weight gains were found in either the males or females
among the five groups. The authors reported that there were no consistent, remarkable
compound-related, dose-dependent effects on any of the hematological, serological, or urinary
parameters evaluated. No significant changes in organ weights or relative organ weights
(expressed either as a ratio of organ weight to body weight or organ weight to brain weight) were
seen in males, and only absolute kidney weights (Table 4-2) and kidney weights relative to brain
weights were statistically significantly elevated in the mid- and high-dose groups of female rats.
These increases were 8 and 9% for absolute kidney weight and 11 and 11% for kidney weights
relative to brain weights for the mid- and high-dose female rats, respectively. Dose-related
increases (although not statistically significant) in liver weights occurred in both sexes. A
limited number of organs (livers, kidneys, testes, and ovaries from 10 rats/sex/dose) were
examined microscopically at termination, and no compound-related histopathological changes
were reported.
Table 4-2. Absolute kidney weights in rats treated with trans-l,2-DCE via
drinking water for 90 days
MaleJ3
Dose (mg/kg-day)
Vehicle
402
1,314
3,114
Kidney weight (g)a
4.26± 0.07
4.36± 0.10
4.44± 0.10
4.41± 0.09
Female£
Dose (mg/kg-day)
Vehicle
353
1,257
2,809
Kidney weight (g)a
2.20 ±0.04
2.26 ±0.04
2.37 ± 0.04c
2.40 ± 0.03c
"Mean ± standard error (SE).
b17-20 animals per group.
'Statistically significant, p < 0.05.
Source: Hayes et al. (1987).
NTP (2002) conducted a 14-week study with trans-l,2-DCE in rats and mice. F344/N
rats, 10/sex/dose, were fed diets containing microcapsules with a chemical load of 45%
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trans-l,2-DCE at dietary concentrations of 0, 3,125, 6,250, 12,500, 25,000, and 50,000 ppm,
resulting in average daily trans-1,2-DCE doses of 0, 190, 380, 770, 1,540, and 3,210 mg/kg-day
for males and 0, 190, 395, 780, 1,580, and 3,245 mg/kg-day for females, respectively. B6C3Fi
mice (10/sex/group) received 0, 480, 920, 1,900, 3,850, and 8,065 mg/kg-day for males and 0,
450, 915, 1,830, 3,760, and 7,925 mg/kg-day for females (NTP, 2002). Additional groups
(10 males and 10 females) of rats and mice served as untreated and vehicle controls (animals that
received feed with unloaded microcapsules). Animals were evaluated for survival, body weight
(weekly), and feed consumption (weekly). Necropsies were performed on all animals. Organ
weights were measured for the heart, right kidney, liver, lung, right testis, and thymus. Clinical
findings, including hematology (rats only), clinical chemistry, and histopathology, were
performed. Hematology parameters included hematocrit; hemoglobin concentration;
erythrocyte, reticulocyte, and platelet counts; erythrocyte and platelet morphology; mean cell
volume; mean cell hemoglobin; mean cell hemoglobin concentration; and leukocyte count and
differentials. Complete histopathology was performed on all rats and mice in the untreated
control, vehicle control, and 50,000 ppm groups (3,210 and 3,245 mg/kg-day in male and female
rats; 8,065 and 7,925 mg/kg-day in male and female mice, respectively). In addition to gross
lesions and tissue masses, the following tissues were examined: adrenal gland, bone with
marrow, brain, clitoral gland, esophagus, gallbladder (mice only), heart, large intestine (cecum,
colon, rectum), small intestine (duodenum, jejunum, ileum), kidney, liver, lung, lymph nodes
(mandibular and mesenteric), mammary gland, nose, ovary, pancreas, parathyroid gland,
pituitary gland, preputial gland, prostate gland, salivary gland, skin, spleen, stomach
(forestomach and glandular), testis with epididymis and seminal vesicle, thymus, thyroid gland,
trachea, urinary bladder, uterus, and Zymbal's gland.
In the rat study, there were no exposure-related deaths. The final mean body weight and
body weight gain of male rats exposed to trans-l,2-DCE in the 3,210 mg/kg-day group were
reduced by about 6% (statistically significant) below controls (Table 4-3). Feed consumption in
the exposed groups was similar to that in the vehicle controls. On day 21 and at week 14, there
were mild decreases (generally less than 5% below the values in controls) in hematocrit values,
hemoglobin concentrations, and erythrocyte counts in 1,540 and 3,210 mg/kg-day males and
1,580 and 3,245 mg/kg-day female rats. At week 14, these effects were also seen in male rats
exposed to 3 80 and 770 mg/kg-day trans-1,2-DCE. Females exposed to >780 mg/kg-day had
statistically significantly decreased serum alkaline phosphatase (ALP) activities compared with
the vehicle controls on day 21. These decreases were noted by the authors to be minimal in
severity, no greater than about 13%, and transient, with activities in the affected groups returning
to vehicle control levels by week 14. On day 21, it was also noted that there was a minimal
suppression of serum 5'-nucleotidase activities in the 3,210 mg/kg-day male and the 3,245
mg/kg-day female rats. According to the authors, these sporadic differences in clinical chemistry
parameters at various time points generally did not demonstrate exposure concentration
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relationships or were inconsistent between males and females. These differences were not
considered by the authors to be toxicologically relevant.
Table 4-3. Final body weights in rats exposed to trans-l,2-DCE in the feed
for 14 weeks
Males
Dose (mg/kg-day)
Vehicle
190
380
770
1,540
3,210
Body weight (g)a
360 ±6
365 ±5
361 ±3
357 ±5
350 ±6
339 ± 4b
Females
Dose (mg/kg-day)
Vehicle
190
395
780
1,580
3,245
Body weight (g)a
190 ±4
198 ± 3
203 ± 2C
198 ± 3
196 ±3
191 ±2
"Values are mean ± SE.
bStatistically significant difference from controls, p < 0.01.
'Statistically significant difference from controls, p < 0.05.
Source: NTP(2002).
NTP (2002) reported mild decreases in hematocrit values, hemoglobin concentrations,
and erythrocyte counts at week 14 in male and female rats in all but the lowest dose groups. Of
these parameters, only the changes in RBC counts were dose related and statistically
significantly different from the vehicle control (p < 0.01) (see Table 4-4). The maximum
decrease in RBC was 7% in males and 5% in females at the highest dose (3,210 mg/kg-day in
male rats and 3,245 mg/kg-day in female rats).
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Table 4-4. Red blood cell counts in rats exposed to trans-l,2-DCE in the
feed for 14 weeks
Malesp
Dose (mg/kg-day)
Vehicle
190
380
770
1,540
3,210
RBC (106/|aL)a
8.14± 0.08
8.17 ± 0.05
7.93 ± 0.10c
7.84 ± 0.09c
7.79 ± 0.08d
7.56 ± 0.15d
Female/
Dose (mg/kg-day)
Vehicle
190
395
780
1,580
3,245
RBC (106/|aL)a
7.59 ±0.06
7.58 ± 0.10
7.50 ±0.08
7.49 ±0.04
7.34 ± 0.05d
7.20 ± 0.08d
"Values are mean ± SE.
bTen animals in each group except for the male 380 mg/kg-day group with only nine animals.
'Statistically significant difference, p < 0.05.
Statistically significant difference, p < 0.01.
Source: NTP (2002).
In female rats exposed to >395 mg/kg-day the absolute and relative liver weights
(expressed as a ratio of liver weight to body weight) were approximately 10-17% and 6-10%
higher (statistically significant), respectively, than those of the vehicle controls (see Table 4-5).
The greatest increases were observed at the 395 mg/kg-day dose. Absolute kidney weights of
male rats exposed to 1,540 or 3,210 mg/kg-day trans-l,2-DCE were decreased by about 7%. No
gross or histological lesions were observed in rats that were attributed to exposure to trans-
1,2-DCE.
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Table 4-5. Relative liver weights in mice and rats exposed to trans-l,2-DCE
in the feed for 14 weeks
Mice -
malesh
Dose (mg/kg-day)
0
480
920
1,900
3,850
8,065
Relative liver
weights2
4.347 ±0.056
4.552± 0.113
4.597 ±0.115
4.745 ±0.084c
4.736 ±0.079c
4.979± O.llT
Mice - femalesb
Dose (mg/kg-day)
0
450
915
1,830
3,760
7,925
Relative liver
weights2
4.621 ±0.07
4.738 ±0.068
4.970 ±0.127
4.813 ± 0.05
5.115 ± 0.139c
5.117 ± 0.08c
Rats-
males*
Dose (mg/kg-day)
0
190
380
770
1,540
3,210
Relative liver
weights2
3.465 ±0.058
3.538 ±0.032
3.658 ±0.099
3.524 ±0.050
3.492 ±0.048
3.634 ±0.056
Rats - femalesb
Dose (mg/kg-day)
0
190
395
780
1,580
3,245
Relative liver
weights2
2.937 ±0.038
3.040 ±0.052
3.220 ±0.066c
3.100±0.051c
3.132 ±0.052c
3.216 ± 0.051c
"Value are mean ± SE
bTen animals per group
'Statistically significant, p < 0.01.
Source: NTP (2002).
In the mouse study, no exposure-related deaths occurred. Mean body weights of
8,065 mg/kg-day males and 7,925 mg/kg-day females were significantly less (both by about 7%)
than those of the vehicle controls. Mean body weight gains of female mice in the 1,830 and
3,760 mg/kg-day groups were also significantly less (6 and 4%, respectively) than in vehicle
controls. Feed consumption in the exposed groups was similar to that in the vehicle controls.
No exposure-related alterations in clinical chemistry parameters were observed.
As shown in Table 4-5, the relative liver weights (expressed as a ratio of liver weight to
body weight) of male mice exposed to >1,900 mg/kg-day and female mice exposed to 3,760 or
7,925 mg/kg-day were significantly greater than those of the vehicle controls. Relative liver
weight increases in the male mice were 10% or less except for the high dose (8,065 mg/kg-day),
which showed a 14% increase compared with the vehicle control. The relative liver weights in
the two highest female mice dose groups (3,760 mg/kg-day and 7,925 mg/kg-day) were
increased by about 12% over vehicle controls. Other than a statistically significant increase of
16% in the 915 mg/kg-day female mice, there was no significant dose-related change in absolute
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liver weight. No gross or microscopic lesions were observed in mice that could be attributed to
trans-1,2-DCE exposure.
It was concluded by NTP (2002) that little toxicity was associated with ingestion of
microencapsulated trans-1,2-DCE, and that the histopathology and clinical chemistry data
combined with organ and body weight data revealed that the maximum tolerated dose (MTD)
had not been reached in this study.
In a 90-day drinking water study conducted by Barnes et al. (1985), groups of male and
female CD-I mice (24 mice/sex in the control group and 16 mice/sex in the treatment groups)
were exposed to trans-1,2-DCE, purity 98%, dissolved in deionized water at 0, 0.1, 1.0, or
2.0 mg/mL. Target daily doses were 1/100, 1/10, and 1/5 the acute oral LD50; actual time-
weighted average daily doses calculated on the basis of water consumption were 0, 17, 175, and
387 mg/kg-day for males and 0, 23, 224, and 452 mg/kg-day for females. Body and organ
weights, hematology, serum chemistries, and hepatic microsomal activities were measured.
Fluid consumption by both male and female mice progressively decreased throughout the
duration of the experiment, with comparable changes occurring in the control group. Few
trans- 1,2-DCE-induced changes in terminal body weight gain or gross pathology were observed
at the time of necropsy in either sex. As shown in Table 4-6, male mice receiving 175 mg/kg-
day trans-1,2-DCE demonstrated a statistically significant increase in mean absolute liver
weights and relative liver weights (expressed as a ratio of liver weight to body weight); however,
absolute liver weights were less than those of the controls in the low- and high-dose groups.
Females receiving 452 mg/kg-day demonstrated an 11% decrease (statistically significant) in
absolute lung weights. Additionally, absolute thymus weight was reduced by 24% (statistically
significant) at the high dose in females and relative thymus weights were statistically
significantly reduced in the mid- and high-dose females.
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Table 4-6. Results of 90-day study in CD-I mice exposed to trans-l,2-DCE in
the drinking water
Parameterb
Malesa
Dose (mg/kg-day)
Vehicle
17
175
387
Liver weight (mg)
[% body weight]
2,029 ± 43
[5.10]
2,007 ± 62
[5.01]
2,288 ±60c (8%)
[5.53c]
2,022 ± 85
[5.17]
Lung weight (mg)
[% body weight]
232 ±4
[0.58]
228 ±5
[0.57]
236 ±6
[0.57]
223 ±5
[0.58]
Thymus weight (mg)
[% body weight]
48 ±3
[0.12]
47 ±3
[0.12]
54 ±2
[0.13]
48 ±2
[0.12]
Kidney weight (mg)
[% body weight]
651 ±21
[1.64]
637 ± 22
[1.59]
658 ±19
[1.59]
634 ± 27
[1.63]
Prothrombin time (sec)
10.0 ± 0.2
8.5 ± 0.2C (15%)
8.8 ± 0.3C (12%)
9.8 ±0.2
Leukocytes (103/mm3)
5.30 ±0.32
4.95 ±0.42
4.83 ±0.24
5.16 ± 0.40
Glucose (mg %)
153 ± 7
195 ± 8C (27%)
184 ± 5C (20%)
190 ±7C (24%)
LDH (IU/L)
677 ± 33
605 ± 47
449 ± 22c (34%)
587 ±56
SGPT (IU/L) (or ALT)
44.3 ±3.3
55.1 ±7.1
45.0 ± 6.8
41.2 ± 4.7
SGOT (IU/L) (or AST)
74.0 ±6.5
110.0 ± 7.8C (48%)
65.3 ±5.0
69.9 ± 5.8
SAP (IU/L) (or ALP)
34.3 ±1.8
37.6 ±5.1
55.5 ± 5.4C (62%)
45.6 ± 2.4C (33%)
Parameter
Femalesd
Dose (mg/kg-day)
Vehicle
23
224
452
Liver weight (mg)
[% body weight]
1,712 ± 57
[5.27]
1,839 ± 51
[5.44]
1,864 ±38
[5.49]
1,741 ±57
[5.46]
Lung weight (mg)
[% body weight]
254 ± 11
[0.79]
255 ±7
[0.76]
244 ±7
[0.72]
222 ±8C (11%)
[0.70c]
Thymus weight (mg)
[% body weight]
71 ±3
[0.22]
67 ±4
[0.20]
61 ±4
[0.18C]
54 ± 4C (24%)
[0.1 T]
Kidney weight (mg)
[% body weight]
461 ± 12
[1.43]
456 ±9
[1.35]
465 ±13
[1.37]
428 ±8
[1.35]
Prothrombin time (sec)
9.7 ±0.2
9.8 ±0.3
9.1 ±0.2
9.0 ±0.6
Leukocytes (103/mm3)
7.27 ±0.32
6.98 ±0.50
8.95 ± 0.61c (23%)
7.79 ±0.60
Glucose (mg %)
122 ±3
156 ±6C (28%)
147 ±5C (20%)
156 ±6C (28%)
LDH (IU/L)
511 ±22
377 ± 20c (26%)
452 ± 23
559 ±42
SGPT (IU/L) (or ALT)
49.9 ±6.4
38.3 ± 3.0
33.5 ± 3.6C (33%)
30.4 ± 1.6C (39%)
SGOT (IU/L) (or AST)
91.7 ± 6.6
77.8 ±6.0
66.9 ± 4.5C (27%)
58.9 ± 8.8C (36%)
SAP (IU/L) (or ALP)
44.0 ±2.3
47.6 ±4.2
51.0 ± 3.0
45.4 ±2.9
aTwenty-three animals/sex in the control group and 15-16 animals/sex in the treatment groups.
bValues presented are mean± SE.
cDiffers statistically significantly from controls, p < 0.05; Duncan's multiple range test was used to determine
statistical significance.
dTwenty-four animals/sex in the control group and 16 animals/sex in the treatment groups.
Source: Barnes et al. (1985).
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Few changes in hematological parameters were seen; prothrombin time was significantly
decreased by 15 and 12% in male mice exposed to 17 and 175 mg/kg-day trans-1,2-DCE,
respectively, and only decreased by 2% at the high dose (387 mg/kg-day) in the male mice. In
female mice exposed to 224 mg/kg-day, an increase (23%) in blood leukocytes and decrease
(42%) in polymorphonuclear leukocytes occurred. Slight changes in several clinical chemistry
parameters were observed. Although some values were statistically significantly different from
those of the controls, there were no consistent trends or any large deviations from historical
control values. The most noteworthy change was a statistically significant increase in the serum
glucose levels at all dose levels in both males and females. However, the toxicological
significance of these increases is unknown because the values were well within the wide range of
measured values for control mice and a dose response was not demonstrated, even though the
range of doses was 20-fold.
In male mice significant changes in serum LDH and, AST (SGOT) enzymes, and ALP
activities, which provide some indication of hepatocellular injury, were reported. Significant
increases of 62 and 33% were observed in serum ALP levels at the 175 and 387 mg/kg-day
levels, respectively, in male mice. Such increases, however, showed no dose-response
relationship and were not found in the female mice. In female mice, ALT and AST were
depressed at all levels of exposure to trans-1,2-DCE; the decreases were statistically significant
at the two higher dose levels. In both sexes, sporadic elevations in serum potassium and
depressions in serum sodium and calcium were not considered biologically significant. In males,
serum glutathione (GSH) levels were depressed 21% in the highest dose group.
In this same study (Barnes et al., 1985), possible effects of trans-l,2-DCE exposure on
hepatic microsomal drug metabolism potential were assessed by measuring hexobarbital sleeping
time and by evaluating microsomal protein/g liver, CYP450 and cytochrome b5 concentrations,
and microsomal activities for aminopyrine N-demethylase and aniline hydroxylase.
Hexobarbital sleeping times were not affected in the various dose groups exposed to
trans-l,2-DCE in either sex. In male mice, exposure to 175 mg/kg-day trans-1,2-DCE
significantly decreased the microsomal metabolizing activities of both aminopyrine
N-demethylase (17%) and aniline hydroxylase (27%). In contrast, no significant changes in
these enzyme activities were observed in the 387 mg/kg-day exposure group in males. In female
mice, aniline N-hydroxylase activity was statistically significantly depressed in all exposure
groups, although the decreases (21, 33, and 28%, respectively) were not dose-dependent.
4.2.1.2.3. Mixtures of cis- andtrans-l,2-DCE. McMillan (1986) conducted a 30-day
subchronic study with a 50% mixture of the cis- and trans-isomers orally administered in sesame
seed oil (1 mL/kg) to male Sprague-Dawley-derived rats (six/group) at a daily dose of 5 mmol
(485 mg/kg-day); control rats received the vehicle alone. At termination, the mean relative liver
weight in the treated group (expressed as a ratio of liver weight to body weight) was 19% greater
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(statistically significant) than that of control rats, while the mean relative weight of the lungs was
significantly reduced by 14%. Mean values for plasma AST activity and creatinine levels in the
treated group were significantly reduced by 25 and 17%, respectively, at sacrifice. The mean
plasma calcium level in the treatment group was elevated by about 14%, as was plasma chloride
by about 3%, while plasma K and CO2 were slightly depressed. Reductions in erythrocyte count,
hemoglobin, and hematocrit were also recorded as decreases of 6, 5, and 5%, respectively.
4.2.1.3. Chronic Studies
No chronic toxicity studies for the cis- and trans- isomers of 1,2-DCE administered by
the oral route were found.
4.2.2. Inhalation Exposure
4.2.2.1. Short-term Studies
No short-term inhalation studies of cis-1,2-DCE, trans-1,2-DCE, or mixtures of cis- and
trans-1,2-DCE were identified.
4.2.2.2. Subchronic Studies
4.2.2.2.1. cis-l,2-DCE. No subchronic inhalation studies of cis-l,2-DCE were identified.
4.2.2.2.2. trans-l,2-DCE. Freundt et al. (1977) exposed six mature female SPR Wistar
rats/group for 8 hours/day, 5 days/week to air containing 200 ppm (792 mg/m3) trans- 1,2-DCE
for 1, 2, 8, and 16 weeks. Concentrations were monitored by GC. Selected tissues (lung, heart,
liver, kidney, spleen, brain, muscle, and sciatic nerve) were examined for gross and
histopathological changes. Blood samples were analyzed for clinical chemistry and hematology
parameters.
Changes in alveolar septal distension of the lungs and slight to severe fatty degeneration
of the liver lobules and Kupffer cells were observed (see Table 4-7). Pathological changes in the
lung were noted and consisted of pulmonary capillary hyperemia and alveolar septal distention in
all six rats in all four exposure duration groups. These changes in the lung were considered by
the authors to be slight. These changes were also seen in one of the control animals exposed for
1 week and in two of the control animals exposed for 2 weeks, but not in any of the control
animals exposed at either 8 or 16 weeks. This is the only reported study of lung pathology in
animals exposed to trans-1,2-DCE. The evaluation of respiratory effects is limited by the
following: pulmonary capillary hyperemia and alveolar septal distention was also present in the
control animals, there were a small number of animals examined, and the upper respiratory tract
was not examined for pathology. A statistical evaluation of the histological data on the
respiratory system was not presented in this study.
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Table 4-7. Histopathological changes in subchronic inhalation study of
trans-l,2-DCE
Exposure
Rat
Liver effect:
fat accumulation—
liver lobule3
Liver effect:
fat accumulation—
Kupffer cells
Rat
Lung effect: capillary
hyperemia, alveolar
septum distention
Controls
1-6
0
0
1-5
6
0
+
200 ppm/8 hr for 1
week (5 days)
1-4
5,6
0
+
0
+
1-6
+
Controls
1-6
0
0
1-4
5,6
0
+
200 ppm/8 hr for 2
weeks
(5 days/week)
1,2
3-6
0
+
0
+
1-6
+
Controls
1-5
6
0
0
0
++
1-6
0
200 ppm/8 hr for 8
weeks
(5 days/week)
1-3
4-6
0
+
0
++
1-6
+
Controls
1-4
5,6
0
+
0
+
1-6
0
200 ppm/8 hr for 16
weeks
(5 days/week)
1
2,3
4-6
0
+
++
0
+
+
1-6
+
0 = no change; + = slight change; ++ = severe change.
Source: Freundt et al. (1977).
Histopathological changes were also observed in the liver and included fat accumulation
of liver lobules and Kupffer cells. After exposure to 792 mg/m3 for 1 week, slight fat
accumulation in liver lobules and Kupffer cells occurred in two of six rats but not in any of the
controls. When rats were exposed for 2 weeks under the same conditions, slight fat
accumulation in liver lobules and Kupffer cells occurred in four of the six rats but not in any of
the controls. After exposure to 792 mg/m3 for 8 weeks, three of the six rats showed evidence of
slight changes in the liver lobules and severe changes in the Kupffer cells. At this exposure for
8 weeks, severe fat accumulation was also noted in the Kupffer cells in one of the six controls.
When rats were exposed for 16 weeks under the same conditions, slight changes in the Kupffer
cells and severe changes in the liver lobule were noted in three of the six exposed rats. Slight
changes in both the Kupffer cells and in the liver lobules occurred in two other treated animals in
this 16-week exposure group for a total of five of the six rats showing some liver effect in this
exposure group. However, slight changes in both the Kupffer cells and in the liver lobule also
occurred in two of the control animals in the 16-week exposure group. For each of the exposure
durations (1, 2, 8, and 16 weeks) there was no statistically significant difference between the
controls and the exposed groups with respect to the incidence of liver effects (fat accumulation
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or Kupffer cells). However, in general, the incidence and severity of fat accumulation increased
with increasing exposure duration.
As described in an unpublished report by DuPont (1998), the subchronic toxicity of trans-
1,2-DCE (>99.4% pure) was evaluated in Crl:CD BR male and female rats (15/sex/group)
exposed to analytically determined mean concentrations of 0, 200, 1,000, or 4,000 ppm (0, 792,
3,960, or 15,800 mg/m3) for 6 hours/day, 5 days/week for 90 days. An abstract of this study was
published by Kelly et al. (1999). Ten of the 15 rats/sex/group were sacrificed at 90 days while
the remaining 5 rats/sex/group were allowed to recover for 1 month. Study parameters included
clinical observations, clinical laboratory evaluations (hematology, clinical chemistry), urinalysis,
and pathological evaluations (organ weights, gross observations, and microscopic evaluations).
Clinical pathology was evaluated in rats (10/sex/group) at 45 and 90 days. Pathology was
evaluated at 90 days (10/sex/group) and after a 1-month post exposure period (5/sex/group). An
additional four groups (15/rats/sex/group) designated for cell proliferation evaluations were
exposed to the same trans-1,2-DCE concentrations; 5 rats/sex/group were sacrificed for cell
proliferation evaluation on test days 9, 89-99, and 134-135 (tissues from animals in the recovery
group, i.e., 134-135 days, were not evaluated).
There were no treatment-related effects on body weight or on food consumption during
exposures or during an observation period of 1 month post exposure. The incidence of wet or
stained perineal areas was increased in the 1000- and 4000-ppm female rats. The incidences of
stained perineum and wet perineum were 0/15, 1/15, 4/15, and 5/15 and 0/15, 1/15, 6/15, and
8/15 in 0, 200-, 1000-, and 4000-ppm female rats, respectively. This condition was characterized
as transient as there were no wet or stained perineal areas observed in these rats in the morning
during the rat weighings. The authors of the study note that this observation is common in rats
after inhalation exposure and may have been related to the stress of the exposure
Clinical chemistry changes observed in the DuPont (1998) study were generally not dose
dependent, were transient (i.e., observed at 45 but not 90 days), or were of small magnitude, and
were not considered by the investigators to be toxicologically important. No statistically
significant or toxicologically important analytical urine changes occurred during this study.
Hematological analysis of blood samples collected on days 45 and 90 revealed a few statistically
significant hematological findings in male and female rats. Mean hemoglobin concentration and
hematocrit were statistically significantly decreased in the 1000- and 4000-ppm males at the 45-
day sampling time and mean monocyte count was statistically significantly decreased in the
4000-ppm females at the 45-day sampling time; these changes were not considered to be
toxicologically important because the changes were small in the context of historical controls
and a similar change did not occur at the 90-day sampling time. Generally dose-related
decreases in white blood cell (WBC) and lymphocyte counts were observed at both 45 and 90
days in male and female rats. These data are summarized in Table 4-8.
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Table 4-8. Selected hematology findings in rats exposed to trans-l,2-DCE by
inhalation for 90-days
Parameter
Concentration
(ppm)
Sampling time
45-day
90-day
Males
WBC (x 103/j_iL)a
0
17.2 ± 2.3
15.7 ± 2.0
200 (792 mg/m3)
15.0 ± 2.3
13.6 ± 2.5
1000 (3,960 mg/m3)
16.5 ±4.1
13.6 ± 3.4
4000 (15,800 mg/m3)
13.9 ± 1.6b
12.6 ±1.8
Lymph (/jaL)a
0
13953 ±2321
12901±1961
200 (792 mg/m3)
12187 ±2293
10670 ±2189
1000 (3,960 mg/m3)
13766 ± 3455
10706 ±2766
4000 (15,800 mg/m3)
10451 ±900b
9597 ± 1230b
Females
WBC (x 103/j_iL)a
0
15.5 ± 4.9
11.7 ± 4.5
200 (792 mg/m3)
13.5 ± 2.7
10.1 ±0.9
1000 (3,960 mg/m3)
13.2 ± 3.3
9.0 ±2.3
4000 (15,800 mg/m3)
12.1 ±2.2
9.6 ±2.1
Lymph (/jaL)a
0
13295 ±4389
10239 ±4147
200 (792 mg/m3)
11508 ±2792
8337 ±892
1000 (3,960 mg/m3)
11244 ±2880
7705 ±2147
4000 (15,800 mg/m3)
10516 ±1989
7948 ±1943
"Group means ± SD.
bSignificantly different from the control (p < 0.05) by Dunnett's criteria.
Source: DuPont(1998)
At the high dose, WBC decreased about 20% in male rats and 18% in female rats
compared to the control; lymphocyte levels decreased about 25% in males and about 22% in
females.
In general, organ weights in exposed rats showed no statistically or biologically
significant changes (less than 10%) relative to the control (DuPont, 1998). Relative adrenal
weight (as percent of body weight) was statistically significantly reduced in male rats at 1 month
recovery at the mid concentration (1000 ppm or 3,960 mg/m3) but not at the high concentration
(4000 ppm or 15,800 mg/m3) of trans- 1,2-DCE. In male rats at zero day recovery, increases in
relative liver weight (as percent of body weight) ranged from 4 to 8% and were not dose related.
Similarly, increases in relative liver weight (as organ to brain weight ratio) were 4, 4, and 6% at
concentrations of 200, 1000, 4000 ppm, (or 792, 3,960, 15,800 mg/m3), respectively. In female
rats at zero day recovery, increases in relative liver weight (as percent of body weight) were 1, 5,
and 6% at concentrations of 200, 1000, 4000 ppm (or 792, 3,960, 15,800 mg/m3), respectively;
increases in relative liver weight (as organ to brain weight ratio) in female rats at zero day
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recovery were 2, 8, and 8% at concentrations of 200, 1000, 4000 ppm [792, 3,960, 15,800
mg/m3], respectively. Increases in relative kidney weight for male and female rats (both as
percent of body weight and as organ to brain ratio) were less than 10% and were not generally
dose-related. For male rats, relative kidney weights ranged from 0-6% and 3-5% as percent of
body weight and as organ to brain weight ratio, respectively. Relative kidney weight for female
rats ranged from 2-5% as percent of body weight and from 5-8 % as organ to brain weight ratio.
No histopathological changes were related to exposure to trans-l,2-DCE.
4.2.2.2.3. Mixtures of cis- and trans-l,2-DCE. A subchronic inhalation study was conducted
by Dow (1962) in which rats, rabbits, guinea pigs (strains not stated), and beagle dogs were
exposed to 0, 500, or 1,000 ppm (0, 1,980, or 3,960 mg/m3) 1,2-DCE mixture (58% cis-, 42%
trans-isomer), 7 hours/day for 6 months. The 1,980 mg/m3 exposure groups consisted of 24 male
and 35 female rats, 7 male and 8 female guinea pigs, 3 male and 3 female rabbits, and 2 female
dogs, while the 3,960 mg/m3 exposure groups consisted of 12 male and 12 female rats and
2 male and 2 female rabbits. In addition to the animals receiving daily 7-hour exposures,
separate groups of 10 male rats were exposed to 1,980 mg/m3 1,2-DCE for 4, 2, or 1 hour(s)/day
for a total duration of 5 months. In all studies each animal was weighed twice per week until
growth was determined to be normal; afterwards, each animal was weighed once per week.
Hematological analyses and clinical chemistry determinations were performed on all rabbits, on
five male and five female rats exposed to 3,960 mg/m3, and on all dogs exposed to 1,980 mg/m3.
Rats and rabbits exposed to 3,960 mg/m3 of 1,2-DCE, 7 hours/day (136 exposures in
195 days) did not exhibit increased mortality or clinical signs of toxicity. Growth of animals was
normal, and final body weights and weights of lungs, heart, spleen, and testes were not
significantly different from controls. Hematology and biochemical values were within normal
limits. The average relative kidney weight (expressed as a ratio of kidney weight to body
weight) in male and female rats were increased by 16 and 9%, respectively (statistically
significant only in the males). The average relative liver weight in female rats (expressed as a
ratio of liver weight to body weight) was statistically significantly increased by 23%. Liver
weight in both male and female rabbits were also increased, but statistical significance was not
determined because of the small number of rabbits tested.
Rats exposed to 1,980 mg/m3 of 1,2-DCE, 7 hours/day, for six months did not
demonstrate excess mortality or adverse clinical effects. Hematology and clinical chemistry
values were within normal limits. Terminal body weights and relative lung, heart, spleen, and
testes weight were not significantly different from controls, but relative kidney weight of male
and female rats were increased by 9 and 18%, respectively (statistically significant only in the
male rats). Liver weights of female rats were significantly increased by 19%. No noteworthy
effects on mortality, behavior, or appearance were observed in the guinea pigs exposed to
1,980 mg/m3 on 81 of 117 days. Final average body weights and organ weights were not
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significantly different from controls. Rabbits exposed to 1,980 mg/m3 for 131 exposures in
181 days exhibited no effects, except that increases in liver weights of both male and female
rabbits occurred at termination (statistical evaluations were not performed because of the small
number of experimental animals). Female dogs exposed to 1,980 mg/m3 tolerated 129 exposures
in 183 days without biologically significant effects. Clinical chemistry and hematology data
were essentially identical to values obtained prior to initiation of the 1,980 mg/m3 exposure
regimen.
In rats exposed to 1,980 mg/m3 1,2-DCE for shorter periods of 4, 2, or 1 hour/day, 5 days
per week, for 5 months (136 exposures in a period of 195 days), no clinical or behavioral
abnormalities were seen, and final body weight and organ weight data were not significantly
different from control values. The BUN and ALP values were within normal limits.
4.2.2.3. Chronic Studies
No chronic inhalation exposure studies were identified for either cis- or trans-1,2-DCE.
4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND INHALATION
The available studies of reproductive and developmental outcomes are limited for both
the cis- and trans- isomers of 1,2-DCE. In an inhalation teratogenicity study of trans-l,2-DCE
(DuPont, 1988a), few developmental parameters were affected by treatment and these were
observed only in the high-exposure groups. Range-finding studies on the developmental toxicity
of a mixture of 1,2-DCE (composition of isomers unknown) via the oral route of exposure (NTP,
1991a, b, c) showed no signs of developmental or maternal toxicity at any of the initial doses
tested (up to 2,918 mg/kg-day), but at higher doses (up to 6,906 mg/kg-day) showed maternal
toxicity in the form of reduced maternal body weight and reduced maternal weight gain.
4.3.1. Oral Exposure
4.3.1.1. cis-l,2-DCE
No studies of reproductive or developmental toxicity of cis- 1,2-DCE in animals
following oral exposure were found.
4.3.1.2. trans-l,2-DCE
In a 14-week toxicity study described above in Section 4.2.1.2.2, NTP (2002) fed
F344/N rats and B6C3Fi mice diets containing microcapsules with a chemical load of 45%
trans-1,2-DCE. The rats (10/sex/group) received average daily trans-1,2-DCE dietary doses of 0,
190, 380,770, 1,540, mid 3,210 mg/kg-day for males and 0, 190, 395,780, 1,580, and
3,245 mg/kg-day for females. In the mouse study (10/sex/group), males received 0, 480, 920,
1,900, 3,850, and 8,065 mg/kg-day and females received 0, 450, 915, 1,830, 3,760, and
7,925 mg/kg-day (NTP, 2002). There were no organ weight changes or gross or microscopic
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lesions observed in the reproductive organs of rats or mice that would suggest that
trans-1,2-DCE targets the reproductive system.
4.3.1.3. Mixtures of cis- and trans-l,2-DCE
NTP conducted a series of developmental toxicity range-finding studies in mice and rats
with a mixture of 1,2-DCE isomers (composition unknown). The test compound was
administered in the feed in the form of microcapsules. Macroscopic or microscopic testing for
malformations was not conducted. In the mouse study (NTP, 1991a), 12 pregnant CD-I
mice/group were given feed containing 1,2-DCE mixture at concentrations of 0, 0.05, 0.25, 0.5,
1, and 1.5% on gestation days (GDs) 6-16. Doses were calculated based on feed intake and
body weight as 0, 97, 505, 979, 2,087, and 2,918 mg/kg-day, respectively. Body weight, feed
consumption, and any signs of toxicity were monitored. Dams were sacrificed on GD 17 and
their uteri examined. Gravid uterus weights, fetal body weights, and numbers of fetuses
(live/dead), implantation sites, and resorptions were recorded. None of the parameters showed
any deviation from control values. The authors concluded that, based on this range-finding
study, 1,2-DCE treatment did not cause maternal or developmental toxicity in mice at any of the
tested dose levels.
Ten pregnant Sprague-Dawley rats/group were subjected to the same experimental
protocol as above (NTP, 1991b) and received feed containing 0, 0.2, 1, 2, or 4% 1,2-DCE
mixture on GDs 6-16, resulting in doses of 0, 135, 672, 1,228, 1,966, and 2,704 mg/kg-day,
respectively. Dams were sacrificed on GD 20. The same parameters as in the mouse study were
examined. There were no signs of developmental or maternal toxicity at any of the levels tested.
Four animals of the highest dose group were found not to be pregnant; the study authors
considered this to be an isolated event not related to chemical treatment. A repeat of this study
was then undertaken with higher doses (NTP, 1991c). Pregnant Sprague-Dawley rats were
exposed on GDs 6-16 to feed containing 0, 4, 7.5, or 10% 1,2-DCE mixture, corresponding to
doses of 0, 3,134, 5,778, and 6,906 mg/kg-day. Dams were sacrificed on GD 20. The same
parameters as above were monitored. There was no mortality. Feed intake was dose-
dependently reduced. Maternal weight gain was dose-dependently reduced and statistically
significantly different from controls in the mid- and high-dose groups. Maternal body weights
were statistically significantly reduced at the 3,134 (3 and 15% on GDs 14 and 16, respectively),
5,778 (6, 7, 7, 9, mid 8% on GDs 9, 11, 14, 16, and 20, respectively), mid 6,904 mg/kg-day doses
(11, 12, 12, 14, and 10% on GDs 9, 11, 14, 16, and 20, respectively). Pregnancy outcome
numbers or fetal body weights were not affected by the treatment. Based on the dose ranges
used in this dose range-finding study, the authors concluded that DCE treatment caused maternal
toxicity at all dose levels based on reduced body weight. However, no changes were noted in the
limited number of fetal parameters evaluated in the study.
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4.3.2. Inhalation Exposure
4.3.2.1. cis-l,2-DCE
No studies of reproductive or developmental toxicity of cis- 1,2-DCE in animals
following inhalation exposure were found.
4.3.2.2. trans-l,2-DCE
In a study conducted by DuPont (1988a), and published in Hurtt (1993), trans-l,2-DCE
was administered to 24 pregnant female Crl:CD®BR rats/group by inhalation, 6 hours daily, on
GDs 7-16. Selection of the exposure levels was based on an MTD study conducted with
pregnant female rats prior to the actual experiment. On the basis of the pilot study, exposure
levels chosen for the actual study were 0, 2,000, 6,000, and 12,000 ppm (0, 7,920, 23,760, and
47,520 mg/m3, respectively). The low-exposure group concentration level was chosen to be
10 times the ACGIH TLV. Maternal body weight and feed consumption data were observed and
analyzed. Fetal weights were also noted. During the first two days of dosing, dams exposed to
23,760 mg/m3 showed slight weight gain suppression, and the 47,520 mg/m3 exposure group
showed statistically significant weight loss. Additionally, a statistically significant suppression
of body weight gain was noted in animals at the 23,760 mg/m3 concentration on GDs 11-13. For
the entire dosing period, a significantly reduced weight gain was observed only at the
47,520 mg/m3 concentration. Feed consumption was significantly reduced in the 23,760 and
47,520 mg/m3 groups throughout the exposure period. In the 7,920 mg/m3 group, there was a
significant decrease in feed consumption during GDs 13-15, although no significant effects on
body weight were reported. No significant changes in body weight or food consumption were
observed in any other group. As seen in maternal body weight change, no significant differences
were noted in feed consumption in the pre- or post exposure periods. Eye irritation was observed
in rats at all exposure levels. The only other compound-related clinical or postmortem findings
were increases in alopecia, salivation, and lethargy in rats during the periods of exposure,
especially in the high-exposure group.
Dams were sacrificed on GD 22 and their uteri examined. The mean number of
resorptions per litter was statistically significantly increased in dams in the 23,760 mg/m3 and
47,520 mg/m3 exposure groups. The values for resorptions in mid- and high-exposure groups
were within the range of historical controls in recent studies (past 2 years) conducted by the
laboratory and were not considered to be biologically significant but rather an artifact of the
unusually low resorption rate in the concurrent control group (0.3 mean resorptions per litter).
There were no differences in the pregnancy rate, fetuses per litter, number of stunted fetuses, or
number of corpora lutea observed per female. Developmental variations per litter were not
significantly increased in any of the exposure groups. No significant differences were detected
in the mean percent of fetuses per litter with malformations at any exposure level.
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In conclusion, treatment-related maternal and developmental toxicities were only
observed in high-concentration groups. Maternal toxicity was evidenced by statistically
significant decreases in body weight and feed consumption at 47,520 mg/m3 (the highest
exposure concentration tested), and by significant decreases in feed consumption at 23,760 and
7,920 mg/m3. The decrease in feed consumption at 23,760 mg/m3 was expressed as an observed
effect on body weights only on GDs 11-13. At 7,920 mg/m3, the effect on feed consumption,
seen only on GDs 13-15, was minimal and not accompanied by a significant body weight
change. Although the body weight change was lower for this group compared with controls,
their feed consumption was consistently lower than controls throughout the study. According to
the authors, this change was not accompanied by a statistically significant decrease in body
weight, and, therefore, its biological significance is questionable. Additionally, a statistically
significant trend was noted in the incidence of females with clinical findings on GDs 7-16, but
this was the result of ocular irritation in most animals. Significant developmental toxicity
(decreased mean fetal weight) was evident among fetuses exposed to 47,520 mg/m3
trans-1,2-DCE.
4.3.2.3. Mixtures of cis- and trans-l,2-DCE
No studies of reproductive or developmental toxicity of mixtures of 1,2-DCE in animals
following inhalation exposure were found.
4.4. OTHER DURATION- OR ENDPOINT-SPECIFIC STUDIES
4.4.1. Acute Studies
4.4.1.1. Oral Exposure
4.4.1.1.1. cis-l,2-DCE. In an acute hepatotoxicity study reported in a dissertation by McMillan
(1986), cis-l,2-DCE was administrated orally by gavage at doses of 26 mmol/kg (2,521 mg/kg)
and 51 mmol/kg (4,944 mg/kg) in sesame seed oil to six male Sprague-Dawley rats/dose. The
GSH levels were statistically significantly elevated by 19 and 28% at doses of 2,521 and
4,944 mg/kg, respectively. ALT activity was unchanged at either dose, but AST activity was
statistically significantly elevated by 56% at 4,944 mg/kg cis-1,2-DCE.
4.4.1.1.2. trans-l,2-DCE. Male and female Sprague-Dawley-derived CD rats, 22-30 days of
age, were administered a range of single doses of trans-1,2-DCE via corn oil gavage in a study
conducted by Hayes et al. (1987). The total volume of solution administered in this acute study
was 10 mL/kg. There were five dosage groups (exact doses not given), consisting of
10 rats/sex/group. Symptoms of dose-dependent central nervous system (CNS) depression,
ataxia, and depressed respiration were observed at all doses; all deaths occurred within 30 hours
of dosing. Although the exact dosages were not reported in this study, the authors determined
that the LD50 was 7,902 mg/kg (95% confidence interval [CI] 6,805-9,175 mg/kg) and
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9,939 mg/kg (CI 6,494—15,213 mg/kg) for male and female rats, respectively. Gross necropsy
findings of all rats that died were negative. No consistent compound-related gross pathological
findings were observed at necropsy.
An oral LD50 test was performed by Freundt et al. (1977) in which mature female SPF
Wistarrats received doses of 2-8 mL/kg trans-1,-2-DCE dissolved in olive oil (totaling
10 mL/kg each dose) via gavage. The LD50 was reported to be 1.0 mL/kg (95% CI: 0.9-
1.1 mL/kg) (1,280 mg/kg) for the rats treated via gavage. (The action of trans-l,2-DCE given
orally is more pronounced than after i.p. doses; described in Section 4.4.4.1.). One rat exhibited
gross pathology including pulmonary capillary hyperemia and alveolar septal distention and
fibrous swelling and hyperemia with incipient disorganization of the cardiac muscle. In two rats,
severe fatty infiltration of the liver lobules and Kupffer cells was found.
In an acute hepatotoxicity study reported in a dissertation by McMillan (1986),
trans-1,2-DCE was administered orally as a single dose of 51 mmol/kg (4,944 mg/kg). As with
the study conducted on the cis- isomer, six male rats were in each dose group. No differences
were seen between the controls and rats administered the single 4,922 mg/kg-day dose of
trans-1,2-DCE.
Barnes et al. (1985) also evaluated the acute oral toxicity of trans-l,2-DCE in 6-week-old
male and female CD-I mice. Trans-1,2-DCE was administered via gavage as a single dose after
an 18-hour fast. Nine different doses, ranging from 800-3,500 mg/kg of trans-l,2-DCE in
0.01 mL 1:9 emulphor:water vehicle per gram body weight, were used to generate dose-response
curves for DCE-induced mortality. The mice were observed continuously for 4 hours following
gavage and then twice daily for 14 days. All decedents and mice surviving the 14-day period
were subjected to gross necropsy. The LD50 was determined by log probit analysis to be
2,122 mg/kg (95% CI: 1,874-2,382) for male mice and 2,391 mg/kg (95% CI: 2,055-2,788) for
female mice. Upon gross necropsy, target organs were the lungs and liver. The lethality of the
test agent was attributed to depression of the CNS, as signs of decreased activity, ataxia,
suppression of the righting reflex, ruffled fur, and hunched back were seen. Hyperemia of the
mucosal surface of the stomach and small intestines was also observed at necropsy.
4.4.1.1.3. Mixtures of cis- and trans-l,2-DCE. Dow (1960) reported that a dose of 2,000 mg/kg
1,2-DCE (isomer composition not stated) as a 10% solution in corn oil administered by gavage
was not lethal to the exposed rats in this range-finding study. The Dow (1960) study noted that
"some kidney injury" was observed at necropsy, but that no other reactions were noteworthy.
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4.4.1.2. Inhalation Exposure
4.4.1.2.1. cis-l,2-DCE. A 4-hour inhalation median lethal concentration (LC50) study with
groups of five male and five female Crl:CD®BR rats was conducted with cis-l,2-DCE (DuPont,
1999). The exposure concentrations were 0, 12,100 (47,900 mg/m3), 13,500 (53,400 mg/m3),
15,700 (62,200 mg/m3), and 23,200 ppm (91,900 mg/m3). During a 14-day recovery period, rats
were weighed and observed for signs of clinical toxicity. All rats underwent gross pathological
examination immediately after death or at the end of the recovery period and the liver, kidney,
heart, and lung were evaluated histologically. The LC50 was calculated to be 54,200 mg/m3. It
was noted that the rats were prostrate, with eyes open, but unresponsive to alerting stimuli during
exposure. Other clinical signs included weakness and irregular respiration immediately after
exposure. After the 47,900 mg/m3 cis-l,2-DCE exposure, rats showed weakness, but there were
no effects on body weight at this concentration. Rats that did not die (two of five males and two
of five females) at the 53,400 mg/m3 exposure showed weakness and irregular respiration
immediately after the exposure and showed slight to severe weight loss for one day after
exposure, followed by a normal weight gain rate. All five of the male rats exposed (all of which
had died during the study) to 62,200 mg/m3 of the cis-isomer had minimal hepatic centrilobular
vacuolation. One male rat in each of the lower exposure levels (47,900 and 53,400 mg/m3) also
had minimal hepatic centrilobular vacuolation upon microscopic examination. There was one
male rat in the highest exposure group with this lesion. Since all five male rats at the high
concentration died during exposure, these animals may not have survived long enough to
develop the lesion. There were no exposure-related effects observed in female rats exposed to
cis-l,2-DCE. No effects were seen in the heart, kidney, or lungs in exposed rats.
4.4.1.2.2. trans-l,2-DCE. In studies similar to those conducted with the cis-isomer, DuPont
(1999) conducted a 4-hour acute inhalation study with trans-1,2-DCE, using five male and
five female Crl:CD®BR rats per exposure level. The exposure concentrations used in this study
were 0, 12,300 (48,700 mg/m3), 22,500 (89,100 mg/m3), 28,100 (111,300 mg/m3), and
34,100 ppm (135,000 mg/m3). At the 48,700 mg/m3 exposure, rats recovered and resumed a
normal appearance within about 30 minutes after the end of the exposure. There were no effects
on body weight at this concentration. Rats that survived the 89,100 mg/m3 exposure showed
lethargy and irregular respiration immediately after exposure and showed slight weight loss for
one day, followed by a normal weight gain rate. Rats exposed to 111,300 mg/m3, showed
weakness immediately after exposure and slight to severe weight loss for one day. Unlike effects
seen with cis-1,2-DCE, no compound-related effects were observed in livers of rats exposed to
trans-l,2-DCE at concentrations up to 135,000 mg/m3. No effects were seen in heart, kidneys, or
lungs of exposed rats. The LC50 was determined at 95,400 mg/m3; the authors concluded that
trans-1,2-DCE was about half as acutely toxic as the cis-isomer by the inhalation route.
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Freundt et al. (1977) exposed six mature female SPR Wistar rats/group once to
trans-l,2-DCE (purity unspecified) at concentrations of 0, 200, 1,000, and 3,000 ppm (0, 792,
3,960, and 11,880 mg/m3, respectively) for 8 hours. Test agent concentrations were monitored
by GC. Various parameters including symptoms of CNS depression, quantitative determination
of serum components (cholesterol, calcium, inorganic phosphate, total bilirubin, albumin, total
protein, uric acid, urea nitrogen, glucose, ALP), and hematological parameters (hemoglobin, cell
volume, mean corpuscular volume, mean corpuscular hemoglobin) were evaluated.
No symptoms of CNS depression were observed at the concentrations used. Clinical
chemistry values for serum albumin, BUN, and ALP were statistically significantly depressed
(11, 20, and 16%, respectively) in the rats exposed to 3,960 mg/m3; values for 11,880 mg/m3
were not reported. Freundt et al. (1977) noted pathological changes in the hearts of rats exposed
to trans-l,2-DCE after a single 8-hour exposure to 11,880 mg/m3 but not after exposures to lower
levels. These changes were described as severe fibrous swelling of the myocardium and
hyperemia. Blood leukocyte counts were reduced in rats exposed to 792 and 3,960 mg/m3, while
erythrocyte counts were significantly reduced in rats exposed to 3,960 mg/m3 trans- 1,2-DCE.
(The actual blood leukocyte counts and erythrocyte counts were not given.) Histopathological
changes were seen, including fatty degeneration of liver lobules and Kupffer cells and capillary
hyperemia with distension of the alveolar septa of lungs. At 792 mg/m3 (the TLV), slight fatty
degeneration of the liver occurred in one of six rats, and hyperemia of the lung with alveolar
septum distention occurred in all members of the group. At 3,960 mg/m3, liver degeneration was
seen in two of six rats, and lung changes were seen in five of six rats. The same incidence (2/6
rats) of liver degeneration was noted at 11,880 mg/m3, and five of six rats had lung effects. In
addition to effects on liver and lungs, severe fibrous swelling and hyperemia with barely
maintained striation of the myocardium were noted (% not indicated) at 11,880 mg/m3.
Gradiski et al. (1978) evaluated the toxicity of trans-l,2-DCE by using groups of
20 female OF1 mice exposed for 6 hours to five airborne concentrations. The LC50 was
determined graphically to be 21,723 ppm (86,000 mg/m3). On the basis of the high LC50, the
inhalation toxicity of trans- 1,2-DCE was judged by the author to be lower than that of nine other
chlorinated aliphatic solvents that were tested concurrently.
4.4.1.2.3. Mixtures of cis- and trans-l,2-DCE. Lehmann (1911) reported inhalation
experiments in which four cats were exposed to 50-72 mg/L (50,000-72,000 mg/m3) 1,2-DCE;
the isomer composition was not stated. The cats demonstrated varying degrees of narcosis; with
symptoms of salivation, sneezing, disturbance of balance, and prostration. Two of the cats died.
In acute inhalation studies (Dow, 1960), male rats (9/group) were exposed to 0, 7,297,
14,814, 16,810, 29,035, and 50,123 ppm (28,900, 58,740, 66,650, 115,120, and 198,700 mg/m3)
1,2-DCE (isomer composition not stated) for periods up to 7 hours. Rats exposed to
concentrations above 115,120 mg/m3 rapidly became unconscious, with rapid breathing and
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tremors, and exposures lasting longer than 0.2 hours were fatal. Six of nine rats exposed to
66,650 mg/m3 for 4 or 7 hours died. One hour after exposure at this concentration, the rats
developed tremors and made running movements while lying on their sides. Slight liver and
lung pathology were observed 1 day after exposures; 1 week after exposures, slight liver and
lung injury plus moderate kidney injury were reported. No rats died after exposure to
28,900 mg/m3 1,2-DCE for 7 hours or after exposure to 58,740 mg/m3 for 1 hour. No LC50 was
calculated.
4.4.2. In Vivo Neurological Behavioral Studies
Inhibition of propagation and maintenance of an electrically evoked seizure discharge
was used as a criterion for neurotropic effects in experimental animals by Frantik et al. (1994).
The studies were designed to measure the concentrations of 1,2-DCE (isomer not stated) and
47 other volatile solvents required to inhibit electrically evoked acute neurotoxic symptoms, a
measure of subclinical CNS depression. Effect-air concentration regressions for 1,2-DCE and
47 other volatile solvents were determined after 4-hour inhalation exposures in adult male
Wistar-derived rats (0.5-1 year old) and H strain female mice (2-4 months old). Four exposed
animals and four untreated controls were tested at four to five solvent concentrations, ranging
from 90 to 21,000 ppm (373 to 83,370 mg/m3) in rats and 300 to 24,000 ppm (1,191 to 95,280
mg/m3) in mice. Each experiment was repeated.
As measured by the tonic extension of hind limbs in rats and the velocity of tonic hind-
limb extension in mice, the mean latency of responses to short electrical stimuli (0.2 seconds,
50 Hz, 180 V in rats and 90 V in mice) was evaluated graphically. The critical level of effect
(the effect in the lower third part of the dose-response function corresponding to the shortening
of the tonic extension of hind-limbs by 3 seconds in rats and the lengthening of the latency of
extension by 0.6 seconds in mice) and the threshold for slowing the propagation or shortening of
the duration of seizures by 10% of the maximum effect possible (ECio) were determined. Values
for this critical level of effect were generally several times lower than airborne concentrations
evoking behavioral inhibition in animals and 1-2 orders of magnitude lower than concentrations
inducing narcosis. The mean concentration of 1,2-DCE evoking a 30% depression in response in
rats was 1,810 ppm with a one-sided 90% CI of 245 ppm and a slope of the regression line of
0.022%/ppm. Equivalent data for mice were 3,400 ppm for a 30% depression in response with a
one-sided 90% CI of 490 ppm and a slope of the regression line of 0.02%/ppm. Frantik et al.
(1994) proposed that their ECio values could be used to evaluate the efficacy of short-term
exposure limits for protection of workers from acute nervous depression and other subnarcotic
effects, such as headaches, impairment of vigilance, and lowered reliability of performance.
Concentration-dependent behavioral changes in male Swiss OF1 mice, following a
4-hour exposure to 1,2-DCE (isomer not stated) and to 12 other aliphatic and aromatic solvents,
were evaluated by DeCeaurriz et al. (1983). Tests were conducted to determine whether the test
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agent reduced immobility developed in the "behavior despair" swimming test, a test of
nonconditioned neurobehavioral performance. Concentration-related reductions in immobility
during a 3-minute test were seen for all solvents; the percent decreases in immobility vs.
exposure concentration for each agent (4-5 concentrations) were graphically depicted. The
concentration of 1,2-DCE required for a 50% decrease in immobility (ID50) was 1,983 ppm (95%
CI: 1,708-2,309 ppm). Most of the solvents tested were considerably more effective than
1,2-DCE as inhibitors of immobility; only methyl ethyl ketone and 1,1,1-trichloroethane were
less effective than 1,2-DCE. A good correlation (r = 0.93) between ID50 values and 1981
ACGIH occupational exposure standards for the chemicals tested was demonstrated.
Kallman and Balster (1983) studied disruption of reinforced operant behavior in groups
of nine mice. The animals were trained to depress a lever, and the correct behavior was
reinforced with sweetened milk. Mice were gavaged with daily doses of 100 mg/kg or more
1,2-DCE (isomer not stated) 30 minutes after the daily operant session for a minimum period of
1 week. A dose of 300 mg/kg-day or more disrupted the reinforced operant behavior. Continued
exposure at this level produced initial decreases with a gradual return to baseline performance
within 2 days. This pattern was maintained at doses below the MTD (800 mg/kg-day). When
1,2-DCE exposure was terminated, the mice recovered their conditioned behavior within
15 days.
Taste aversion to saccharin induced by 1,2-DCE in male CD-I mice was also reported by
Kallman et al. (1983). In conditioning trials for a period of 7 days, groups of seven mice were
accustomed to 30-minute sessions of drinking from two spouts that provided access to 0.3%
sodium saccharin or deionized water. Five minutes after session completion, gavage doses of
30-2,000 mg/kg 1,2-DCE in 1:9 emulphor:water were administered. Twenty-four hours after the
final conditioning treatment, the groups of mice were subjected to the 30-minute two-bottle
choice test (saccharin vs. deionized water), with careful monitoring of fluid consumption of
saccharin and water from the bottles. Doses of 300-2,000 mg/kg 1,2-DCE (but not lower
concentrations of 30 and 100 mg/kg) significantly depressed consumption of sodium saccharin
offered in the 30-minute preference test. The effective dose (ED50), the dose of 1,2-DCE that
reduced saccharin solution consumption by 50%, was graphically determined to be 144.5 mg/kg.
Intake of deionized water was also reduced when offered after daily gavage doses of 1,2-DCE
but was statistically significantly reduced only after a dose of 2,000 mg/kg. Other halogenated
compounds, such as chloral, 1,1,2-trichloroethane, and 1,2-dichloroethane, were more potent
than 1,2-DCE in inducing conditioned taste aversion. The threshold for behavioral effects of
1,2-DCE in these studies was about 100 mg/kg.
4.4.3. Immunological Studies
4.4.3.1. cis-l,2-DCE
No immunotoxicity studies of cis-1,2-DCE were located.
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4.4.3.2. trans-l,2-DCE
In a short-term assessment of immunotoxicity, 4-week-old male CD-I mice (10-
12/group) with an average initial weight of approximately 30 g were gavaged with solutions of
trans-l,2-DCE (0, 22, and 222 mg/kg, or 1/100 and 1/10 the LD50) on 14 consecutive days
(Munson et al., 1982). At necropsy, there were no significant effects on liver, spleen, lungs,
thymus, kidney, or brain weights. Leukocyte counts for the experimental groups did not differ
significantly from the untreated control group. Munson et al. (1982) evaluated humoral immune
function as indicated by the ability of the spleen cells to produce IgM antibody-forming cells
(AFCs) following challenge with sheep red blood cells (sRBCs). The authors reported the
antibody response to sRBCs challenge on day 11 as the number of AFCs per spleen and per
106 spleen cells from animals killed 24 hours after the last treatment. A trend towards
suppression of the number of AFCs expressed per spleen basis (significant at/? < 0.1 level) was
observed with trans-l,2-DCE; however, this response was not statistically significant at the
p < 0.05 level or when expressed per 106 spleen cells. Munson et al. (1982) also assessed cell-
mediated immune response as measured by the delayed-type hypersensitivity (DTH) response to
sRBCs. The response was characterized as slight but significant (p < 0.05) and not dose-
dependent in the abstract of the journal article. However, in the results section of the article the
authors stated that trans- 1,2-DCE showed no effect in the DTH response. This contradictory
presentation of the data between the abstract and results sections renders these study findings
unreliable. It is unknown whether the slight reduction in DTH was associated with trans-
1,2-DCE or trichloroethylene, another chemical tested in the study. The authors concluded that
mice exposed to trans-1,2-DCE for 14 days at doses up to 222 mg/kg-day showed no significant
change in cell-mediated or humoral immunity (Munson et al., 1982).
The immunotoxicity of trans-1,2-DCE was also investigated in studies in which three
concentrations, 0.1, 1.0, and 2.0 mg/mL, were provided to male and female CD-I mice
(10 mice/group) in drinking water containing 1% emulphor (Barnes et al., 1985; Shopp et al.,
1985). These drinking water concentrations were equivalent to doses of 17, 175, and 387 mg/kg-
day in male mice and 23, 224, 452 mg/kg-day in female mice. The study by Shopp et al. (1985)
reported assays for effects of the test agent on the immune system, while Barnes et al. (1985)
reported the study details and systemic toxicity findings.
In a preliminary 14-day study involving gavage exposure to the test agent at 0.1 or
1.0 mg/mL (21 or 210 mg/kg), Shopp et al. (1985) reported no statistically significant effects of
trans- 1,2-DCE on the humoral immune status of male mice as measured by the production of
AFCs against sRBCs. Cell-mediated immune status, measured by the DTH response to sRBCs,
was also unaffected in male mice dosed with trans-l,2-DCE for 14 days. Body weight was not
affected in male or female mice at either dose of trans-1,2-DCE in the 14-day study.
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In the same study by Shopp et al. (1985), three assays were utilized to evaluate humoral
immune status in both male and female mice following 90 days of exposure to trans-1,2-DCE in
drinking water at concentrations up to 2.0 mg/ml. These assays included quantification of spleen
AFCs directed against sRBCs on days 4 and 5 after antigen presentation, hemagglutinin titers to
sRBCs, and spleen cell response to the B cell mitogen lipopolysaccharide (LPS).
Body weight was not affected in male or female mice at any dose of trans-1,2-DCE
following 90 days of exposure. The AFC results are shown in Table 4-9. The number of AFCs
per 106 spleen cells was reduced by 26% in male mice exposed to trans-l,2-DCE at doses of
175 and 387 mg/kg-day (significantly different atp< 0.05 from control mice given deionized
water). When expressed on a per spleen basis, the numbers of AFCs in male mice were
significantly reduced at all exposure concentrations tested (equivalent to doses 17, 175, and 387
mg/kg-day). However, the expression of AFCs on a per spleen basis is affected by changes in
the relative size of the spleen. Therefore, to avoid effects due to differences in relative spleen
size, the number of AFCs per 106 spleen cells is considered the preferred measure. Spleen
weights were not significantly affected by the treatments. Females responded normally except
for mice in the 0.1 mg/mL group (23 mg/kg-day), which demonstrated a 32% decrease in AFC
response on a total spleen basis.
Table 4-9. Humoral immune response to sRBCs in CD-I mice exposed to
trans-l,2-DCE in drinking water for 90 days (day 4)
Exposure group
Spleen weight (nig)
Antibody-forming cells
per spleen ( 10~5)
Antibody-forming cells
per 106 cells
Males"
Control
202 30
4.48 0.32
2,200 125
0.1 mg/mL (17 mg/kg-day)
164 13
3.28 0.28b
2,048 152
1.0 mg/mL (175 mg/kg-day)
oo
r-
3.34 0.39b
1,625 136b
2.0 mg/mL (387 mg/kg-day)
173 10
2.87 0.37b
1,618 226b
Females"
Control
228 13
4.38 0.37
1,765 110
0.1 mg/mL (23 mg/kg-day)
176 llb
2.97 0.49b
1,478 211
1.0 mg/mL (224 mg/kg-day)
230 12
4.51 0.24
1,967 89
2.0 mg/mL (452 mg/kg-day)
191 13b
3.47 0.50
1,518 184
"Values are mean ± SE for 12 mice in the control group and 8 mice in treatment groups, measured on day 4 after
antigen presentation.
bValues differ significantly from control group (p < 0.05).
Source: Shopp et al. (1985).
Hemagglutinin titers in CD-I mice exposed to trans-1,2-DCE at all dose levels were not
significantly changed from control values. Spleen lymphocyte responsiveness to LPS was not
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altered in the males, but the female mice at the highest dose level demonstrated a statistically
significantly enhanced spleen cell response to LPS.
Three assays were also used to evaluate the status of cellular immunity: (1) DTH
response to sRBCs challenge, (2) popliteal lymph node proliferation in response to sRBCs, and
(3) spleen cell response to concanavalin A (Con A). Male mice exposed to trans-1,2-DCE did
not show changes in either the DTH or popliteal lymph node proliferation response to sRBCs,
but females exposed to 1.0 mg/mL had a slight increase in the DTH response. No alterations in
spleen lymphocyte response to Con A were noted. In addition, the ability of bone marrow cells
125
from mice exposed to trans-1,2-DCE for 90 days to incorporate I-labeled deoxyuridine was
essentially unaffected by the treatments (Shopp et al., 1985).
In summary, repeated exposure of mice to trans-1,2-DCE in drinking water for 90 days
had no effect on the cell-mediated immune status of either sex or on the humoral immune status
of females. Shopp et al. (1985) concluded that there was marked suppression in humoral
immune status in male mice and that the decrease in AFCs was significantly decreased in these
mice. However, the authors also suggested that the decrease in AFCs was not severe enough to
depress the functional ability of the humoral immune system because there was no change in
hemagglutination titers to sRBCs or lymphoproliferative response of spleen cells to the B-cell
mitogen LPS. Overall, the authors concluded that the immune system of CD-I mice was not
overly sensitive to the effects of trans-1,2-DCE and that the few effects that were seen were
probably the result of general toxicity rather than specific target organ toxicity. Additional
discussion of these study findings is given in Section 4.6.1.2.
Freundt et al. (1977) reported that inhalation exposure of female SPF Wistar rats to
>200 ppm caused slight to severe fatty degeneration of Kupffer cells in the liver. In addition,
decreased leukocyte counts were observed in rats exposed to 200 ppm and 1,000 ppm
trans-1,2-DCE for 8 hours, and pneumonic infiltration was observed in the lungs after exposure
to 200 ppm for 8 and 16 weeks, suggesting that inhalation of the test agent may have
immunological effects.
4.4.3.3. Mixtures of cis- and trans-l,2-DCE
No immunotoxicity studies of mixtures of cis- and trans-1,2-DCE were located.
4.4.4. Toxicity Studies by Other Routes
4.4.4.1. Intraperitoneal Injection
4.4.4.1.1. cis-l,2-DCE. In an acute hepatotoxicity study reported in a dissertation by McMillan
(1986), cis-l,2-DCE was administrated i.p. at doses of 21 mmol/kg (2,039 mg/kg) and
26 mmol/kg (2,521 mg/kg) in sesame seed oil to six male Sprague-Dawley rats/dose, so that each
rat received 4 mL/kg of body weight. The GSH levels and plasma enzyme activities (ALT, AST,
sorbitol dehydrogenase [SDH]), indicators of liver toxicity, were measured. The GSH and ALT
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activities were not significantly altered by either dose, but both plasma AST and SDH activities
were statistically significantly elevated. For AST, both doses had more than a twofold increase
over that seen in controls, while for SDH there was at least a threefold increase when compared
with the control.
In a study (Plaa and Larson, 1965) to obtain data regarding the relative nephrotoxic
properties of a series of chlorinated methane, ethane, and ethylene derivatives in mice, kidney
function was assessed for cis- and trans-1,2-DCE as well as other chlorinated derivatives by
measurement of the excretion of phenolsulfonephthalein (PSP) and by the use of an indicator
strip to measure protein and glucose in the urine. In addition to these tests of kidney function,
kidney sections were examined histologically. For evaluating cis-l,2-DCE's potential for
nephrotoxicity, doses of 0.1 (10 mice), 1.0 (10 mice), and 2.0 mL/kg (6 mice) were dissolved in
corn oil and administered i.p. The results show that cis-1,2-DCE failed to cause renal
dysfunction. None of the mice examined histologically showed necrosis or swelling.
4.4.4.1.2. trans-l,2-DCE. An LD50 test was performed by Freundt et al. (1977) in which mature
female SPF Wistar rats and mature female NMRI mice were exposed to trans-1,2-DCE via the
i.p. route. The LD50 was reported to be 6.0 mL/kg (95% CI: 5.1-7.1 mL/kg) (7,680 mg/kg) for
the rats. The LD50 for mice was 3.2 mL/kg (95% CI: 2.8-3.7 mL/kg) (4,096 mg/kg). The mouse
was more sensitive to the effects of trans-l,2-DCE than the rat after i.p. dosage, whereas the rat
appears to be more sensitive to trans-l,2-DCE given orally (see Section 4.4.1.1.2.). The
postmortem gross pathology in the mice after administration of trans-1,2-DCE showed
hyperemia involving the liver, kidneys, urinary bladder, and intestines. The number of dead
mice ranged from 1 to 10/10 per dose group. Clinical signs of toxicity were not reported for
mice. In one rat gross pathology included pulmonary capillary hyperemia and alveolar septal
distention and fibrous swelling and hyperemia with incipient disorganization of the cardiac
muscle.
In an acute hepatotoxicity study reported in a dissertation by McMillan (1986),
trans-l,2-DCE was administered i.p. at doses of 20 mmol/kg (1,939 mg/kg-day) and 25 mmol/kg
(2,424 mg/kg) in sesame oil (4 mL/kg) to male Sprague-Dawley rats (6/group). The results of
this set of experiments show that trans-1,2-DCE administered i.p. depressed GSH content
(statistically significant) at the 2,424 mg/kg dose and increased plasma AST and SDH activities
in a dose-related (although not statistically significant) manner. Plasma ALT activity was
unchanged. The effects of trans-l,2-DCE on GSH content and plasma ALT, AST, mid SDH
activities with respect to time were also examined. Two groups of 30 male rats each were used;
one group was the control and the other group was treated i.p. with trans-1,2-DCE. At time
intervals of 2, 4, 8, 12, 24, and 48 hours, five animals from each group were killed. Blood was
collected for plasma enzyme determination, and livers were removed for GSH analysis. SDH
activity was maximally elevated (fivefold) at 4 hours after administration of trans-1,2-DCE and
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remained elevated. AST activity was elevated over the entire time course (fivefold greater than
the control at the maximally elevated point at 4 hours), while ALT activity was elevated over
control levels for the first 8 hours of the study (fivefold greater than the control at the maximally
elevated point of 4 hours). Histopathological results of slight to moderate necrosis show that the
greatest potential hepatotoxicity occurred at 4 and 8 hours. The effects of the trans-isomer were
maximal at 4 hours after administration, using plasma enzyme elevations as the indicator of
toxicity. Glutathione depression occurred between 4 and 8 hours after administration. It is
important to note that all the parameters measured, with the exception of AST activity, returned
to near control levels by 12 hours (i.e., the effects were not sustained). The histopathological
results indicated the same time course.
The LD50 for trans-1,2-DCE in female OF1 mice via i.p. injection was reported by
Gradiski et al. (1978) to be 2,940 mg/kg. Freundt et al. (1977) reported that the i.p. LD50 for
trans-l,2-DCE was 3.2mL/kg (4,100 mg/kg) for female NMRI mice and 6.0 mL/kg
(7,680 mg/kg) for mature female Wistar rats. Nakahama et al. (2000) treated 7-week-old Wistar
rats (sex not stated) with 0.5 g/kg cis- or trans-1,2-DCE with or without co-treatment with
phenobarbital (80 mg/kg). Animals were sacrificed 24 hours after treatment, and body weights
as well as relative liver and lung weights (as compared to body weight) were measured. The cis-
1,2-DCE caused a small but statistically significant decrease in body weight gain. Both isomers
caused increases (although not statistically significant) in relative liver weights. Lung weights
were not affected. Pretreatment with phenobarbital had no noteworthy effect on these
observations.
4.4.4.2. Dermal Application
In dermal toxicity studies (Brock, 1990; DuPont, 1988b), a single dose of 5,000 mg/kg
trans-1,2-DCE was applied onto the clipped, intact skin of two male and three female New
Zealand white rabbits under an occlusive wrapping. At the end of a 24-hour exposure period, the
test material was removed. Test rabbits were examined for clinical signs of toxicity and
mortality for 14 days after treatment. No animals died, but signs of severe skin irritation
remained throughout the observation period. Mild-to-severe erythema and no-to-severe edema,
necrosis, and fissuring of the skin with raw areas and epidermal scaling were observed. Body
weight losses of up to 3% of initial weight were observed in three rabbits 1 day following
treatment. Under conditions of the assay, the dermal LD50 was greater than 5,000 mg/kg body
weight.
4.4.4.3. Eye Irritation
Brock (1990) reported results of an irritation test with trans-l,2-DCE (99.64% pure) that
was conducted at DuPont (1988c). The test agent (0.01 mL) was instilled into the lower
conjunctival sac of two female New Zealand White rabbits. Twenty seconds later, the eyes of
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one rabbit were washed with lukewarm tap water, while the eye of the other rabbit remained
unwashed. Eyes were scored for irritation at 1 and 4 hours and after 1, 2, and 3 days. Severe
corneal opacity was observed in the washed eye, and moderate iritis and conjunctivitis were
observed in both the washed and unwashed treated eyes. Copious blood-tinged discharge was
seen in both treated eyes, with moderate and mild chemosis in the washed eye and unwashed
eye, respectively. The maximum Draize score was 17/110 for the unwashed eye and 41/110 for
the washed eye. Fluorescein stain examinations were positive for corneal opacity in the washed
eye and negative in the unwashed eye. Three days after treatment the eyes of both rabbits had
returned to normal. Under conditions of the study, trans-1,2-DCE was a severe eye irritant.
Moderate pain and conjunctivitis were reported after 1,2-DCE (isomer not stated) was
administered to the eyes of rabbits (Dow, 1960). Some of the eyes were washed after
administration. Reactions to the test agent had not completely subsided 1 week after dosing.
4.4.4.4. Skin Irritation
Brock (1990) reported results of a skin irritation test with trans-l,2-DCE (99.64% pure)
conducted at DuPont (1988d). The test agent, 0.5 mL, was applied onto the clipped, intact skin
of one female and five male New Zealand White rabbits under an occlusive wrapping. At the
end of a 24-hour exposure period, the material was removed. The site of application was scored
for irritation at 24, 48, and 72 hours post treatment. Mild or moderate erythema was observed at
all observation times. Under conditions of the study, trans-l,2-DCE was a moderate skin irritant.
In skin irritation studies, 1,2-DCE mixture was applied undiluted 10 times to the intact
skin of ears of white rabbits (Dow, 1960). Essentially no irritation was reported following the
first eight applications, but slight hyperemia was observed thereafter. The ears of the rabbits
appeared normal 21 days after cessation of treatments. Four applications of undiluted 1,2-DCE
to the intact belly skin of rabbits caused slight to moderate hyperemia. Slight edema and
moderate necrosis of the skin appeared after the third and fourth applications. Undiluted
1,2-DCE was also applied twice to the abraded belly of rabbits. Slight to moderate hyperemia
and edema with slight necrosis occurred after the first application, and moderate edema and
necrosis were seen after the second application. Slight exfoliation, scabs, and scars were seen
21 days after treatments.
4.5. MECHANISTIC DATA AND OTHER STUDIES IN SUPPORT OF THE MODE OF
ACTION
4.5.1. Hepatotoxicity Studies
Trans-1,2-DCE (20 mmol [1,940 mg]/kg i.p.) depressed liver aniline hydroxylation (AH),
ethylmorphine N-demethylation (EN-D), and CYP450 content, but no effect was seen on
NADPH:cytochrome c reductase activity (McMillan, 1986). Administration of phenobarbital
with trans-l,2-DCE caused a further depression of all parameters except CYP450. Treatment
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with p-naphthoflavone and trans-l,2-DCE elevated AH and CYP450 levels but slightly
depressed NADPH:cytochrome c reductase activity. When given prior or subsequent to
trans-1,2-DCE, SKF-525A had a cumulative depressive effect on all parameters except
NADPH:cytochrome c reductase activity. According to McMillan (1986) these data indicate an
enhanced depressive effect of trans-1,2-DCE plus phenobarbital with respect to aniline
hydroxylase and ethylmorphine N-demethylase activity. The author also concluded that, with
respect to CYP450 and cytochrome c reductase, exposure to phenobarbital plus trans-l,2-DCE
partially alleviated the depression seen in these parameters with trans-1,2-DCE treatment only.
Freundt and Macholz (1978) exposed adult female Wistar rats (10/group) to single 8-hour
inhalation exposures of cis- or trans-1,2-DCE and evaluated hexobarbital sleeping time,
zoxazolamine paralysis time, and formation of 4-aminoantipyrine from aminopyrine. Exposure
levels of 0, 200, 600, or 1,000 ppm (0, 792, 2,380, or 3,960 mg/m3) of either isomer resulted in
statistically significant, concentration-related increases in hexobarbital sleeping time and
zoxazolamine paralysis time at all exposure levels, with the exception of the zoxazolamine
paralysis time for rats exposed to the trans-isomer at 200 ppm. The cis-isomer was the more
potent of the two isomers at every exposure level. Inhalation of cis- and trans-1,2-DCE for 8
hours also caused a statistically significant, exposure-dependent inhibition of renal excretion of
4-aminoantipyrine (AAP) after gavage administration of 20 mg aminopyrine immediately
following the exposures. The cis-isomer was also the more potent isomer in this experiment.
The 1,2-DCE-induced effect was reversible by 6 hours after termination of exposures.
These experiments indicated that phase I oxidative metabolism of hexobarbital,
zoxazolamine, and aminopyrine was inhibited by exposures to either isomer of 1,2-DCE; the
lowest-observed-adverse-effect level (LOAEL) was 200 ppm (792 mg/m3). The dose received
by these rats was estimated to be 198 mg/kg, based on a conversion factor of 3.96 mg/m3 per
ppm, 0.30 m3/day breathing rate for adult female rats weighing 0.2 kg, and 8 hours/day of
exposure with an estimated net inhalation retention of 50%. N-acetylation of AAP and O-
glucuronidation of 4-hydroxyantipyrine were not affected by an 8-hour exposure of rats to 1,000
ppm trans-1,2-DCE, indicating that phase II enzymes are considerably less sensitive to induction
or inhibition by 1,2-DCE. In in vitro studies, N-demethylation of aminopyrine and
O-demethylation of p-nitroanisole were competitively inhibited by addition of trans-1,2-DCE to
a reaction mixture containing liver microsomes from untreated rats (Freundt and Macholz, 1978).
On the basis of these findings, the authors concluded that 1,2-DCE competed for the type I
binding site of CYP450.
Jenkins et al. (1972) evaluated the effects of cis- and trans-l,2-DCE on enzyme activities
in liver and plasma of adult male Holtzman rats. Twenty hours after administration of 400 or
1,500 mg/kg (4.1 or 15.5 mmol/kg) cis- or trans-l,2-DCE in corn oil by gavage (2 mL/kg) to 3—
4 rats/group, liver glucose-6-phosphatase (G-6-Pase), ALP, and tyrosine transaminase activities
were statistically significantly increased by the cis-isomer; with the exception of G-6-Pase
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activity in animals receiving 400 mg/kg, these enzyme activities were not significantly elevated
after treatment with trans-1,2-DCE. In most instances, plasma ALP and ALT activities did not
differ significantly from controls after treatment with either isomer. The authors concluded that
the cis-isomer caused a slightly greater biochemical response than the trans-isomer.
Moore (1978), in an abstract, reported that administration of 0.1-1 mL/kg 1,2-DCE
(isomer not stated) to rats (strain not given) inhibited the hepatic microsomal calcium pump
24 hours after treatment in a dose-dependent manner by up to 70% at the highest dose. This
finding is of some interest in light of several other studies that reported increases in blood
calcium levels following 1,2-DCE treatment (McCauley et al., 1995; McMillan, 1986; Barnes et
al., 1985).
Sipes and Gandolfi (1980) used uninduced and phenobarbital-, 3-methylcholanthrene-, or
Aroclor 1254-induced rat liver microsomes to demonstrate that several halogenated
hydrocarbons, including 1,2-DCE (isomer not stated), bind covalently to protein and lipid, and
that binding was increased up to eightfold when induced microsomes were used. However,
1,2-DCE displayed low protein and lipid binding compared with halogenated methanes and
ethanes. Furthermore, these authors could not demonstrate any DNA binding with 1,2-DCE,
even when phenobarbital-induced microsomal preparations were used, while all other
compounds tested positive for this characteristic. The lowest DNA-binding activities were
observed with dichloromethane, iodomethane, 1,2-dichloroethane, and 1,1,1-trichloroethane,
while the highest activities came from 1,2-dibromoethane, bromotrichloromethane, chloroform,
and carbon tetrachloride (Sipes and Gandolfi, 1980).
4.5.2. Nephrotoxicity Studies
The degree of nephrotoxicity elicited by cis- or trans-1,2-DCE administered i.p. to male
Swiss mice was evaluated by monitoring urinary excretion of PSP and by detection of urinary
protein and glucose excretion with indicator strips (Plaa and Larson, 1965). The chemicals were
dissolved in corn oil and administered to 10 mice/dose at doses of 0.1, 1.0, or 2.0 mL/kg (128,
1,280, and 2,560 mg/kg) and 1.0, 2.0, or 4.0 mL/kg (1,280, 2,560, and 5,120 mg/kg) for the cis-
and trans-isomer, respectively. Surviving mice in the treatment groups where lethality occurred
exhibited a delay in excretion (<40% excreted within 2 hours) of the administered dose of 1
mg/kg PSP, while normal untreated mice excreted 67% of the administered dose within 2 hours.
Mice in dose groups where no animals died exhibited normal urinary excretion of PSP. Urine
was collected from mice surviving 24 hours for evaluation of excretion of protein and glucose.
An increased occurrence of urinary excretion of protein ( 100 mg/100 mL) in high-dose animals
(Table 4-10) was observed with both isomers. There was no detectable amount of glucose in 60
control mice, and no detectable protein in 32 of the controls; 23 controls exhibited trace amounts
of protein and 5 contained 30 mg. Histologic examination of kidneys of mice treated with 1.0
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mL/kg cis-isomer or 2.0 mL/kg trans-isomer failed to reveal proximal convoluted tubule necrosis
or swelling.
Table 4-10. Effect of 1,2-DCE isomers on urinary protein and glucose
24 hours after intraperitoneal treatment of male Swiss mice
Agent
Dose (mL/kg)
Number of mice
tested3
Number of mice with
urinary proteinb
cis-l,2-DCE
0.1
10
2
1.0
10
2
2.0
6
3
trans-l,2-DCE
1.0
10
0
2.0
10
1
4.0
5
3
"Each group originally contained 10 mice; at the high dose, only survivors were tested.
bSignificant urinary protein excretion if 100 mg/100 mL.
Source: Plaa and Larson (1965).
4.5.3. Studies with Cell Cultures
Mochida et al. (1995) used cultured human oral carcinoma cells (KB cells) to compare
the toxicity of cis- and trans-1,2-DCE. Cells were exposed to various concentrations of the test
agent for 72 hours. On the basis of cell counts, the 72h-ID5o value (which represesents 50%
inhibitory dose to growth of cells) was determined to be 3,900 jig/mL culture medium for the
trans-isomer but 5,800 jag/mL culture medium for the cis-isomer. No explanation for the relative
toxicity of the two isomers in KB cells was provided. Isomers of 1,2-DCE were considerably
less toxic to KB cells than other chlorinated organic compounds that are common contaminants
in groundwater. It is noted that these ID50 values are in the range of the solubility limits for these
isomers.
The cytotoxicity of cis-1,2-DCE to isolated rat hepatocytes was reported by Suzuki et al.
(1994). Hepatocyte cultures were incubated for 2 hours in Eagle's medium, containing 10% calf
serum and 10 mM cis-1,2-DCE. No effect was seen on release of LDH or formation of
thiobarbituric acid-reactive substances (TBARS). Extracted cellular lipids from hepatocytes
exposed to cis-l,2-DCE did not show significant increases in phospholipid hydroperoxides, in
contrast to cells exposed to carbon tetrachloride, 1,1,1-trichloroethane, and 1,3-dichloropropene.
Thus, no indication of cis-l,2-DCE-induced lipid peroxidation in hepatocyte membranes was
seen at the dose level tested.
No effect of trans-l,2-DCE on lipid peroxidation, measured as TBARS, was observed in
cultured bovine pulmonary arterial endothelial cells or rabbit aortic smooth muscle cells (Tse et
al., 1988). However, lipid peroxidation was seen in the presence of 2% (volume/volume)
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trans-l,2-DCE in Medium 199 supplemented with 10-20% fetal calf serum in the presence of
extracellular Fe(III)ADP (6.2 suggesting that a synergistic interaction between iron and the
test agent may occur (Tse et al., 1990).
4.5.4. Genotoxicity
A number of studies have evaluated the genotoxicity of cis- and trans-l,2-DCE as
individual isomers and as a mixture of both. In vitro studies include tests in prokaryotic
organisms such as Salmonella typhimurium (Ames assay) and Escherichia coli, and eukaryotic
organisms including Saccharomyces cerevisiae and Aspergillus nidulans. Genotoxic effects of
1,2-DCE have also been studied in Chinese hamster cells and human lymphocytes in vitro.
Further, in vivo studies have been conducted in the host-mediated assay, micronucleus test in
mice, and mitotic recombination in Drosophila melanogaster.
4.5.4.1. In Vitro Studies
Gene mutation studies in S. typhimurium and E. coli using cis- or trans-1,2-DCE or a
mixture of both isomers were mainly nonpositive (Mersch-Sundermann, 1989; Mersch-
Sundermann et al., 1989; Zeiger et al., 1988; Calandra et al., 1987; Strobel and Grummt, 1987;
Mortelmans et al., 1986; Nohmi et al., 1985; Cerna and Kypenova, 1977; Greim et al., 1977,
1975; Simmon et al., 1977). However, Mersch-Sundermann (1989) reported positive results in a
Salmonella strain, TA98, for the trans-isomer with or without metabolic activation by S9.
Nonpositive results were reported in the same study for the cis-isomer with or without metabolic
activation.
Strobel and Grummt (1987) tested trans-l,2-DCE at concentrations of 0.01-1 mg/plate in
Salmonella strains TA97, TA98, TA100, and TA104. Although there were increases in the
number of revertants in some strains (TA97, TA100), they were not dose-dependent. Up to a
5.5-fold increase in revertant numbers was observed in strain TA97. 1,2-DCE did not have any
effect on strains TA97 and 98 in the absence of S9. However, in the presence of S9, TA97
showed a maximum (5.5-fold) response both at doses 0.025 and at 0.25 mg/plate. Furthermore,
even at the lowest concentration (0.01 mg/plate), a fivefold increase in revertants was observed.
In the case of strain TA98, an approximate twofold increase in revertants at 1 mg/plate was
observed compared with the control. TA100 displayed up to a 2.5-fold increase in revertants at
different doses both in the absence and presence of S9. Exposure of 1,2-DCE to the TA104
strain resulted in an increase in the number of revertants both in the absence and presence of S9.
However the response was not dose-dependent. Since many strains had responses, even at the
lowest concentrations, that were close to the maximum (TA97 and TA98, +S9; TA104, + and -
S9), the results are difficult to interpret. The authors offered no discussion or rationale for the
high revertant rates that occurred at the low concentrations. Cerna and Kypenova (1977)
similarly reported decreasing number of revertants with increasing concentrations of DCE.
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However, these concentrations were lower than in assays that measured similar gene reversions
and, therefore, it is unlikely that these concentrations were causing cell toxicity.
Studies in yeast, using the diploid S. cerevisiae strain D7 for gene conversion, reverse
mutation, or mitotic recombination, were mostly nonpositive for cis- and trans-1,2-DCE (Koch et
al., 1988; Galli et al., 1982; Bronzetti et al., 1981; Simmon et al., 1977). However, Bronzetti et
al. (1984) reported positive results in S. cerevisiae D7 for both isomers with metabolic activation
and for the cis-isomer only without metabolic activation. In addition, a positive result was
reported by Koch et al. (1988) for aneuploidy in S. cerevisiae D61.M with the trans-isomer with
or without metabolic activation. However, Koch et al. (1988) cautioned that the effect noted in
the D61.M strain could have been intensified due to the long incubation period required with this
strain, and storage of the test tubes in an ice bath during part of the incubation period. Positive
results were also seen for aneuploidy and mitotic segregation in A. nidulans diploid strain PI
exposed to a mixture of both isomers (Crebelli et al., 1992; Crebelli and Carere, 1987).
No chromosome aberrations or sister chromosome exchanges were reported in Chinese
hamster cells for either cis- or trans-1,2-DCE (Sawada et al., 1987; Sofuni et al., 1985).
However, Doherty et al. (1996) investigated the activation and deactivation of chlorinated
hydrocarbons including 1,2-DCE in metabolically competent human cells. The authors used
human lymphoblastoid cell line AHH-1 (containing native CYP1A1 activity), MCL-5 (stably
expressing human CYP1A2, 2A6, 3A4, 2E1 and microsomal epoxide hydrolase) and h2El
(containing cDNA for CYP2E1) cell lines. 1,2-DCE produced an increase in micronuclei at
concentrations between 0 and lOmM in the AHH-1 and h2El cell lines. The micronuclei
contained approximately equal frequencies of both kinetochore-positive and kinetochore-
negative signals. At concentrations up to and including 10 mM, no increase in micronuclei was
observed in the MCL-5 cell line.
Tafazoli and Kirsch-Volders (1996) compared the cytototoxic, genotoxic, and mutagenic
activity of a number of chlorinated aliphatic hydrocarbons including 1,2-DCE. The mutagenicity
and cytotoxicity of 1,2-DCE was evaluated in an in vitro micronucleus assay using human
lymphocytes in the presence or absence of S9. A low but positive response (p < 0.05) was
obtained at 20 mM concentration both with and without S9. The authors stated that this increase
was not accompanied by a substantial decrease in cell proliferation. In addition to the
micronucleus assay, a comet assay was employed to examine the capacity of 1,2-DCE to induce
DNA damage in in vitro isolated human lymphocytes. Positive responses for tail length were
found at 6 and 8 mM (p < 0.01) and for tail movement at 2 mM (p < 0.01) with S9. A summary
of the in vitro genetic toxicology studies is presented in Table 4-11.
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Table 4-11. In vitro genotoxicity studies using cis- and trans-1,2-
dichloroethylene
Strain/
Result
Test system
cell line
-S9
+S9
Dose/plate
Compound
Effect
Reference
Bacterial systems
S. typhimurium
TA98
TA100
TA1535
TA1538
TA1950
TA1951
TA1952
-
NT
0.5-50 jaL
cis, trans
Reverse
mutation
Cerna and
Kypenova, 1977
TA1535
TA1538
-
-
NA
cis, trans
Reverse
mutation
Greim etal., 1977
TA98
TA100
TA1535
TA1537
TA1538
-
NT
Up to 5 mg
cis, trans
Reverse
mutation
Simmon et al.,
1977
TA98
TA100
TA1535
TA1537
-
-
33-5,555 jag
lO'-lO4 |ag
mixture
trans
Reverse
mutation
Mortelmans et al.,
1986
TA97
TA98
TA1535
TA1537
-
-
33-10,000 jag
cis
Reverse
mutation
Zeiger et al., 1988
TA97a
TA98
TA100
TA102
NT
-
NA
trans
Reverse
mutation
Calandra et al.,
1987
TA97
TA98
TA100
-
-
NA
cis
Reverse
mutation
Mersch-
Sundermann, 1989
TA97
TA98
TA100
+
+
+
+
NA
trans
Reverse
mutation
Mersch-
Sundermann, 1989
TA97
TA98
TA100
TA104
+
+?
+?a
+?
+
+?
0.01-1.0 mg
trans
Reverse
mutation
Strobel and
Grummt, 1987
TA87
TA98
TA100
TA102
-
-
1.0-50 mg
cis
Reverse
mutation
Nohmi et al., 1985
E. coli
K12
-
-
2.9 mM
2.3 mM
cis
trans
Greim etal., 1977,
1975
PQ37
-
-
NA
cis, trans
DNA damage
Mersch-
Sundermann et al.,
1989
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Table 4-11. In vitro genotoxicity studies using cis- and trans-1,2-
dichloroethylene
Strain/
Result
Test system
cell line
-S9
+S9
Dose/plate
Compound
Effect
Reference
S. cerevisiae
D7
-
NA
100 mM
cis, trans
Bronzetti et al.,
1981
D7
+
+
+
100 mM
40 mM
100 mM
80 mM
cis
trans
Gene
conversion,
reverse
Bronzetti et al.,
1984
D7
-
-
100 mM
cis, trans
mutation, or
mitotic
recombination
Galli et al., 1982
D3
-
NT
Up to 0.2 mL
cis, trans
Simmon et al.,
1977
D7
-
-
77.3 mM
trans
Koch et al., 1988
D61.M
+
+
77.3 mM
trans
Aneuploidy
Koch et al., 1988
A. nidulans
Diploid PI
+
NT
1-2.5 mL
in 20 L
(24-hour
vapor)
mixtureb
Mitotic
recombination,
mutation
Aneuploidy
Crebelli and
Carere, 1987
Diploid PI
+
NT
0.05-0.175%
(v/v)
mixture
Mitotic
segregation
Crebelli etal., 1992
Mammalian cells
Chinese hamster
CHL
-
-
7.5 mg/mL
cis
Chromosome
aberrations
Sofunietal., 1985
V79 lung
+
NT
6.5 x 10~3 M
trans
c-Mitosis,
aneuploidy
Onfelt, 1987
CHL
-
-
2.0 mg/mL
cis, trans
Chromosome
aberrations,
sister chromatid
exchange
Sawada et al., 1987
CHO
ND
?
?
160-5,000
Mg/mL
cis
trans
Sister chromatid
exchange
Galloway et al.,
1987
CHO
+
+
126-12,630
Mg/mL
mixture
Galloway et al.,
1987
CHO
-
-
500-5,000
1,600-5,000
455-12,630
Mg/mL
cis
trans
mixture
Chromosome
aberrations
Galloway et al.,
1987
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Table 4-11. In vitro genotoxicity studies using cis- and trans-1,2-
dichloroethylene
Test system
Strain/
cell line
Result
Dose/plate
Compound
Effect
Reference
-S9
+S9
Human
lymphoblastoid
AHH-1
h2El
MCL-5
+
+
NT
2.5 mM
2.5 mM
10 mM
mixture
Micronucleus
assay
Doherty etal., 1996
Human
lymphocytes
+
+
20 mM
mixture
Micronucleus
assay
Tafazoli and
Kirsch-Volders,
1996
+
+
6 mM
4 mM
mixture
Comet assay,
DNA breakage
Tafazoli and
Kirsch-Volders,
1996
increase in revertants in mid-dose range, decrease at high doses; poor dose response (see text).
bAuthors state CASRN for mixture, but chemical name is given as 1,2-dichloroethane.
+ = positive; - = nonpositive; ? = inconclusive; NT = not tested; NA = not available
4.5.4.2. In Vivo Studies
In the host-mediated assay in mice, Cerna and Kypenova (1977) reported an increase in
mutation and chromosomal aberrations for the cis-isomer, with no increase noted for the trans-
isomer. Also, in a similar host-mediated assay, Bronzetti et al. (1984) reported positive results
for the cis-isomer and nonpositive results for the trans-isomer. No increase in micronucleus
induction was reported in the bone marrow of CD-1 mice exposed by i.p. injection to a mixture
of the cis- and trans-isomers (Crebelli et al., 1999). Since none of the 10 halogenated aliphatic
hydrocarbons studied (including 1,2-DCE) showed any evidence of micronucleus induction, the
authors concluded that the in vivo mouse bone marrow test may not be sensitive enough to detect
the genotoxic effects of this group of compounds. However, an increase in mitotic
recombination was observed in Drosophila larvae exposed to the vapors of a mixture of both
isomers at 2,000 ppm (Vogel and Nivard, 1993). See Table 4-12 for a summary of the in vivo
genetic toxicology studies using cis- and trans-1,2-DCE.
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Table 4-12. In vivo genotoxicity studies using cis- and trans-
1,2-dichloroethylene
Test system
Strain/cells
Result3
Dose
(LED/HID)b
Compound
Effect
Reference
Host: mouse;
S. cerevisiae
CD
D7
-
3,000 mg/kg
cis, trans
Host-mediated
assay
Bronzetti et al.,
1981
+
1,300 mg/kg
cis
trans
Bronzetti et al.,
1984
Host: mouse;
S. typhimurium
ICR
TA1950
TA1951
TA1952
+
*/2, 1 LD50
(i.p.)
cis
trans
Cerna and
Kypenova, 1977
Mouse, female
Bone marrow
+
5 1/6 LD50
(i.p.)
cis
trans
Chromosomal
aberrations
Cerna and
Kypenova, 1977
Mouse, male
-
500-2,000
mg/kg
cis, trans
Tice etal., 1987
Mouse, male and
female
Peripheral
erythrocytes
-
280^490
mg/kg (i.p.)
mixture
Micronucleus
test
Crebelli et al.,
1999
Mouse, male and
female
-
3,125-50,000
ppm in feed for
14 weeks
trans
MacGregor et al.,
1990
Mouse, male
Bone marrow
-
500-2,000
mg/kg
cis, trans
Sister chromatid
exchange
Tice etal., 1987
D. melanogaster
larvae
Cross of
y w
+
2,000 ppm
(vapor)
mixture
Eye mosaic
assay
Vogel and Nivard,
1993
a+ = Positive; - = nonpositive.
bLED = lowest effective dose, HID = highest ineffective dose.
In conclusion, both cis-l,2-DCE and trans- 1,2-DCE have been evaluated for genotoxicity
and mutagenicity using various in vitro and in vivo assays in both non-mammalian and
mammalian systems. Most gene mutation assays both in S. typhimurium strains and E. coli were
nonpositive as a result of exposure to 1,2-DCE. Studies in yeast, using the diploid S. cerevisiae
strain for gene conversion, reverse mutation, or mitotic recombination, were also mostly
nonpositive for cis- and trans-1,2-DCE. No chromosome aberrations or sister chromosome
exchanges were reported in Chinese hamster cells when exposed to either isomer of 1,2-DCE;
however, micronucleus formation was observed in human lymphocytes. Overall, data for
1,2-DCE are generally nonpositive for genotoxicity and mutagenicity. The positive studies are
inconsistent and need further confirmation.
4.5.5. Quantitative Structure-Activity Relationship Studies
Greim et al. (1975) used a number of chlorinated ethylenes in an E. coli mutation assay
with metabolic activation by S9. They observed that ethylenes with an asymmetric arrangement
of the chlorines across the double bond were mutagenic (vinyl chloride > trichloroethylene >
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1,1 -DCE), while those with a symmetric arrangement (tetrachloroethylene, cis- and
trans-1,2-DCE) were not. No mutagenic activity was observed in any test in the absence of S9.
Greim et al. (1975) explained this finding on the basis of likely differences in the chemical
stabilities of the respective oxiranes that were formed by S9. Jones and Mackrodt (1982)
developed a theoretical model for oxirane reactivity, specifically targeting the energy required to
split the weaker of the two C-O bonds in the oxirane ring. They compared these energies to the
mutational potencies observed by Greim et al. (1975) and confirmed the C-O bond split energy
to be a good predictor of mutational potency. In a correlation of mutagenic potency vs. C-O
bond split energy, there was a region of no mutational potency with decreasing bond strength,
and both cis- and trans-1,2-DCE fell within that portion of the curve. There followed a region
where mutational potency increased strongly with decreasing bond strength (trichloroethylene;
maximal potency with chloroethylene) but then turned and decreased with decreasing bond
strength; this part of the curve was represented by 1,1-DCE. In a subsequent paper, Jones and
Mackrodt (1983) included carcinogenicity data for several halogenated ethylenes in their
calculations and found that carcinogenicity and mutagenicity data formed almost overlapping
bell-shaped curves when correlated with C-O bond split energy.
Loew et al. (1983) used a molecular orbital method to evaluate the carcinogenic potencies
of chloroethylenes, including cis- and trans-1,2-DCE. They proposed that CYP450 metabolism
of the parent compound results in an initial radical intermediate from which either epoxides or
reactive carbonyl compounds could be formed, suggesting three possible alternate pathways of
toxic activation. The authors considered the carbonyl compound (i.e., the acyl chloride or
aldehyde) rather than the epoxide to be the ultimate carcinogen in a genotoxic (DNA adduct
formation) or epigenetic (macromolecule alkylation and necrosis) process. Their findings
indicated that the amount of reactive carbonyl compound formed by metabolism, rather than its
electrophilic reactivity, was a determinant of carcinogenic potency. For the purpose of their
evaluations, they assumed that carbonyl compounds were formed by both the radical and the
epoxidation pathway and that the protonated forms of the epoxides could also act as the ultimate
carcinogens.
Loew et al. (1983) used three parameters as molecular indicators of carcinogenic
potency: the activation energy required to create a reactive intermediate in all three alternate
pathways, the electrophilic potency of the putative active carcinogen to form a covalent bond,
and its long-range electrostatic interactions. They used the "Modified Neglect of Diatomic
Overlap" method in their calculations. Activation energy turned out to be a useful predictor only
for compounds with few chlorine substituents, and predictions became increasingly inaccurate
with increasing degree of chlorination. Electrophilicity turned out to be unsuitable as a predictor
of carcinogenicity. Assuming the carbonyl compounds to be the ultimate carcinogen, the authors
proposed a carcinogenicity ranking of vinyl chloride = 1,1-DCE > 1,2-DCE >
tetrachloroethylene > trichloroethylene. Assuming the epoxide carbocation as the ultimate
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carcinogen, the authors proposed a similar ranking in which 1,2-DCE was placed tentatively in
the same position. Vinyl chloride is a confirmed human and animal carcinogen; the other
compounds, with the exception of 1,2-DCE, are all animal carcinogens without evidence of
carcinogenicity in humans. The approach and ranking, as proposed by Loew et al. (1983), are
therefore of limited use. In the evaluation of Jones and Mackrodt (1983), vinyl chloride was
listed as about 100 times as potent a carcinogen as 1,1-DCE, with no oncogenicity attributed to
1,2-DCE.
Crebelli et al. (1995) evaluated a set of 55 halogenated hydrocarbons for their ability to
induce mitotic chromosome malsegregation, mitotic arrest, and lethality in A. nidulans. The
1,2-DCE isomer mixture was about one-half as potent in inducing malsegregation as 1,1-DCE
but only slightly less potent in inducing growth arrest. The most potent malsegregation inducers
in this test were 1,1,1-trichloropropene and 1,1,2,2-tetrabromoethane, with more than 20 times
the potency of 1,2-DCE. Tetrabromomethane was about 2,500 times more effective in arresting
A. nidulans growth than 1,2-DCE. The quantitative structure-activity relationship (QSAR)
evaluations did not include 1,2-DCE; although the authors reported high correlation coefficients
between measured and calculated values for 20 compounds that were subjected to QSAR, a
visual inspection of the values for individual compounds was not convincing.
Liu et al. (1997) used the computer software MultiCASE (designed to identify as yet
unknown portions in a chemical structure that confer specific reactivity) to evaluate
93 chemicals, including trans-1,2-DCE and several mono-, tri-, or tetrachlorinated alkanes and
alkenes that had been tested experimentally for their ability to induce chromosome
malsegregation in S. cerevisiae. A subset of the NTP salmonella mutagenicity database was used
for comparison. They identified the trans-1,2-DCE structure and the vinylidene chloride portion
of the tetrachloroethylene structure as having activity in the malsegregation assay.
Cronin (1996) used the original values of Frantik et al. (1994) (see Section 4.4.2) for a
QSAR analysis of the neurotoxicity of 47 compounds that comprised benzene and many of its
alkylated congeners, halogenated alkanes and alkenes (including the 1,2-DCE mixture), alcohols,
ketones, esters, and a few unsubstituted alkanes (n-pentane, n-hexane, n-heptane, and
cyclohexane). In a first attempt, using rat data, stepwise regression through all 44 data sets (that
included boiling and melting points, a hydrophobicity factor, and several other specific
molecular parameters for each compound) afforded an equation to describe the neurotoxicities of
these compounds that included the respective boiling points and specific molecular parameters
but no term for hydrophobicity. After removing four evident outliers from the data set—
interestingly, those were the four unsubstituted alkanes mentioned above—another equation was
obtained that was based only on a specific hydrophobicity factor. By using mouse data, an
equation was obtained that again did not contain the hydrophobicity factor. The same four
unsubstituted alkanes were outliers, and their removal resulted in an equation that was based on
boiling point, hydrophobicity, and an additional molecular factor. Regression curves described
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by the respective equations for rats and mice had similar slopes and intercepts. The author was
able to obtain a description that separated highly neurotoxic compounds from less neurotoxic
ones. 1,2-DCE fell in the middle range of the 47 chemicals tested for neurotoxicity. The author
pointed out that solubility of the compounds played a minor role in their neurotoxicity, while
hydrophobicity was more useful in predicting this effect.
4.6. SYNTHESIS OF MAJOR NONCANCER EFFECTS
A general overview of toxicity studies conducted with cis- or trans-1,2-DCE indicates
that both compounds display low toxicity. Although most of the 1,2-DCE literature suggests that
the liver is the organ most affected by exposure to 1,2-DCE at high doses, evidence of any
specific pathological event is limited based on the available information (NTP, 2002; DuPont,
1998; McCauley et al., 1995; Hayes et al., 1987; Freundt et al., 1977). Table 4-13 presents a
summary of the major subchronic studies and the observed effects for both oral and inhalation
exposure to cis- and trans-1,2-DCE.
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Table 4-13. Summary of major noncancer subchronic studies for oral exposure to 1,2-DCE
Reference
Isomer
Dosing
vehicle
Treatment
period
Species,
number of
animals
Dose (mg/kg-day)
Effects observed at the LOAEL
NOAELa
(mg/kg-d)
LOAELb
(mg/kg-d)
NTP
(2002)
trans
Oral
(feed)
14 weeks
Rat,
9-10/sex/dose
M: 0, 190, 380, 770,
1,540, or 3,210
F: 0, 190, 395, 780,
1,580, or 3,245
M: Decreased BW gain (6%) at the high
dose was not considered a LOAEL.
F: | Rel liver wt (10%).
The biological significance of I RBC
count in M (>380) andF (>1,580) was
unclear and not used to identify the
LOAEL.
M: 3,210
F: 190
M:ND
F: 395
M: 0, 480, 920, 1,900,
3,850, or 8,065
F: 0, 450, 915, 1,830,
3,760, or 7,925
M: t Rel liver wt (9%). At 8,065, j BW
NTP
(2002)
trans
Oral
(feed)
14 weeks
Mouse,
10/sex/dose
gain (-7%).
F: t Rel liver wt (11%). Decreased body
weight gain (6%) at 1,830 was not
considered a LOAEL.
M: 920
F: 1,830
M: 1,900
F: 3,760
M: 0, 402, 1,314, or
Hayes et
al. (1987)
trans
Oral
(dw)
90 days
Rat,
20/sex/dose
3,114
F: 0,353, 1,257, or
2,809
F: | Abs kidney wt (8%)
M: 3,114
F: 353
M:ND
F: 1,257
M: | Rel liver wt (8%)
Barnes et
al. (1985)
trans
Oral
(dw)
90 days
Mouse,
15-23/sex/dose
M: 0, 17, 175, or 387
F: 0, 23, 224, or 452
F: | Rel thymus wt (18%)
Changes in clinical chemistry parameters
were sporadic and not used to identify a
LOAEL.
M: 17
F: 23
M: 175
F: 224
Shopp et
al. (1985)
trans
Oral
(dw)
90 days
Mouse,
8-12/sex/dose
M: 0, 17, 175, or 387
F: 0, 23, 224, or 452
M: | sRBC-responsive cells (26%)
M: 17
F: 452
M: 175
F:ND
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Table 4-13. Summary of major noncancer subchronic studies for oral exposure to 1,2-DCE
Reference
Isomer
Dosing
vehicle
Treatment
period
Species,
number of
animals
Dose (mg/kg-day)
Effects observed at the LOAEL
NOAELa
(mg/kg-d)
LOAELb
(mg/kg-d)
McCauley
et al.
(1995)
cis
Gavage
90 days
Rat,
10/sex/dose
Reported: 0, 10, 32,
98, 206 mg/kg-day;
EPA calculated: 0, 32,
97, 291, 872
M: t Rel. kidney wt (14%). Liver wt
significantly increased at >97.
F: | Rel. liver wt (14%)
Changes in clinical chemistry and
hematology parameters were sporadic and
not used to identify the LOAEL.
M: ND
F: 32
M: 32
F: 97
''NOAEL = No-observed-adverse-effect level
bLOAEL = Lowest -observed-adverse-effect level
M = males; F = females; t = increase; I = decrease; BW = body weight; abs = absolute; rel = relative; wt = weight; dw = drinking water; ND = not determined
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Table 4-14. Summary of major noncancer subchronic studies for inhalation exposure to 1,2-DCE
Reference
Isomer
Treatment
period
Species,
number of
animals
Concentration (mg/m3)
Effects observed at the
LOAEL
NOAELa
(mg/m3)
LOAELb
(mg/m3)
Dow
(1962)
report
mix
7 hr/day for
6 months
Rat,
12-
35/sex/conc.
0, 1,980, or 3,960
M: t Rel liver and kidney wt
F: t Rel liver and kidney wt
M: ND
F: ND
M: 1,980
F: 1,980
Freundt
(1977)
trans
8 hr/day, 5
days/week
for 1, 2, 8,
and 16
weeks
Rat (female),
6/conc.
0, 792
F: Fat accumulation in the
liver and Kupffer cells
F: NA
F: 792
DuPont
(1998)
trans
6 hr/day, 5
days/week
for 90 days
Rat,
15/sex/conc
0, 792, 3,960, or 15,800
M and F: Decreased
lymphocyte count reported
(statistically significant in
high-concentration M only);
biological significance
unclear and not used to
identify the LOAEL.
M: 15,800
F: 15,800
M: ND
F: ND
"NOAEL = No-observed-adverse-effect level
bLOAEL = Lowest -observed-adverse-effect level
M = males; F = females; t = increase; I = decrease; rel = relative; wt = weight; cone = concentration; ND = not determined
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4.6.1. Oral
4.6.1.1. cis-l,2-DCE
No studies of the effects of oral exposure to cis-l,2-DCE in humans were identified, and
the experimental toxicity database for this isomer is limited. The only investigation of repeat-
dose toxicity of cis-l,2-DCE by the oral route is McCauley et al. (1995, 1990). As noted in
Sections 4.2.1.1 and 4.2.1.2, comparison of the unpublished report (McCauley et al., 1990) and
the published study (McCauley et al., 1995) revealed certain errors and inconsistencies in the
documentation of the study protocol and results that were determined not to have compromised
the integrity of the data since the inconsistencies were more likely an issue of quality of the
report writing than an issue with the findings themselves.
McCauley et al. (1995, 1990) reported statistically significant increases in relative (to
body weight) liver weights in rats following both 14- and 90-day exposures (up to 38 and 32% in
males and 38 and 30% in females, respectively). These investigators reported that there were no
histopathological changes in the liver. Clinical chemistry indicators of liver function were
limited to serum AST. There were no statistically significant changes in AST levels at any dose
in either sex. The absence of compound-related histopathological changes in the liver, the
absence of AST changes, and the lack of measurements of other clinical chemistry indicators of
liver function in the McCauley et al. (1995, 1990) study create difficulties in interpreting the
relative liver weight findings.
Limited data from studies of shorter (acute and short-term) duration suggest that the liver
is a target organ for cis-1,2-DCE, although responses, at least under the conditions studied, were
minimal. Liver enzymes were measured in an acute toxicity study of cis-l,2-DCE (McMillan,
1986). Following a single oral gavage dose of 4,944 mg/kg-day cis-l,2-DCE, GSH levels were
elevated by 28% and AST activity by 56%; ALT activity was unchanged. Minimal hepatic
centrilobular vacuolation was reported in rats exposed to lethal or near lethal concentrations
(>47,900 mg/m3) of cis-l,2-DCE in an acute inhalation study (DuPont, 1999), and concentration-
related increases in relative liver weights were reported in male and female rats in the 14-day
oral toxicity study by McCauley et al. (1995, 1990).
In studies of 1,2-DCE mixtures, slight liver pathology was observed in rats 1 week after
an acute 7-hour inhalation exposure to 66,740 mg/m3 1,2-DCE (isomer composition not stated)
—a concentration lethal to 6 of the 9 exposed rats (Dow, 1960)—and relative liver weight was
increased following exposure to 1,2-DCE (58% cis-, 42% trans-isomer) by inhalation
intermittently for up to 195 days at concentrations of 1,980 mg/m3 (rabbits) or 3,960 mg/m3
(rats) (Dow, 1962). In the Dow (1962) study, clinical chemistry values were reported to fall
within normal values. Overall, these studies show elevated liver weight to be the most consistent
finding in studies of cis-l,2-DCE or mixed isomers of 1,2-DCE. Reported increases in liver
enzymes have been small, and slight liver pathology has been documented only following acute
inhalation exposure at or near lethal exposure concentrations.
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The increased liver weight observed in the McCauley et al. (1995, 1990) study was
related to administration of cis-l,2-DCE; however, in the absence of elevated liver enzymes or
histopathology, the change in liver weight is difficult to interpret.
Increased relative kidney weight (as a percent of body weight) was also observed by
McCauley et al. (1995, 1990). In male rats, statistically significant increases in relative kidney
weight were observed at all doses in the 90-day study but not in the 14-day study. In female rats,
relative kidney weight was not statistically elevated following 90 days of exposure, but was
elevated in the two highest dose groups following 14 days of exposure. The absence of
compound-related histopathological changes in the kidney in the McCauley et al. (1995, 1990)
study raises questions about the biological significance of the relative kidney weight findings.
BUN and creatinine, two clinical chemistry parameters that are indicators of kidney function, did
not provide supporting evidence for functional damage to the kidney (McCauley et al., 1995,
1990). In the 90-day study, BUN and creatinine were only marginally decreased (although
statistically significant) in high-dose (872 mg/kg-day) male rats; values in treated females were
similar to controls.
The observed increases in liver and kidney weight could represent early indicators or
precursors of liver and kidney toxicity; however, toxicity information is limited. Additionally, it
is not possible to predict whether liver or kidney damage would or would not occur at higher
concentrations or in studies of longer exposure duration (i.e., chronic studies).
McCauley et al. (1995, 1990) also reported decreases in hematological parameters
(hemoglobin and hematocrit) in male and female rats at doses >97 mg/kg-day that were not dose-
related and were considered by the study investigators to be marginal. Comparison of
hemoglobin and hematocrit findings with normal values for Sprague-Dawley rats suggests that
these hematological parameters were not affected by cis-l,2-DCE treatment. Based on blood
samples collected from 25 male and 25 female Sprague-Dawley rats, Leonard and Rubin (1986)
reported the following means (and ranges): hemoglobin—16.1 g/dL (13.3-17.3 g/dL) in males
and 16.2 g/dL (14.6-17.2 g/dL) in females and hematocrit—42.4% (36.7-46.4%) in males mid
41.6% (37.6^14.3%) in females. With the exception of the 291 mg/kg-day-dosed female rats,
values for hemoglobin and hematocrit in cis-l,2-DCE-treated rats were within the normal range
reported by Leonard and Rubin (1986). Matsuzawa et al. (1993) examined hematological data
from >2,700 male and >2,600 female Sprague-Dawley rats. Values within two standard
deviations of the mean were considered by the study authors to be within the normal range. For
hemoglobin and hematocrit in male and female rats, two standard deviations from the mean as
reported by Matsuzawa et al. (1993) is equivalent to approximately 12-13% of the mean. This
compares to decreases in hematological parameters in the McCauley et al. (1995, 1990) study of
only 6-10% of the control mean. Thus, based on normal ranges for hematologic parameters, the
hematology findings in McCauley et al. (1995, 1990) are not considered biologically significant.
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There is limited evidence that oral exposures to cis-l,2-DCE affect the CNS. In a 14-day
oral gavage study, McCauley et al. (1995, 1990) reported signs of CNS depression in male and
female rats. Immediately following gavage dosing, animals appeared agitated followed by
lethargy and ataxia. In the 90-day study (McCauley et al., 1995, 1990), however, the
investigators reported no compound-related clinical effects.
There are no oral studies of chronic, reproductive, or developmental toxicity of
cis-l,2-DCE. The findings from developmental toxicity range-finding studies for a mixture of
1,2-DCE isomers (NTP, 1991a, b, c) are summarized in Section 4.6.1.3.
4.6.1.2. trans-l,2-DCE
No human studies involving oral exposure to trans-1,2-DCE were identified. The oral
toxicity of trans-1,2-DCE was evaluated in four subchronic toxicity studies—NTP (2002) (rats
and mice), Barnes et al. (1985) (mice), Hayes et al. (1987) (rats), and Shopp et al. (1985) (mice).
The drinking water study by Barnes et al. (1985) exposed mice at doses up to approximately 400
mg/kg-day, whereas the drinking water study by Hayes et al. (1987) and dietary study by NTP
(2002) exposed mice and rats to doses almost an order of magnitude higher. These three studies
identified a range of effects associated with trans- 1,2-DCE exposure, including decreased body
weight gain, effects on organ weights (liver, kidney, thymus, and lung), minimal changes in liver
function enzymes, decreased mean body weight, and minimal decreases in hematological
parameters. A fourth subchronic study in mice (Shopp et al., 1985) evaluated the immunotoxic
potential of trans-1,2-DCE. Information on developmental toxicity is limited to developmental
range-finding studies on a mixture of 1,2-DCE isomers (NTP, 1991a, b, c), which are
summarized in Section 4.6.1.3. No chronic bioassays of trans-1,2-DCE toxicity have been
performed.
Statistically significant effects on the liver were observed by Barnes et al. (1985) and
NTP (2002), but not Hayes et al. (1987). In the 90-day Barnes et al. (1985) study, male and
female mice were exposed to trans-l,2-DCE in drinking water at doses up to 387 mg/kg-day for
males and up to 452 mg/kg-day for females. A significant increase in mean liver weights was
noted at the mid-dose (175 mg/kg-day), but not at the highest dose, in male mice. No DCE-
induced changes in terminal body weight were observed. Significant increases in serum ALP
levels of 62 and 33% were reported at the 175 and 387 mg/kg-day doses, respectively, in male
mice. These increases showed no dose-response relationship, were within the normal range for
this mouse strain, and were not observed in female mice. In female mice, ALT and AST levels
were depressed at all doses, with statistical significance at the two highest dose levels. Increases
in ALT and AST levels are indications of liver damage; the implication of decreases in these
enzymes is unknown. The findings of Barnes et al. (1985) suggest that trans-1,2-DCE, via
drinking water, does not induce hepatotoxicity at doses up to 387 mg/kg-day in male mice and
up to 452 mg/kg-day in female mice.
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NTP (2002) conducted a 14-week dietary study of trans-1,2-DCE in rats and mice at
doses ranging from approximately 190 to 3,200 mg/kg-day in rats and approximately 450 to
8,000 mg/kg-day in mice. Table 4-5 shows the relative liver weight changes (expressed as a
ratio of liver weight to body weight) in mice and rats. Absolute and relative liver weights of
female rats exposed to >395 mg/kg-day were statistically significantly higher by 8-17% and 6—
10%, respectively, than those of the vehicle controls; liver weights of male rats were not affected
by trans-1,2-DCE exposure. In mice, relative liver weights were statistically significantly
increased over controls in males (by 9-15%) exposed to doses of >1,900 mg/kg-day and in
females (by 11%) exposed to doses of >3,760 mg/kg-day. Clinical chemistry data did not
suggest hepatotoxicity in either species. Statistically significant decreases in serum ALP
activities were reported in female rats exposed to the three highest doses compared with the
vehicle controls; these decreases were minimal in severity (<13%) and transient (i.e., present at
day 21 but not week 14). No exposure-related changes in ALP activities were observed in male
rats or mice of either sex. No changes were observed in other clinical chemistry parameters,
including cholesterol, ALT, and SDH levels, in rats or mice of either sex.
As with cis-1,2-DCE, consideration was given to the possibility that the observed liver
effects for trans-1,2-DCE are early indicators or precursors of liver toxicity. It is not possible to
predict whether liver damage would or would not occur at higher concentrations or in studies of
longer exposure duration (i.e., chronic).
In interpreting the liver weight findings from subchronic oral exposure to trans-1,2-DCE,
consideration was given to the entire database for trans-1,2-DCE, including acute oral gavage
studies, inhalation study findings, and 1,2-DCE mixture information. McMillan et al. (1986)
reported no effects on ALT in rats following a single gavage dose of 4,944 mg/kgtrans-
1,2-DCE; acute i.p. injection caused transient increases in liver enzymes and histopathology
(maximally elevated at 4 hours and near control levels at 12 hours post exposure). In an oral
LD50 test in female Wistar rats (Freundt et al., 1977), severe fatty infiltration of the liver lobules
and Kupffer cells was observed in some animals receiving a single oral gavage dose of trans-1,2-
DCE. Barnes et al. (1985) indicated that based on gross necropsy findings, the liver was a target
following single acute gavage administration of trans-1,2-DCE. By inhalation, Freundt et al.
(1977) found fatty effects on the liver following single and repeated (up to 16 weeks) exposures
to 792 mg/m3; in contrast, DuPont (1999) found no compound-related liver pathology in rats
following acute inhalation exposures at lethal concentrations (48,700-135,000 mg/m3) or a 90-
day inhalation exposure at concentrations up to 15,800 mg/m3. In studies of 1,2-DCE mixtures,
slight liver pathology was observed in rats one week after an acute 7-hour inhalation exposure to
66,740 mg/m3 1,2-DCE (isomer composition not stated)—a concentration lethal to 6/9 exposed
rats (Dow, 1960)—and relative liver weights were increased following exposure to 1,2-DCE
(58% cis-, 42% trans-isomer) by inhalation intermittently for up to 195 days at concentrations of
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1,980 mg/m3 (rabbits) or 3,960 mg/m3 (rats) (Dow, 1962). In the Dow (1962) study, clinical
chemistry values were reported to be within normal values.
The increased liver weight observed in NTP (2002) and Barnes et al. (1985) was related
to administration of trans-1,2-DCE; however, in the absence of elevated liver enzymes or
histopathology, the change in liver weight is difficult to interpret.
In the 90-day drinking water study by Hayes et al. (1987), kidney weight (absolute and
relative to body weight) was statistically significantly increased in female rats (by 11 to 13%) at
doses of 1,257 and 2,809 mg/kg-day trans- 1,2-DCE, but not in male rats in any dose groups. The
kidney weight changes in female rats were not accompanied by histopathologic changes. In the
dietary study by NTP (2002), absolute kidney weight was decreased (up to 9%) in female rats
(1,580 and 3,245 mg/kg-day) and female mice (7,925 mg/kg-day), but relative kidney weight (as
a ratio to body weight) was similar to controls in all dosed groups. No gross or histopathological
lesions in the kidney were observed in rats or mice that were attributed to exposure to trans-
1,2-DCE (NTP, 2002). Similarly, clinical chemistry findings, BUN, creatinine, total protein, and
albumin levels, did not provide evidence of any functional changes in the kidney. NTP (2002)
observed that sporadic differences in clinical chemistry parameters at various time points
generally did not demonstrate an exposure response relationship or were inconsistent between
males and females.
Overall, the findings from Hayes et al. (1987) and NTP (2002) provide limited evidence
that trans-1,2-DCE affects the kidney. The findings from these two studies are inconsistent, with
Hayes et al. (1987) reporting an increase in relative kidney weight and NTP (2002) reporting a
decrease. Neither NTP (2002) nor Hayes et al. (1987) found any treatment-related
histopathological changes of the kidney in rats and mice. Additionally, NTP (2002) did not find
any clinical chemistry changes indicative of nephrotoxicity. Therefore, the kidney weight data
are difficult to interpret.
There is limited evidence in the trans-l,2-DCE database for effects on the thymus. In a
90-day drinking water study, Barnes et al. (1985) reported decreased relative (as a ratio of body
weight) and absolute thymus weight in mid- (224 mg/kg-day) and high-dose (452 mg/kg-day)
female mice, but not in any of the treated male mice. Hayes et al. (1987) reported no changes in
absolute and relative thymus weight or histopathologic changes in the thymus at doses almost
10-fold higher than the doses used in Barnes et al. (1985). NTP (2002) reported no changes in
absolute and relative thymus weight in rats and mice, except for a statistically significant
increase in absolute (27%) and relative (25%) thymus weight in female mice at the low dose, and
no significant histopathologic lesions.
Inconsistent hematological findings have been associated with trans-l,2-DCE exposure.
In a 90-day drinking water study, Barnes et al. (1985) reported sporadic changes in hematology
parameters (prothrombin time, leukocytes, and polymorphonuclear leukocytes) in mice; changes
in these parameters were not dose-related or consistent across sexes. In a second subchronic
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drinking water study of trans-1,2-DCE, Hayes et al. (1987) reported no treatment-related effects
on hematologic parameters in rats at doses up to approximately 3,000 mg/kg-day.
NTP (2002) reported mild decreases in hematocrit values, hemoglobin concentrations,
and erythrocyte counts at week 14 in male and female rats in all (380-3,245 mg/kg-day) but the
lowest dose groups (190 mg/kg-day). Only decreased RBC counts showed a dose-response and
statistical significance. This effect was demonstrated in male rats with significant decreases at
doses >380 mg/kg-day (p < 0.05), although the maximum decrease in RBC was only 7% in
males and 5% in females at the highest dose (3,210 and 3,245 mg/kg-day for males and females,
respectively). NTP (2002) further observed the following: (1) McCauley et al. (1995) reported a
decrease in the circulating erythroid mass in Sprague-Dawley rats exposed to 872 mg/kg cis-
1,2-DCE by gavage for 90 days, but this response was not dose-related and not considered by the
study investigators to be biologically relevant; and (2) no hematologic response was observed by
Barnes et al. (1985) in male or female CD-I mice exposed to 387 mg/kg-day trans-1,2-DCE for
90 days. NTP (2002) concluded that the trans- (and cis-) isomer may have an effect on
hematologic endpoints but more consistency between studies is necessary before the biological
significance (if any) is known.
The immunotoxicity associated with oral exposure to trans-1,2-DCE was investigated in
mice treated for 14 or 90 days (Shopp et al., 1985; Munson et al., 1982). In male CD-I mice
administered trans-1,2-DCE for 14 consecutive days at doses up to 222 mg/kg-day by gavage,
Munson et al. (1982) evaluated humoral immune function as indicated by the ability of spleen
cells to produce IgM AFCs following challenge with sRBCs. Munson et al. (1982) also assessed
cell-mediated immune function as indicated by the DTH response to sRBCs. The authors
described the antibody response to sRBCs as the number of AFCs per spleen and per 106 spleen
cells. Munson et al. (1982) reported a trend toward suppression of the number of AFCs
expressed on a per spleen basis (significant atp <0.1), but this response was not statistically
significant at the p < 0.05 level or when expressed per 106 spleen cells. The DTH response was
characterized as slight but significant (p < 0.05) and not-dose dependent in the abstract of the
journal article. However, in the results section of the article the authors state that trans-l,2-DCE
showed no effect in the DTH response. This contradictory reporting of the DTH response data
renders these findings questionable. It is unknown whether the slight reduction in DTH was
associated with trans-1,2-DCE or some other test chemical evaluated in the study. The authors
concluded that mice exposed to trans-1,2-DCE for 14 consecutive days at doses up to
222 mg/kg-day showed no significant change in cell-mediated or humoral immunity (Munson et
al., 1982). Data from a longer duration (90-day) oral exposure study demonstrate that
suppression of the antibody response to sRBCs is associated with exposure trans-1,2-DCE, but
do not support an effect of trans-1,2-DCE on the DTH response.
In a 90-day drinking water study in mice (Shopp et al., 1985), a dose-related suppression
of the humoral immune status, as measured by spleen cell antibody production directed against
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sRBCs, was observed in male mice treated with trans-1,2-DCE. When expressed as AFCs per
106 spleen cells, the number of AFCs was reduced by 26% in male mice at doses of 175 and
387 mg/kg-day (significantly different atp < 0.05 from control mice). Shopp et al. (1985)
reported that there was marked suppression in humoral immune status in male mice as indicated
by the significantly decreased number of AFCs in these mice. However, the authors also stated
that the decrease in AFCs was not severe enough to depress the functional ability of the humoral
immune system because there was no change in hemagglutination titers to sRBCs or
lymphoproliferative response of spleen cells to the B-cell mitogen LPS. The authors concluded
that the immune system of CD-1 mice does not appear to be overly sensitive to trans-1,2-DCE
and that the observed effects were probably the result of general toxicity as opposed to specific
target organ toxicity.
EPA evaluated the findings from Shopp et al. (1985) and determined, in contrast to the
study authors, that suppression in the number of AFCs in male CD-I mice represents functional
suppression of the humoral immune system and not general toxicity. No indicators of general
toxicity, such as reduced body weight, were reported.
Loss of immune function has been shown to increase susceptibility to infection. Luster et
al. (1993) identified animal studies where decreases of less than 20% in the AFC (compared to a
26% reduction in AFC in male mice treated with trans-l,2-DCE in the Shopp et al. [1985] study)
are likely to increase susceptibility to infection when the animal is challenged with an infectious
agent. Measurement of the antibody response to a T-dependent antigen, such as sRBCs that are
used as a surrogate for a typical foreign material or infectious agent in toxicology studies,
represents a sensitive and reproducible biomarker of immune function (WHO, 1996). A
functional immune system is required by the host to properly defend against infections, as most
recently evidenced by the numerous fatalities from infectious diseases in individuals with AIDS
or increased deaths from influenza in individuals with compromised immune systems. The exact
quantitative association between loss in immune function (or the AFC response) and
development of infectious disease is difficult to ascertain. Luster et al. (1993) conducted mouse
studies in which groups of mice were administered increasing doses of an immunosuppressive
drug (cyclophosphamide) and increasing amounts of infectious agents (e.g., Listeria). Although
no longer used as an endpoint, survival was monitored as the indicator to resist Listeria infection.
The data were modeled and the results indicated that decreases of even less than 20% in the AFC
response decreases the ability of the host to survive infection.
Luster et al. (2004) summarized effects of moderate losses in immune function on
infectious diseases in humans. A number of studies have examined both infectious disease
incidence and immune function in groups of individuals with moderately compromised immune
systems. For example, studies in immunosuppressed patients following hematopoietic stem cell
transplantation showed a 1.7-fold higher rate of infections with only twofold decreases in certain
types of CD4 T cells (Storek et al., 2000). Similarly, transplant patients, even when on very low
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levels of immunosuppressive therapy, show a 1.5-fold increased risk of in immune function and
associated increased levels of antibodies to latent viruses, such as CMV, EBV and HSV, an
indication of viral reactivation (Kiecolt-Glaser et al, 2002).
EPA's testing guidelines for immunotoxicity require testing the antibody response to a
T-cell-dependent antigen (and suggest the use of sRBCs) as the primary assay to determine
functional responsiveness of major components of the immune system following chemical
exposure in mice and rats (U.S. EPA, 1998b). To evaluate the antibody response, EPA's testing
guidelines require the measurement of splenic anti-sRBC AFCs or serum levels of anti-sRBC
IgM. Reduction in either of these measures of the antibody response represents evidence of
chemical immunosuppression. The AFC assay is a well validated endpoint in immunotoxicology
and has been characterized across multiple labs (Ladies, 2007; Loveless, 2007). While serum
measurements of anti-sRBC IgM levels such as the hemagglutination test are more convenient
than AFC assays because of the ability to obtain values from frozen serum rather than viable
animal cells, the two assays provide an evaluation of different aspects of the antibody response.
The AFC assay provides a measure of the antibody producing cells of the spleen, and this
measure is highly predictive of the overall immunotoxicity of a chemical (Luster et al., 1993,
1992). Serum anti-sRBC IgM values are a general measure of the antibody response because
these values reflect antibodies produced from multiple sources, including spleen, lymph nodes,
and bone marrow. Therefore, the AFC assay is not expected to provide evidence of chemical
immunosuppression at the level of splenic antibody production that might not be identified by
measurements of serum levels of anti-sRBC IgM. Data on the antibody response for two well
known immunotoxicants (cyclophosphamide and dexamethasone) demonstrate that the AFC
assay can be a more sensitive assay for the determination of suppression of the antibody response
than measurement of serum levels of anti-sRBC IgM in an enzyme-linked immunosorbent assay
(Loveless et al., 2007). Loveless et al. (2007) reported that the AFC assay was consistently
better at identifying suppression of the T-dependent antibody response across laboratories and
that the AFC detected suppression at lower concentrations for dexamethasone than were
observed by measurement of serum levels of anti-sRBC IgM. In addition to the negative data
from the hemagglutination assay, Shopp et al. (1985) uses a lack of an observed effect of trans-
1,2-DCE on the proliferative response of splenocytes to LPS to suggest that the functional ability
of the humoral immune system in male mice was not suppressed.
The proliferative response to LPS is not a reliable indicator of humoral immune
suppression, and is listed as one of the poorest predictors for potential immunotoxicity in the
review of sensitivity and predictability of immune tests by Luster et al. (1992). In addition, the
LPS response is a non-specific activation of certain B cells, is a fairly insensitive assay, and is
seldom used today in immunology testing (Luster et al., 1992). In contrast, suppression of T-
cell-dependent antibody response as determined by the AFC assay is one of the most predictive
assays for chemical immunotoxicity (Herzyk and Holsapple, 2007; Luster et al., 1992).
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Therefore, EPA determined that the 26% suppression in the number of sRBC-specific
AFCs per 106 spleen cells of male mice in Shopp et al. (1985) is a biologically significant
measure indicating suppressed immune function associated with oral exposure to trans-1,2-DCE
that is not contradicted by a lack of observed change in the hemagglutination assay to sRBCs or
proliferative response to LPS.
4.6.1.3. Mixtures of cis- and trans-l,2-DCE
There is inadequate information available on the mixtures of cis- and trans-1,2-DCE to
support a separate health assessment; however, effects observed in studies of the mixture of
isomers are generally consistent with those of the individual isomers. Dow (1960) reported that
a dose of 2,000 mg/kg 1,2-DCE (isomer composition not stated) as a 10% solution in corn oil
administered to rats by gavage was not lethal, but some kidney injury was observed at necropsy.
In a 14-day oral gavage study in which a 50% mixture of both 1,2-DCE isomers was
administered in a sesame seed oil vehicle (1 mL/kg) at a dose of 5 mmol (485 mg/kg-day) to
male Sprague-Dawley-derived rats, McMillan (1986) reported a statistically significant increase
in kidney weights, slight but significant reductions in plasma creatinine and BUN, and an
increase in plasma calcium. In a 30-day study by McMillan (1986), the mean relative weight of
the liver (expressed as a ratio to body weight) in the treated group at termination was
significantly greater by 19% than that of control rats. Additionally, the treated rats exhibited
significant reductions in mean relative weight of the lungs (14%), mean values for plasma AST
(25%), and creatinine levels (17%); erythrocyte count, hemoglobin, and hematocrit levels were
also reduced by 6, 5, and 5%, respectively. A series of developmental range-finding studies in
rats and mice (NTP, 1991a, b, c) conducted with a mixture of 1,2-DCE (composition of isomers
unknown) via the oral route found no signs of developmental or maternal toxicity at any of the
initial doses tested (up to 2,918 mg/kg-day), but found maternal toxicity in the form of reduced
maternal body weight and reduced maternal weight gain at higher doses (up to 6,906 mg/kg-
day).
4.6.2. Inhalation
4.6.2.1. cis-l,2-DCE
No studies of the effects of cis-1,2-DCE by inhalation exposure in humans were
identified. There are no inhalation studies of subchronic, chronic, reproductive, or
developmental toxicity of cis-l,2-DCE. Investigation of the inhalation toxicity of cis-l,2-DCE is
limited to an acute 4-hour inhalation LC50 study in rats (DuPont, 1999). The LC50 was calculated
to be 54,200 mg/m3. Effects associated with acute inhalation exposure to cis-l,2-DCE at
concentrations near the LC50 included severe weight loss and clinical signs suggestive of effects
on the CNS, including unresponsiveness, weakness, and irregular respiration immediately after
exposure, and minimal hepatic centrilobular vacuolation upon microscopic observation.
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4.6.2.2. trans-l,2-DCE
The human database for trans-1,2-DCE is limited to one study from the 1930s involving
only two subjects (Lehmann and Schmidt-Kehl, 1936). This study provides limited evidence
that trans-1,2-DCE can cause eye irritation and CNS depression (nausea, drowsiness, fatigue,
vertigo) following acute inhalation exposures. Information on the potential health effects of
inhaled trans-1,2-DCE comes from studies in animals, including two acute inhalation studies
(DuPont, 1999; Freundt et al., 1977) and two subchronic inhalation studies (DuPont, 1998;
Freundt et al., 1977), of which only one (Freundt et al., 1977) is a published peer-reviewed study.
In addition, one study evaluated the effects of inhalation exposure on developmental outcomes
(DuPont, 1988a; published in Hurtt et al., 1993).
Evidence for liver toxicity associated with inhaled trans-1,2-DCE is inconsistent. In the
only published peer-reviewed subchronic inhalation study, Freundt et al. (1977) reported slight to
severe fatty accumulation in the liver lobules and Kupffer cells in rats exposed for 8 hours/day, 5
days/week to air containing 792 mg/m3 trans-1,2-DCE for 1, 2, 8, or 16 weeks. These effects
occurred in two of the six rats exposed for 1 week, in four of the six rats exposed for 2 weeks, in
three of the six rats exposed for 8 weeks, and in five of the six rats exposed for 16 weeks. These
effects were also seen in one of the six controls at 8 weeks and in two of the six controls at 16
weeks. In general, the incidence and severity of fat accumulation increased with increasing
exposure duration. Similar effects were reported in an acute inhalation study by the same
investigators (Freundt et al., 1977).
In the 90-day inhalation toxicity study (DuPont, 1998), rats were exposed to analytically
determined mean concentrations of 0, 792, 3,960, or 15,800 mg/m3 trans-1,2-DCE for 6
hours/day, 5 days/week. Effects on the liver were limited to increases in relative and absolute
liver weights of 8% or less compared to the control. No evidence of fatty accumulation was
observed at any exposure concentration.
Changes in some hematological parameters were reported in the DuPont (1998) study.
Significantly decreased mean hemoglobin concentrations and hematocrit were observed in male
rats and decreased monocyte count in female rats at the 45-day sampling time; similar changes
did not occur at the 90-day sampling time and thus were not considered to be toxicologically
important.
Decreases in WBC and lymphocyte counts were also observed at the 45-day and 90-day
sampling times in male and female rats that were generally concentration related.5 The decreases
were statistically different from the control at 45 and 90 days only in the 15,800 mg/m3 (high-
dose) males. WBC decreased by about 20% in male rats and by approximately 18% in female
5 WBCs (or leukocytes) consist of five difference cell types, including neutrophils, basophils, eosinophils,
lymphocytes, monocytes, and macrophages. In the DuPont (1998) study, the reduced WBC count generally
reflected the reduced lymphocyte count.
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rats, and lymphocyte levels decreased by about 25% in male rats and by about 22% in female
rats. These findings were not considered by the authors to be toxicologically important because
the magnitude of the changes was considered small in the context of historical controls and
because a common cause of decreased lymphocyte counts in rodents is the release of endogenous
glucocorticoids. The endogenous glucocorticoids cause redistribution of lymphocytes from the
circulation into the lymphoid tissue and is a secondary effect associated with stress (Jensen,
1969; Brondeau et al., 1990). The authors of the DuPont (1998) report noted that this type of
stress-related response has been observed in other inhalation studies at that laboratory and
elsewhere (Brondeau et al., 1990). Brondeau et al. (1990) examined the effects of a single 4-
hour exposure to airborne chemicals (but not 1,2-DCE) at irritant levels on blood cell counts in
rats and found that leucopenia was related to the irritant potencies of the test compounds. Since
stimulation of the hypophysis-adrenal axis can account for many of the physiological effects
associated with an extensive variety of stressors (Yannai, 1983), Brondeau et al. (1990) also
examined the effects of exposure to irritant levels in adrenalectomized rats and found that the
leucopenic effect was adrenal dependent. Similarly Shimizy et al. (2000) demonstrated
decreased leukocyte counts after 12 or 24 hours of restraint stress and showed that the
lymphocytopenia induced by restraint stress was absent in adrenalectomized mice. Dhabhar et
al. (1995) suggested that stress-induced increases in plasma corticosterone were accompanied by
significant decreases in numbers and percentages of lymphocytes and that the effects of stress
were largely dependent on adrenal hormones because the magnitude of the stress-induced
changes was significantly reduced in adrenalectomized animals.
In the DuPont (1998) study, the authors failed to identify the cause of stress in exposed
animals. In other studies, trans-1,2-DCE was reported as irritating in humans at a concentration
of 950 ppm (3,772 mg/mg3) (Lehmann and Schmidt-Kehl, 1936) and in rats at a concentration of
2,000 ppm (7,940 mg/mg3) (Hurtt et al., 1993). Therefore, it is plausible that the decrease in
lymphocyte count reflects a stress-related increase in glucocorticoid levels and a secondary effect
on trafficking and redistribution of WBC between the blood and other immune components in
the rat, but direct evidence is not available.
Of note is the fact that no histopathological changes of the spleen and thymus were seen
in rats at any exposure concentration. Statistically significant changes in WBC and lymphocyte
counts were not identified by NTP (2002) in their 90-day oral (feed) study. Therefore, the
decreases in WBC and lymphocyte count, while treatment related, are of uncertain toxicological
significance.
In a single-exposure concentration inhalation study by Freundt et al. (1977),
histopathological changes of the lung (hyperemia and alveolar septal distension) were reported in
rats exposed to 11,880 mg/m3 for 8 hours. Similar effects were reported by these investigators in
animals exposed to 792 mg/m3 trans-l,2-DCE for 8 hours/day, 5 days/week for up to 16 weeks.
The pathological changes in the lung were considered by the authors to be slight in severity and
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were present in all six rats in all four exposure groups, in one of the six control animals exposed
for 1 week, and in two of the six control animals exposed for 2 weeks, but not in any of the
control animals exposed at either 8 weeks or 16 weeks. No lung pathology was observed in the
90-day study by DuPont (1998) at a concentration approximately 20-fold higher than that used
by Freundt et al. (1977). The finding of lung effects in the Freundt et al. (1977) study is difficult
to interpret as this study is the only report of lung pathology in animals exposed to trans-
1,2-DCE, a small number of animals were examined, several of the controls also developed this
effect, and the upper respiratory tract was not examined for pathology.
Only one study investigated the developmental toxicity of trans- 1,2-DCE in pregnant rats
that were exposed via inhalation to 7,930, 23,790, or 47,580 mg/m3 trans-l,2-DCE for
6 hours/day (DuPont, 1988a; published in Hurtt et al., 1993). The two high concentrations were
overtly maternally toxic, while the 7,930 mg/m3 concentration (chosen at 10 times the TLV) was
slightly maternally toxic. There were no changes in numbers of fetuses or implantations, but a
statistically significant decrease in fetal weight was reported at the highest concentration. No
malformations were observed. Oral administration of a 1,2-DCE mixture to pregnant mice and
rats similarly provided no evidence for developmental toxicity and showed maternal toxicity
only at high doses.
Evidence for CNS toxicity following trans-1,2-DCE exposure by the inhalation pathway
comes from studies of acute inhalation exposure only. Freundt et al. (1977) reported no
symptoms of CNS depression in rats that received an 8-hour inhalation exposure to
concentrations up to 11,880 mg/m3. Lethargy and irregular respiration were reported
immediately after exposure to a concentration of 89,100 mg/m3 (DuPont, 1999).
4.6.2.3. Mixtures of cis- and trans-l,2-DCE
There is inadequate information available on the mixtures of cis- and trans-1,2-DCE to
support a separate human health assessment; however, effects observed in studies of the mixture
of isomers are generally consistent with those of the individual isomers. At concentrations of
about 115,270 mg/m3, rats exposed to a mixture of 1,2-DCE isomers (unspecified composition)
rapidly became unconscious, and exposures lasting longer than 0.2 hours were fatal (Dow,
1960). In another Dow study (1962), rats, rabbits, guinea pigs, and beagle dogs were exposed to
0, 1,980, or 3,960 mg/m3 of a 1,2-DCE mixture (58% cis-, 42% trans isomer) 7 hours/day for six
months. The only notable effects in rats and rabbits exposed to the highest concentration (3,960
mg/m3) were an increase in the average relative kidney weights (expressed as a ratio to body
weight) in male and female rats by 16 and 9%, respectively (only statistically significant in
males), and an increase of 23% in the average relative liver weights of female rats (statistically
significant). Liver weights in both male and female rabbits were also increased, but statistical
significance was not determined because of the small number of rabbits tested. At the mid-
concentration of 1,980 mg/m3, the relative kidney weights of male and female rats were
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statistically significantly increased by 9 and 18%, respectively; liver weights of female rats were
also significantly increased by 19%. In addition, increases in liver weights of both male and
female rabbits occurred at termination (statistical evaluations were not performed because of the
small number of experimental animals). Thus, data from studies on the inhalation of mixtures of
1,2-DCE support the conclusion that the liver and kidney may be target organs for 1,2-DCE.
This is consistent with the findings from the oral subchronic studies of cis-1,2-DCE (McCauley,
1995, 1990), the oral subchronic studies of trans-l,2-DCE (NTP, 2002), and the inhalation
subchronic studies of trans-l,2-DCE (Freundt et al., 1977).
4.6.3. Mode-of-Action Information
The available information on the toxic responses to either cis- or trans-1,2-DCE is limited
and precludes the determination of a mode of toxic action. The acute toxicity (and possibly
behavioral toxicity) of both isomers are likely the result of CNS toxicity related to the anesthetic
and narcotic properties of both compounds. The nonspecific effects observed (NTP, 2002;
McCauley et al., 1995; Hayes et al., 1987; Freundt et al., 1977) do not point to any particular
mode of action (e.g., covalent binding of a metabolite). However, elucidating the reaction of
metabolites of cis- and trans-1,2-DCE with cell components and their possible binding to a cell
component may inform a possible mode of action.
In vitro studies indicate that the biotransformation of cis- and trans-1,2-DCE involves the
hepatic CYP450 system. Furthermore, it has been proposed that multiple forms of hepatic
CYP450 bind and metabolize cis- and trans-1,2 DCEs (Costa and Ivanetich, 1982). The study by
Costa and Ivanetich (1982) suggests that the hepatic CYP450 system is closely associated with
the metabolism and toxicity of 1,2-DCEs. In addition, a number of studies indicate that both the
cis- and the trans-isomers are able to induce (at the protein synthesis levels) and/or inhibit (via
suicide inhibition of the enzyme or suppression of protein synthesis) CYP450s (Nakahama et al.,
2000; Hanioka et al., 1998; Mathews et al., 1997; Paolini et al., 1995, 1992; Testai et al., 1982;
Freundt and Macholz, 1978).
Filser et al. (1982, 1978) and Filser and Bolt (1980) observed the production of acetone
following exposure to cis- and trans-1,2-DCE (and other halogenated ethanes). When male
Wistar rats were exposed to cis- or trans-1,2-DCE at various concentrations in a closed-system
chamber, the authors found acetone in the exhaled air (Filser et al., 1978). Chloroacetate, a
known metabolite of haloethanes, also caused acetone exhalation; thus, the study authors
proposed that the effect was caused by inhibition of the citric acid cycle. Similar to the results of
Freundt and Macholz (1978), they found that cis-l,2-DCE was more potent than the trans-
isomer in eliciting acetone production. Subsequently, Filser and Bolt (1980) reported that the
amount of acetone exhaled far exceeded the amount of cis- or trans-1,2-DCE metabolized in the
animals, suggesting that the exhaled acetone was not a metabolite of the test agent. The authors
found that acetone formation did not increase further when cis- or trans-1,2-DCE exposure
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surpassed concentrations that saturated the metabolic capacity of the test animals; however,
induction and inhibition of CYP450 increased and decreased acetone formation, respectively.
They concluded that metabolic transformation of cis- or trans- 1,2-DCE was a prerequisite for
acetone formation. Based on a further study using only trans-1,2-DCE, Filser et al. (1982)
suggested that the exhaled acetone was likely a by-product of increased lipid metabolism (a
ketone body).
In summary, both cis- and trans-1,2-DCE induced synthesis of specific CYP450
isozymes to some extent (e.g., CYP1A1/2 2B1 and CYP2E1 in mice). The trans-l,2-DCE is an
effective but transient inhibitor of CYP2E1 that by itself is metabolized poorly. Downstream
metabolites may affect the citric acid cycle and, secondarily, lipid metabolism, causing
ketogenesis, but, because of the likely low in vivo concentrations of metabolites resulting from
1,2-DCE exposure, it is probably biologically ineffective.
4.7. EVALUATION OF CARCINOGENICITY
4.7.1. Summary of Overall Weight of Evidence
Under the Guidelines for Carcinogen Risk Assessment (U .S. EPA, 2005a), there is
"inadequate information to assess the carcinogenic potential" of cis- and trans-1,2-DCE. This
cancer descriptor is based on the absence of epidemiological studies in humans and lack of
animal studies designed to evaluate the carcinogenic potential of cis- or trans-1,2-DCE.
4.7.2. Synthesis of Human, Animal, and Other Supporting Evidence
No epidemiologic studies evaluating possible long-term health effects of cis-1,2-DCE,
trans-1,2-DCE, or their mixture in humans were identified. The longest duration animal study, a
6-month inhalation study in four species (Dow, 1962), did not evaluate any histology or cancer
endpoints. The 90-day feeding study by NTP (2002) evaluated cancer endpoints but no positive
findings were reported.
Evidence from genotoxicity and mutagenicity studies is inconclusive. For example, cis-
1,2-DCE, trans-1,2-DCE, and their mixture have been mostly nonpositive in bacterial
genotoxicity assays for gene reversion or DNA damage but gave positive results in some
bacterial assays for mitotic recombination or aneuploidy, frequently in the absence of metabolic
activation by S9. Results for chromosomal aberrations or sister chromatid exchanges in
mammalian cells in culture were mixed, providing positive findings in the presence or absence of
metabolic activation. Some in vivo assays gave positive results (host-mediated assay,
chromosomal aberrations) for cis-1,2-DCE only, possibly reflecting the fact that hepatic uptake
of cis-l,2-DCE is higher than that of trans-1,2-DCE.
Both cis- and trans-1,2-DCE are converted into reactive epoxides (oxiranes) by CYP450
enzymes. It is likely that epoxides are responsible for the inactivation of CYP2E1 by binding to
its heme moiety, and protein adduct formation via sulfhydryl groups of amino acids has been
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shown to occur with 1,2-DCE (Maiorino et al., 1982; Sipes and Gandolfi, 1980). However,
DNA adduct formation has not been demonstrated. DNA binding of 1,2-DCE was negative in an
in vitro assay where other chlorinated hydrocarbons gave positive results (Sipes and Gandolfi,
1980).
Positive results have been obtained with cis-l,2-DCE in several genotoxicity assays in
the absence of metabolic activation, suggesting that the C=C double bond positioned next to two
chlorine substituents might be reactive on its own. However, Henschler (1977), in an evaluation
of the mutagenicity of halogenated olefins, pointed out that asymmetric distribution of chlorine
substituents across the C-C bond, such as exists in 1,1 -DCE, was far more likely to give rise to
mutagenic events because the resulting epoxides are unstable, as compared with a symmetric
distribution of the chlorines as exists in both cis- and trans-l,2-DCE. Evidence for other effects
that could potentially lead to tumor formation, such as redox cycling, GSH depletion, or lipid
peroxidation, has not been shown for cis- or trans-1,2-DCE.
The fact that both cis- and trans-1,2-DCE form epoxides and/or radicals as active
metabolites raises the question of whether these intermediates represent structural alerts.
Laurence et al. (1984) performed a computational study of the reactivities of vinyl chloride and
trans- 1,2-DCE by evaluating the bond energies of protonated chlorine or oxygen in the
corresponding chlorooxiranes. Their assessment indicated that the oxirane from trans-l,2-DCE
should form a guanine N7 adduct analogous to the one found after vinyl chloride exposure that is
thought to be the cause of vinyl chloride-related cancer. However, this evaluation also predicted
that the trans-1,2-DCE oxirane would be far more reactive than the one formed by vinyl chloride,
rapidly reacting with other cellular nucleophiles before sufficient quantities could reach critical
targets in the DNA, and thus predicting a lack of carcinogenicity associated with trans-1,2-DCE.
4.8. SUSCEPTIBLE POPULATIONS AND LIFE STAGES
4.8.1. Possible Childhood Susceptibility
No information is available concerning maternal exposure or health effects in developing
humans exposed to cis- and/or trans-l,2-DCE. One animal study (DuPont, 1988a; published in
Hurtt et al., 1993) investigated the potential for trans-1,2-DCE to induce fetotoxicity or
developmental toxicity in pregnant rats exposed to this agent via inhalation at concentrations of
7,930 to 47,580 mg/m3 for 6 hours daily on GDs 7-16. The results were negative; no
malformations were identified, and fetal weight loss was associated only with concentrations that
were overtly maternally toxic. On the basis of this study, trans-1,2-DCE is not expected to cause
fetotoxicity or developmental effects in humans; however, the limited information does not
support an assessment of potential developmental toxicity. No studies were conducted that
would address childhood susceptibility to either cis- or trans-1,2-DCE.
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4.8.2. Possible Gender Differences
Acute toxicity studies in animals provide suggestive evidence that males may be more
sensitive than females to either cis- or trans-1,2-DCE (McCauley et al., 1995; Hayes et al.,
1987). Hanioka et al. (1998) demonstrated that both cis- and trans-l,2-DCE were far more
effective in affecting the activity of CYP450s in male rat hepatic microsomes as compared with
female preparations. However, conclusions about gender differences in response to 1,2-DCE
exposure cannot be drawn based on this limited information.
4.8.3. Other—Genetic Polymorphisms
Four specific enzymes have been associated with the metabolism of cis- or
trans-1,2-DCE: CYP2E1, CYP3A4, ADH, and GSTZ. Of the CYP450s, CYP3A4 is active
toward these compounds in rats (Costa and Ivanetich, 1982) but most likely not in humans
(Guengerich et al., 1991) (for detail, see Sections 3.3.1 and 3.3.2). Alcohol dehydrogenases
represent a whole family of enzymes, several members of which display gene polymorphism.
Because the specific type of ADH, that according to Costa and Ivanetich (1982) and Filser and
Bolt (1980) may be involved in 1,2-DCE metabolism, has not been characterized, possible
variation in susceptibility associated with ADHs is not further considered here. CYP2E1 and
GSTZ, as enzymes whose polymorphisms might affect the susceptibility of humans towards cis-
or trans-1,2-DCE, are discussed below.
4.8.3.1. Cytochrome P450 2E1
CYP2E1 is constitutively expressed in human liver but is inducible by a variety of
factors, prominently by ethanol consumption, diabetes, or hunger, with in vivo activity levels
varying up to 20-fold (Rannug et al., 1995). Variation in the expression of CYP2E1 could
influence susceptibility to the effects of cis- ortrans-l,2-DCE. In addition, at least six allelic
variants of CYP2E1 are known to exist in humans (Bartsch et al., 2000). In Caucasians, so far
no variation in catalytic activity has been associated with genotype; >90% of Caucasians carry
the homozygous wild-type cl/cl allele. Asians, however, also carry the variant c2 allele, and the
homozygous form of that allele, c2/c2, has been shown to have lower catalytic activity than the
wild-type or the cl/c2 heterozygote (Bartsch et al., 2000). The frequency of the c2 allele has
been reported to be 19-24% in Asians, and the frequencies of other variants are also much higher
in Asians than in Caucasians (Rannug et al., 1995). There is evidence that the homozygous
allele c2/c2 and, to a lesser extent, the heterozygote cl/c2 are associated with an increased risk
for several cancer types (Bartsch et al., 2000). Thus, the possibility exists that polymorphism of
the CYP2E1 gene may affect the susceptibility of humans to the effects of cis- and/or trans-
1,2-DCE.
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4.8.3.2. Glutathione S-Transferase
Although DC A is likely a minor metabolite of cis- and trans-1,2-DCE, it is considered a
likely human carcinogen, and therefore genetic polymorphism of the enzyme that metabolizes
DCA, GSTZ, may play a role in human susceptibility. GSTZ is polymorphic in humans; at this
time, five variants have been described that carry combinations of two possible A/G and/or two
possible T/C transitions (U.S. EPA, 2003). The known variants are designated GSTZla-la,
GSTZlb-lb, GSTZlc-lc, GSTZld-ld, and GSTZle-le (Blackburn et al., 2001, 2000; Tzeng et
al., 2000). Blackburn et al. (2000) analyzed blood samples of Caucasians (68 female and 73
male Australians of European descent, ages 16-69) and demonstrated that allele frequencies for
variants la, lb, and lc were 0.09, 0.28, and 0.63, respectively. In the following year, Blackburn
et al. (2001) refined their analysis to comprise all five variants, using 128 Australian subjects of
European descent, and found variant distributions of 0.086, 0.285, 0.473, 0.156, and 0 for
GSTZla-la, GSTZlb-lb, GSTZlc-lc, GSTZld-ld, and GSTZle-le, respectively. Board et al.
(2001) produced recombinant versions of variants GSTZ*A through GSTZ*D—corresponding
to the variant alleles GSTZla-la through GSTZld-ld—and tested their in vitro catalytic
activities toward DCA. GSTZ* A had the highest activity with 1.61 jj.mol/minute/mg protein,
followed by *B and *C each with 0.45, and *D with 0.3 jamol/minute/mg protein. Given the fact
that only 9% of Caucasians carry the high-activity allele, the low-activity allelic variants may
contribute to an increased susceptibility to the effects of DCA and, thus, also of cis- ortrans-
1,2-DCE.
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5. DOSE-RESPONSE ASSESSMENT
5.1. ORAL REFERENCE DOSE
5.1.1. cis-l,2-DCE
5.1.1.1. Choice of Principal Study and Critical Effect—with Rationale and Justification
The effects of oral exposure to cis-l,2-DCE in humans have not been investigated.
McCauley et al. (1995, 1990) is the only published oral toxicity study of cis-l,2-DCE. Male and
female Sprague-Dawley rats were administered 0, 32, 97, 291, or 872 mg/kg-day cis-l,2-DCE by
corn oil gavage for 90 days. Terminal body weights in male rats at the two highest dose groups
were lower than controls by 10-11%, but were not statistically significantly reduced. Relative
liver weight (expressed as a ratio to body weight) was significantly increased in male and female
rats at doses >97 mg/kg-day and relative kidney weight was significantly increased in male rats
at all dose levels. Investigators reported no significant compound-related histopathological
lesions of the liver or kidney. Statistically significant, but marginal, decreases in certain
hematological parameters (primarily hemoglobin and hematocrit) were observed at doses >97
mg/kg-day. As discussed in Section 4.2.1.2.1, some errors and inconsistencies were identified
upon examination of the unpublished (McCauley et al., 1990) and published (McCauley et al.,
1995) versions of the study, principally related to the documentation of administered doses by
the study authors, inconsistencies in reporting of methods, and some transcription or calculation
errors in the unpublished report and published paper. These errors and inconsistencies suggest
issues with the quality of the report writing, but not with the study findings themselves. As the
only repeat-dose study of cis-l,2-DCE toxicity, this study was used as the basis for the oral RfD.
There is an overall increasing trend for both relative liver weight and relative kidney
weight in rats exposed to cis-l,2-DCE. Increases in relative liver weight alone (up to 32 and
30% in high-dose male and female rats, respectively) are difficult to interpret. Liver weight
changes occurred in the absence of compound-related changes in liver histopathology and AST,
and measurements of other clinical chemistry indicators of liver function were not performed as
part of this study. Similarly, increases in relative kidney weight in high-dose male and female
rats (up to 27 and 23%, respectively) occurred in the absence of renal histopathology and BUN
and creatinine levels (indicators of renal dysfunction). As discussed in Section 4.6.1.1, the
biological significance of liver and kidney weight changes in the absence of other
histopathologic and clinical chemistry changes is difficult to interpret. Because these organ
weight changes could represent early indicators of liver and kidney toxicity, increased relative
liver and kidney weight, as reported McCauley et al. (1995, 1990), were considered as candidate
critical effects.
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5.1.1.2. Methods of Analysis, Including Models
Relative liver and kidney weight data are summarized in Table 5-1.
Table 5-1. Relative liver and relative kidney weights of rats exposed to cis-
1,2-DCE by gavage for 90 days
Control
Relative Liver Weight
Dose (mg/kg-day)
32
97
291
872
Males2
2.85 0.26
3.15 0.27
3.28 0.18b
3.34 0.44b
3.75 0.20b
Females2
2.82 0.19
2.91 0.18
3.21 0.22b
3.36 0.18b
3.67 0.27b
Control
Relative Kidney Weight
Dose (mg/kg-day)
32
97
291
872
Males2
0.70 0.06
0.80 0.06b
0.83 0.06b
0.83 0.10b
0.89 0.06b
Females2
0.69 0.06
0.71 0.05
0.82 0.23
0.85 0.21
0.85 0.06
"Values are mean ± SD.
bSignificantly different from control group; p 0.05 by Tukey's multiple comparison test.
Source: McCauley et al. (1995).
Benchmark dose (BMD) modeling methodology (U.S. EPA, 2000b) was used to
determine the PODs for the two candidate critical effects, i.e., increased relative liver weight and
increased relative kidney weight.
All of the models for continuous data in U.S. EPA's BMDS (version 1.4.1 for liver
weight data and version 2.1 for kidney weight data; U.S. EPA, 2008, 2007) were fit to relative
organ weight data in male and female rats from McCauley et al. (1995, 1990). BMDS was used
to calculate PODs for deriving the RfD by estimating the effective dose at a specified level of
response (BMDX) and its 95% lower confidence limit (BMDLx). A 10% change in relative organ
weight compared with the control was selected as the benchmark response (BMR) level for these
endpoints. A BMR of 10% change in relative organ weight was selected by analogy to body
weight, for which a 10% change is generally recognized as a minimally biologically significant
change (U.S. EPA, 2000b). In addition, consistent with EPA BMD guidance (U.S. EPA, 2000b),
a BMR corresponding to a change in the mean response equal to one standard deviation from the
control mean was also used to generate BMDs and BMDLs for comparison purposes.
Relative liver weight
In the female rat, only the Hill model (with the power parameter restricted to be greater
than 1) adequately fit the relative liver weight data (test 4 % p > 0.1). The other two continuous
models fit to these data, the first-degree polynomial and power models, exhibited significant lack
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of fit. Table B-l in Appendix B presents the goodness-of-fit statistics and corresponding BMD
and BMDL estimates for all three continuous models (i.e., first-degree polynomial, power, and
Hill models) fit to these data. The Hill model predicted a BMDio and BMDLio of 80.5 and
42.3 mg/kg-day, respectively. For comparison purposes, this same model was fit to these data
using a BMR corresponding to a change in the mean response equal to one standard deviation
from the control mean, and yielded BMDisd and BMDLisd estimates of 53.2 and 28.8 mg/kg-
day, respectively. In this particular case, one standard deviation from the control mean
represented about a 7% change in relative liver weight.
For the male rat, only the Hill model (with power restricted to be greater than 1)
adequately fit the relative liver weight data (test 4 % p > 0.1). The other two continuous models
fit to these data, the first-degree polynomial and power models, exhibited significant lack of fit.
Table B-2 in Appendix B presents the goodness-of-fit statistics and corresponding BMD and
BMDL estimates for all three continuous models (i.e., first-degree polynomial, power, and Hill
models) fit to these data. Modeling of the variance (i.e., the test 3 statistic in the BMDS output)
was not satisfactory (test 3 *£ p = 0.049), but because the selected BMR is not expressed on a
standard deviation basis, the impact on the POD is minimal. Therefore, the Hill model was
considered to provide an adequate fit to the male rat relative liver weight data. This model
predicted a BMDio and BMDLio of 54.4 and 18.6 mg/kg-day, respectively. For comparison
purposes, this same model was fit to these data using a BMR corresponding to a change in the
mean response equal to one standard deviation from the control mean, and yielded BMDisd and
BMDLisd estimates of 40.4 and 13.0 mg/kg-day, respectively. In this particular case, one
standard deviation from the control mean represented about a 9% change in relative liver weight.
See Appendix B for further details regarding the BMD modeling of male and female rat relative
liver weight data for cis- 1,2-DCE.
The BMDLio estimates for a 10% increase in relative liver weight in male and female rats
were 18.6 and 42.3 mg/kg-day, respectively. The candidate POD for the RfD for cis-l,2-DCE
based on liver weight changes was chosen as 18.6 mg/kg-day, the lower of the male and female
BMDLio values.
Relative kidney weight
BMDS modeling of relative kidney weight data was conducted for the male and female
rat. For the male rat, BMDS modeling of relative kidney weight data showed that only the Hill
model adequately fit the data (test 4 p > 0.1). The other continuous models fit to these data,
the polynomial (linear and degree>2) and power models, exhibited significant lack of fit. Table
B-3 in Appendix B presents the goodness-of-fit statistics and corresponding BMD and BMDL
estimates for all continuous models fit to these data (i.e., linear, polynominal, power, and Hill
models). The Hill model predicted a BMDio and BMDLio of 19.8 and 5.1 mg/kg-day,
respectively. For comparison purposes, this same model was fit to these data using a BMR
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corresponding to a change in the mean response equal to one standard deviation from the control
mean, and yielded BMDisd and BMDLisd estimates of 19.0 and 5.1 mg/kg-day, respectively.
Table B-4 in Appendix B presents the goodness-of-fit statistics and corresponding BMD
and BMDL estimates for all four models fit to female rat kidney weight data (i.e., second-degree
polynominal, first-degree polynomial, power, and Hill models). The best-fitting model was
chosen from those models exhibiting adequate fit by selecting the model with the lowest Akaike
Information Criteria (AIC) value, as well as evaluating how well each model visually fit the data,
especially in the region of the curve near the BMD. The Hill model provided the best fit,
predicting a BMDio and BMDLio of 55.2 and 10.4 mg/kg-day, respectively. For comparison
purposes, this same model was fit to these data using a BMR corresponding to a change in the
mean response equal to one standard deviation from the control mean; BMDS failed to generate
a BMDLisd when the model was fit to the female relative kidney weight data. The candidate
POD based on kidney weight changes was chosen as 5.1 mg/kg-day, the lower of the male and
female BMDLio values.
The BMDLio estimate for a 10% increase in relative liver weight in the male rat was 18.6
mg/kg-day and the BMDLio estimate for 10% increase in relative kidney weight in the male rat
was 5.1 mg/kg-day. Increased relative kidney was selected as the critical effect because it is the
more sensitive of the two endpoints.
5.1.1.3. RfD Derivation—Including Application of Uncertainty Factors
An RfD of 0.002 mg/kg-day for cis-l,2-DCE was derived by applying a composite
uncertainty factor (UF) of 3,000 to the BMDLio of 5.1 mg/kg-day, as follows:
RfD = BMDLi0/UF
= 5.1 mg/kg-day/3,000
= 0.002 mg/kg-day
The composite UF of 3,000 includes factors of 10 to protect susceptible individuals, 10 to
extrapolate from animals to humans, 10 for use of a study of subchronic duration, and 3 to
account for database deficiencies.
• An intraspecies UF (UFh) of 10 was applied to account for potentially sensitive
human subpopulations in the absence of quantitative information on the variability of
response to cis-l,2-DCE in the human population.
• An interspecies UF (UFa) of 10 was applied to account for the variability in
extrapolating from laboratory animals to humans. No information was available to
characterize the toxicokinetic or toxicodynamic differences between experimental
animals and humans for cis-l,2-DCE.
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• An UF of 1 was used to account for extrapolation from a LOAEL to a NOAEL (UFl)
because the current approach is to address this factor as one of the considerations in
selecting a BMR for BMD modeling. In this case, a BMR of a 10% change in
relative kidney weight compared with the control was selected under an assumption
that it represents a minimal biologically significant change.
• An UF of 10 was used to account for extrapolating from a POD for a subchronic
exposure duration to estimate chronic exposure conditions (UFs).
• An UF of 3 was used to account for database deficiencies (UFd). The study used in
this RfD derivation, McCauley et al. (1995, 1990), is the only study of repeat-dose
toxicity available for cis-l,2-DCE. The database for this isomer is missing studies of
reproductive toxicity, including a two-generation reproductive toxicity study, and
developmental toxicity; however, the developmental toxicity potential for
cis-l,2-DCE is informed by a series of range-finding studies of the developmental
toxicity of a mixture of 1,2-DCE isomers (composition of isomers unknown) (NTP,
1991a, b, c). No evidence of developmental toxicity was observed in mice or rats
based on the parameters evaluated in these range-finding studies (gravid uterus
weight, fetal body weight, number of fetuses [live/dead], implantation sites, and
resorptions).
5.1.1.4. Previous Oral Assessment
An oral RfD for cis-l,2-DCE was not previously available on IRIS.
5.1.2. trans-l,2-DCE
5.1.2.1. Choice of Principal Studies and Critical Effects—with Rationale and Justification
The effects of oral exposure to trans-1,2-DCE in humans have not been investigated. No
chronic studies of trans-l,2-DCE in experimental animals are available. There are four
subchronic studies of oral exposure to trans-1,2-DCE (NTP, 2002; Hayes et al., 1987; Barnes et
al., 1985; Shopp et al., 1985). Table 4-13 presents a summary of these studies.
In a 14-week gavage study, NTP (2002) examined the effects of trans-1,2-DCE in both
sexes of F344/N rats and B6C3Fi mice. Doses ranged from 190 to 3,210 mg/kg-day in male
rats; 190 to 3,245 mg/kg-day in female rats; 480 to 8,065 mg/kg-day in male mice; and 450 to
7,925 mg/kg-day in female mice. Both untreated and vehicle controls were used. Rats exhibited
a decrease (approximately 6%) in final mean body weight and body weight gain, a minimal (no
greater than 13%) but transient decrease of ALP activity that was not considered by the authors
to be toxicologically relevant, a significant increase in relative liver weight (up to approximately
10% increase in females), a significant decrease in kidney weight (up to 7% decrease in males),
but no gross or histological lesions. The relative liver weight changes in treated female rats were
statistically significantly increased about 6-10% compared with controls at doses of
>395 mg/kg-day; male rats exhibited slight increases (<6%). Similarly in mice, there were
generally no exposure-related alterations in clinical chemistry parameters and no exposure-
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related deaths. Mice exhibited an approximate 4-7% decrease in final mean body weight and
body weight gain, a significant increase in relative liver weight (9-15% increase at doses
>1,900 mg/kg-day in males; approximately 11% doses at >3,760 mg/kg-day in females), but no
gross or histological lesions.
Barnes et al. (1985) and Hayes et al. (1987) are 90-day drinking water studies. The most
prominent effect observed by Barnes et al. (1985) was a statistically significant increase in serum
ALP levels of 62 and 33% in male mice at the 175 and 387 mg/kg-day doses, respectively.
These increases showed no dose-response relationship, were not found in the female mice, and
were within the normal range for this strain of mouse. A statistically significant increase in
mean absolute liver weights and relative liver weights (expressed as a ratio of liver weight to
body weight) was also demonstrated in male mice at the mid dose in the Barnes et al. (1985)
study; however, absolute liver weights in the low- and high-dose groups were less than those of
the controls. Additionally, absolute thymus weight was reduced by 24% (statistically significant)
at the high dose in females and relative thymus weights were statistically significantly reduced in
the mid- and high-dose females. The only treatment-related effects observed by Hayes et al.
(1987) were small but statistically significant increases in absolute kidney weight (8-9%) in
female rats at doses of 1,257 and 2,809 mg/kg-day.
Immunotoxicity of trans-1,2-DCE in CD-I mice was assessed in a 90-day drinking water
study by Shopp et al. (1985). A dose-related suppression of sRBC-specific AFCs was observed
in the spleens of male mice treated with trans-l,2-DCE. Shopp et al. (1985) reported marked
suppression in humoral immune status in male mice at 175 and 387 mg/kg-day as indicated by
the significantly decreased number of AFCs in these mice (when expressed as AFCs per
106 cells).
The subchronic studies by NTP (2002), Hayes et al. (1987), and Barnes et al. (1985)
provide limited evidence for effects of trans-1,2-DCE on other organs. Although there are some
positive hematological findings associated with trans-1,2-DCE exposure (NTP, 2002; Hayes et
al., 1987; Barnes et al., 1985), changes in these parameters were not dose-related or consistent
across sexes except for decreases in RBC counts (NTP, 2002). Decreases in the RBC count were
small and are not considered biologically significant. Therefore, the available evidence does not
support consideration of changes in hematological parameters as a critical effect for trans-
1,2-DCE. Body weights were dose-dependently reduced in male rats (NTP, 2002) by about 6%.
Such reductions were not observed in other oral studies of trans-1,2-DCE.
Decreased number of AFCs against sRBCs (Shopp et al., 1985), decreased absolute
thymus weight (Barnes et al. (1985), and increased liver weight (NTP, 2002) were considered for
derivation of potential PODs to serve as the basis of the trans-1,2-DCE RfD. The immunological
response reported in Shopp et al. (1985) is regarded as a biologically significant response and
was observed at relatively low doses (>175 mg/kg-day) of trans-1,2-DCE. Absolute thymus
weight was reduced by 24% (statistically significant) at the high dose in females and relative
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thymus weights were statistically significantly reduced in the mid- and high-dose females.
Changes in thymus weight were not consistently observed across oral toxicity studies of trans-
1,2-DCE (see Section 4.6.1.2); however, the reduction in thymus weight observed in Barnes et
al. (1985) is considered consistent with the immunological response reported by Shopp et al.
(1985). A review of the subchronic toxicity studies for trans-1,2-DCE provides support for the
liver as a target of toxicity. Liver weight changes were observed in female rats and male and
female mice exposed to trans-1,2-DCE in NTP (2002). The female rats exposed to trans-
1,2-DCE (190-3,245 mg/kg-day) exhibited statistically significant increases in liver weight at
doses >395 mg/kg-day. Liver weights were increased in both male and female mice exposed to
trans-1,2-DCE (450-8,065 mg/kg-day), although the male mice were more sensitive with
significant increases at doses >1,900 mg/kg-day compared with the increases in females at doses
>3,760 mg/kg-day.
5.1.2.2. Methods of Analysis—Including Models
BMD modeling methodology (U.S. EPA, 2000b) was employed to determine candidate
PODs for the three endpoints selected as candidate critical effects—decreased number of AFCs
against sRBCs (Shopp et al., 1985), decreased absolute thymus weight (Barnes et al., 1985), and
increased liver weight (NTP, 2002). All of the available continuous models in BMDS were fit to
the data for these three endpoints. The models in BMDS were used to calculate PODs for
deriving the RfD by estimating the effective dose at a specified level of response (BMDX), and its
95% lower confidence limit (BMDLx).
AFC response to sRBCs
Immume response data for trans-l,2-DCE (i.e., decreased number of AFCs against
sRBCs) based on Shopp et al. (1985) are summarixed in Table 5-2.
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Table 5-2. Humoral immune response to sRBCs in CD-I mice exposed to
trans-l,2-DCE in drinking water for 90 days (day 4)
Exposure group
Spleen weight (nig)
AFCs per spleen ( 10 5)
AFCs per 106 cells
Malesa
Control
202 30
4.48 0.32
2,200 125
0.1 mg/mL (17 mg/kg-day)
164 13
3.28 0.28b
2,048 152
l.Omg/mL (175 mk/kg-day)
178 6
3.34 0.39b
1,625 136b
2.0 mg/mL (387 mk/kg-day)
173 10
2.87 0.37b
1,618 226b
Females'1
Control
228 13
4.38 0.37
1,765 110
0.1 mg/mL (23 mk/kg-day)
176 llb
2.97 0.49b
1,478 211
1.0 mg/mL (224 mk/kg-day)
230 12
4.51 0.24
1,967 89
2.0 mg/mL (452 mk/kg-day)
191 13b
3.47 0.50
1,518 184
"Values are mean ± SE for 12 mice in the control group and 8 mice in treatment groups, measured on day 4 after
antigen presentation.
bValues differ significantly from control group, p < 0.05.
Source: Shopp et al. (1985).
Little information exists concerning the biological significance of particular changes in
AFC levels in rodents, and what these changes would correspond to in humans. Therefore, as
recommended for continuous data in the Benchmark Dose Technical Guidance Document (U.S.
EPA, 2000b), a change in the mean response equal to one standard deviation from the control
mean was used as the BMR to facilitate a consistent basis of comparison across assessments for
this endpoint in the absence of information regarding the level of change considered to be
biologically significant. In this case, a BMR of 1 standard deviation corresponds to a 20%
decrease in AFCs per 106 spleen cells.
Table B-5 in Appendix B presents the goodness-of-fit statistics and corresponding
BMDisd and BMDLisd estimates for all four models fit to these data (i.e., second-degree
polynominal, first-degree polynomial, power, and Hill models). The best-fitting model was
chosen from those models exhibiting adequate fit by selecting the model with the lowest AIC
value, as well as evaluating how well each model visually fit the data, especially in the region of
the curve near the BMD (see Appendix B for further details). Based on these model selection
criteria, a second-degree polynomial model provided the best fit to these data, yielding a
BMDisd of 125.6 mg/kg-day and a BMDLisd of 65.0 mg/kg-day. The BMDLisd of 65.0 mg/kg-
day was identified as a candidate POD for the trans-1,2-DCE RfD.
Absolute thymus weight
Thymus weight data for female mice from Barnes et al. (1985) are summarized in
Table 5-3. No treatment-related effects on thymus weight were observed in male mice.
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Table 5-3. Relative and absolute thymus weights in female mice exposed to
trans-l,2-DCE in the drinking water for 90 days3
Dose (mg/kg-day)
Parameter
Vehicle
23
224
452
Thymus weight (mg)b
71 ±3
67 ±4
61 ±4
54 ± 4C (24%)
[% body weight]
[0.22]
[0.20]
[0.18C] (18%)
[0.1 T] (23%)
aTwenty-three animals/sex in the control group and 15-16 animals/sex in the treatment groups.
bValues presented are mean± SE.
cDiffers statistically significantly from controls, p < 0.05; Duncan's multiple range test was used to determine
statistical significance.
Source: Barnes et al. (1985).
All of the models for continuous data in U.S. EPA's BMDS (version 2.13) (U.S. EPA,
2007) were fit to the absolute thymus weight data from Barnes et al. (1985) in female mice. A
10% change in absolute thymus weight compared with the control was selected as the BMR for
this endpoint by analogy to body weight, for which a 10% change is generally recognized as a
minimally biologically significant change (U.S. EPA, 2000c).
Table B-6 in Appendix B presents the goodness-of-fit statistics and corresponding BMD
and BMDL estimates for all continuous models fit to these data (i.e., linear polynominal, second-
degree polynomial, power, and Hill models). The linear polynominal model provided the best fit
of the data based on the lowest AIC value and predicted a BMDio and BMDLio of 196.1 and
138.5 mg/kg-day, respectively. Consistent with EPA guidance (U.S. EPA, 2000b), for
comparison purposes, this same model was fit to these data using a BMR corresponding to a
change in the mean response equal to one standard deviation from the control mean, and yielded
BMDisd and BMDLisd estimates of 427.7 and 289.04 mg/kg-day, respectively. The BMDLio of
138.5 mg/kg-day based on a 10% decrease in absolute thymus weight in female mice was
identified as a candidate POD for the trans-1,2-DCE RfD.
Relative liver weight
Relative liver weights for male and female mice and rats from NTP et al. (1985) are
summarized in Table 5-4.
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Table 5-4. Relative liver weights in male and female mice and rats exposed
to trans-l,2-DCE in the feed for 14 weeks
Mice - males
Dose (mg/kg-day)
0
480
920
1,900
3,850
8,065
Relative liver
weights
(mean ± SE)a
4.347 ±0.056
4.552 ± 0.113
4.597 ±0.115
4.745 ±0.084b
4.736 ±0.079b
4.979 ± 0.1 llb
Mice - females
Dose (mg/kg-day)
0
450
915
1,830
3,760
7,925
Relative liver
weights
(mean ± SE)a
4.621 ±0.07
4.738 ±0.068
4.970 ±0.127
4.813 ±0.05
5.115 ± 0.139b
5.117 ± 0.08b
Rats - males
Dose (mg/kg-day)
0
190
380
770
1,540
3,210
Relative liver
weights
(mean ± SE)a
3.465 ±0.058
3.538 ±0.032
3.658 ±0.099
3.524 ±0.050
3.492 ±0.048
3.634 ±0.056
Rats - females
Dose (mg/kg-day)
0
190
395
780
1,580
3,245
Relative liver
weights
(mean ± SE)a
2.937 ±0.038
3.040 ±0.052
3.220 ± 0.066 b
3.100 ± 0.051 b
3.132 ±0.052b
3.216 ± 0.051b
"Ten animals per group.
bStatistically significant, p < 0.01.
Source: NTP (2002).
All of the models for continuous data in U.S. EPA's BMDS (version 1.4.1) (U.S. EPA,
2007) were fit to the relative liver weight data. BMDS was used to calculate PODs for deriving
the RfD by estimating the effective dose at a specified level of response (BMDX) and its 95%
lower confidence limit (BMDLx). A 10% change in relative liver weight compared with the
control was selected as the BMR for this endpoint. A BMR of 10% change in relative liver
weight was selected by analogy to body weight, for which a 10% change is generally recognized
as a minimally biologically significant change (U.S. EPA, 2000c).
Only the male mouse relative liver weight data could be adequately modeled by the
continuous models in BMDS. The Hill model (with the power parameter restricted to be greater
than 1) and two other continuous models, the first-degree polynomial and power models, did not
exhibit significant lack of fit (based on /?-value >0.1). The Hill model exhibited the best fit of
the data based on the lowest AIC value. Table B-7 in Appendix B presents the goodness-of-fit
statistics and corresponding BMD and BMDL estimates for all three continuous models fit to
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these data (i.e., first-degree polynomial, power, and Hill models). The Hill model predicted a
BMDio and BMDLio of 3,241.9 and 867.3 mg/kg-day, respectively. Consistent with EPA
guidance (U.S. EPA, 2000b), for comparison purposes, this same model was fit to these data
using a BMR corresponding to a change in the mean response equal to one standard deviation
from the control mean, and yielded BMDisd and BMDLisd estimates of 1,348.7 and 395.9
mg/kg-day, respectively. In this case, one standard deviation from the control mean represented
about a 4% change in relative liver weight. See Appendix B for further details regarding the
BMD modeling of male mouse relative liver weight data for trans-1,2-DCE. The BMDLio of
867.3 mg/kg-day based on a 10% increase in relative liver weight in male mice was identified as
a candidate POD for the trans-1,2-DCE RfD.
Increased relative liver weight was observed in female mice and rats (Table 5-4),
although these data sets were not amenable to BMD modeling because at least one of the mid-
level dose groups exhibited a decrease in relative liver weight yielding a non-monotonically
increasing dose-response function (see Tables B-8 and B-9 in Appendix B for further model
details). Therefore, a NOAEL/LOAEL approach was applied. In female mice, the NOAEL was
1,830 mg/kg-day and the LOAEL was 3,760 mg/kg-day, based on statistically significant
increases in relative liver weight. In female rats, the NOAEL was 190 mg/kg-day and the
LOAEL was 395 mg/kg-day, based on statistically significant increases in relative liver weight.
The NOAEL of 190 mg/kg-day was identified as a candidate POD for the trans-1,2-DCE RfD.
The dose-response analysis of the immune, thymus, and liver endpoints, with PODs of
65.0, 138.5, and 190 mg/kg-day, respectively, suggests that the immune system is more sensitive
to the effects of trans-l,2-DCE. Suppression of the humoral immune system, as measured by
spleen cell antibody production directed against sRBCs, was selected as the critical effect for the
trans-l,2-DCE RfD, and the Shopp et al. (1985) study was identified as the principal study. The
BMDLisd of 65.0 mg/kg-day was selected as the POD for deriving the RfD for trans-1,2-DCE.
5.1.2.3. RfD Derivation—Including Application of Uncertainty Factors
To derive an RfD for trans-1,2-DCE, the BMDLisoof 65.0 mg/kg-day (Shopp et al.,
1985) was divided by a composite UF of 3,000. Therefore, the RfD for trans-l,2-DCE is
calculated as follows:
RfD = BMDLisd-UF
= 65.0 mg/kg-day3,000
= 0.02 mg/kg-day
The composite UF of 3,000 includes factors of 10 to protect sensitive individuals, 10 to
extrapolate from animals to humans, 10 for use of a study of subchronic duration, and 3 to
account for database deficiencies.
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• An intraspecies UF (UFh) of 10 was applied to account for potentially sensitive
human subpopulations in the absence of quantitative information on the variability of
response to trans-1,2-DCE in the human population.
• An interspecies uncertainty factor of 10 (UFA) was applied to account for variability
in extrapolating from laboratory animals to humans. No information was available to
characterize the toxicokinetic or toxicodynamic differences between experimental
animals and humans for trans-1,2-DCE.
• An UF of 1 was used to account for extrapolation from a LOAEL to a NOAEL (UFl)
because the current approach is to address this factor as one of the considerations in
selecting a BMR for BMD modeling. In this case, a BMR of 1 standard deviation in
spleen cell antibody production was selected under an assumption that it represents a
minimal biologically significant change.
• An UF of 10 was used to account for extrapolating from a POD for a subchronic
exposure duration to estimate chronic exposure conditions (UFs).
• An UF of 3 was used to account for database deficiencies (UFd). There are several
subchronic oral studies of trans-1,2-DCE (NTP, 2002; Hayes, 1987; Barnes, 1985;
Shopp, 1985). One study investigated developmental toxicity of trans-1,2-DCE via
inhalation (DuPont, 1988a) and showed few developmental parameters to be affected
by treatment. In this study developmental toxicity was manifest only in high-dose
groups. Developmental toxicity potential for trans-1,2-DCE is also informed by a
series of oral range-finding studies of the developmental toxicity of a mixture of
1,2-DCE isomers (composition of isomers unknown) (NTP, 1991a, b, c). No
evidence of developmental toxicity was observed in mice or rats based on the
parameters evaluated in these range-finding studies (gravid uterus weight, fetal body
weight, and number of fetuses [live/dead], implantation sites, and resorptions). The
database for trans-1,2-DCE is missing studies of reproductive toxicity, including a
two-generation reproductive toxicity study.
5.1.2.4. Previous Oral Assessment
The previous RfD of 0.02 mg/kg-day for trans-1,2-DCE was based on the 90-day
subchronic drinking water study in mice (Barnes et al., 1985). The critical effect was increased
serum ALP in male mice. The LOAEL-NOAEL approach was used to derive the RfD. A POD
of 17 mg/kg-day (NOAEL) was identified and a combined UF of 1,000 was applied, resulting in
an RfD of 0.02 mg/kg-day. The UF of 1,000 accounted for the uncertainty in the extrapolation
of dose levels from laboratory animals to humans (UFA= 10), uncertainty in the threshold for
sensitive humans (UFh = 10), and uncertainty in extrapolating from subchronic to chronic
exposure (UFs = 10), but did not account for database deficiencies.
The current assessment uses a different principal study and a different approach for the
derivation of the RfD from the previous oral assessment. The Shopp et al. (1985) study was
selected as the principal study. A decrease in spleen cell antibody production directed against
sRBCs was identified as the critical effect. BMD modeling was used to analyze the data from
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the Shopp et al. (1985) study rather than a LOAEL-NOAEL approach as in the previous
assessment. The composite UF of 3,000 includes an UF of 3 for deficiencies in the database that
was not included in the derivation of the previous RfD.
5.2. INHALATION REFERENCE CONCENTRATION
5.2.1. cis-l,2-DCE
There are no human, chronic, or subchronic inhalation studies for cis-l,2-DCE. The
inhalation toxicity database for cis-l,2-DCE is limited to an acute study (DuPont, 1999) in male
and female Crl:CD®BR rats from which an LC50 of 54,200 mg/m3 was calculated. Therefore, in
the absence of repeat-dose toxicity studies, the available inhalation data for cis-l,2-DCE do not
support derivation of an RfC.
An inhalation assessment for cis-l,2-DCE was not previously developed for the IRIS
database.
5.2.2. trans-l,2-DCE
No epidemiological studies of the effects of inhalation exposure to trans- 1,2-DCE in
humans are available, and case reports involving acute exposure to 1,2-DCE do not provide data
useful for derivation of an RfC. There are two 90-day or longer duration studies using trans-
1,2-DCE (DuPont, 1998; Freundt et al., 1977). The Freundt et al. (1977) subchronic study is a
single dose study with liver endpoint data collected over several exposure durations and the
DuPont (1998) report is available only as an unpublished study.
Freundt et al. (1977) exposed six rats/group for 8 hours/day, 5 days/week to air
containing 792 mg/m3 (200 ppm) trans-l,2-DCE for 1, 2, 8, and 16 weeks. As shown in Table 4-
7, histological changes included slight to severe fatty accumulation in the liver lobules and
Kupffer cells after exposure to 792 mg/m3 for 1, 2, 8, and 16 weeks. For each of the exposure
durations there was no statistically significant difference between the controls and the exposed
groups with respect to the incidence of liver effects (fat accumulation). In general, however, the
incidence and severity of fat accumulation increased with increasing exposure duration.
In the DuPont (1998) study, male and female rats (15/sex/dose) were exposed to 0, 792,
3,960, or 15,800 mg/m3 (0, 200, 1,000, or 4000 ppm) trans-l,2-DCE for 6 hours/day,
5 days/week for 90 days. Changes in relative and absolute liver and kidney weight were not
statistically significant compared to controls. No exposure-related effects were seen in clinical
or pathology parameters or on liver cell proliferation. The only hematological changes that
showed dose-related trends at both 45 and 90 days in male and female rats were changes in WBC
and lymphocyte counts. WBC decreased by up to 18 to 20% in male and female rats, and
lymphocyte levels decreased by up to 22 to 25%.
Although Freundt et al. (1977) reported histopathologic changes in the liver of rats, the
DuPont (1998) study did not corroborate the Freundt et al. study findings. DuPont (1998)
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reported relatively small increases in relative and absolute liver weight (1 to 8%) and no gross or
microscopic changes of the liver attributable to trans-1,2-DCE at an exposure concentration 20-
fold higher than that used in the Freundt et al. (1977) study. NTP (2002) similarly found no
histopathologic changes in the liver when trans-1,2-DCE was administered for 90 days by the
oral route at dietary concentrations as high as 50,000 ppm. In light of the results of DuPont
(1998) and NTP (2002), it is difficult to explain the liver findings in the single-exposure
concentration study by Freundt et al. (1977). Given the limitations of the Freundt et al. (1977)
study (i.e., small sample size, use of only one exposure concentration, and observation of fatty
accumulation in the liver lobules and Kupfer cells in control animals at some exposure durations)
and lack of corroboration from other studies, the Freundt et al. (1977) study was not used as the
basis for deriving an RfC for trans-l,2-DCE.
The findings from the DuPont (1998) study were also considered as the basis for RfC
derivation. As noted above, increases in relative and absolute liver and kidney weight ranged
from 1 to 8% but were not statistically significant. As discussed in Section 4.6.2.2, the decreases
in WBC and lymphocyte count reported in DuPont (1998), while treatment related, are of
uncertain toxicological significance. The study authors suggested that the decreases in WBC and
lymphocyte counts were attributable to the release of endogenous glucocorticoids that can cause
redistribution of lymphocytes from the circulation into the lymphoid tissue and may, therefore,
be considered a secondary effect associated with stress (Jensen, 1969; Brondeau et al., 1990).
While plausible, specific support for this hypothesis was not provided. The lack of
histopathological changes of the spleen and thymus in the DuPont (1998) study are not consistent
with a direct effect of trans-l,2-DCE on lymphocytes. Further, the hematological findings from
other trans-1,2-DCE toxicity studies do not support a determination that trans-1,2-DCE induces
toxicologically significant effects on these hematologic parameters. In a 90-day drinking water
study, Barnes et al. (1985) reported sporadic changes in hematology parameters (prothrombin
time, leukocytes, and polymorphonuclear leukocytes) in mice; changes in these parameters were
not dose-related or consistent across sexes. In a second subchronic drinking water study of trans-
1,2-DCE, Hayes et al. (1987) reported no treatment-related effects on hematologic parameters in
rats at doses up to approximately 3,000 mg/kg-day. The NTP (2002) 90-day oral study found
slight changes in WBC and lymphocyte counts in male and female rats that were not statistically
significantly different from the control and generally showed poor dose-response relationships.
In summary, the available inhalation data from Freundt et al. (1977) and DuPont (1998) were
considered insufficient to support reference value derivation and, therefore, an RfC for trans-1,2-
DCE was not derived.
No inhalation assessment for trans-1,2-DCE was previously included in the IRIS
database.
5.3. UNCERTAINTIES IN THE ORAL REFERENCE DOSE (RfD)
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Risk assessments need to describe associated uncertainty. The following discussion
identifies uncertainties associated with the RfDs for cis- and trans-l,2-DCE. RfC values were
not derived for cis- or trans-1,2-DCE in this assessment. As presented earlier in this section, the
uncertainty factor approach, following EPA practices and RfD and RfC guidance (U.S. EPA,
2002b; 1994b) was applied to POD (BMDLio) values for the cis- and trans-1,2-DCE RfDs.
Using this approach, the POD was divided by a set of factors to account for uncertainties
associated with a number of steps in the analysis, including extrapolation from responses
observed in animal bioassays to humans, extrapolation of data from subchronic exposure to
chronic exposure, to account for a diverse population of varying susceptibilities, and to account
for database deficiencies. Because information specific to cis- and trans-l,2-DCE was limited,
default factors were generally applied for these extrapolations.
The human database for 1,2-DCE is limited to two early studies (from the 1930s)
involving acute inhalation exposure; one of the two was a case report of 1,2-DCE (of unknown
isomeric composition) and the other a human subj ect study of trans-1,2-DCE with only two
subjects. The animal database available to assess cis-l,2-DCE hazard is limited, consisting of
limited acute oral and inhalation studies and a 14-day and 90-day toxicity study. The database
for the trans-isomer, which includes multiple studies of acute and subchronic toxicity,
developmental toxicity, and immunotoxicity, is more extensive (see Section 4). Uncertainties
associated with gaps in the databases for both 1,2-DCE isomers and uncertainties associated with
the data sets used to derive the RfDs for both 1,2-DCE isomers are more fully discussed below.
Selection of the critical effect for reference value determination. The selection of the critical
effect is a source of uncertainty for the oral RfD for both cis- and trans-1,2-DCE. For cis-
1,2-DCE, kidney effects were noted as the critical effect. Increases in relative kidney weight up
to 27% in high-dose male rats and up to 23% in female rats occurred in the absence of renal
histopathology, and BUN and creatinine levels did not indicate renal dysfunction (McCauley et
al., 1995, 1990). The biological significance of kidney weight changes in the absence of other
histopathologic and clinical chemistry changes is difficult to interpret. Such increases in relative
kidney weight could represent an early indicator of kidney toxicity. The absence of supporting
evidence for kidney toxicity makes interpretation of the kidney weight findings difficult and the
biological relevance of increased kidney weight uncertain.
The critical effect for the RfD for trans-1,2-DCE is based on decreased antibody
production directed against sRBCs in male mice (Shopp et al., 1985). The AFC response
exhibited a dose response. EPA determined that the 26% suppression in the number of sRBC-
specific AFCs per 106 spleen cells of male mice in Shopp et al. (1985) is a biologically
significant measure indicating suppressed antibody response associated with oral exposure to
trans-1,2-DCE that is not contradicted by a lack of observed change in the hemagglutination
assay to sRBCs or proliferative response to LPS. Suppression of T-cell-dependent antibody
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response as determined by the AFC assay to sRBCs is a well validated endpoint that is highly
predictive for immunotoxicity (Herzyk and Holsapple, 2007; Luster et al., 1992). Support for
this critical effect can be found in the decreased thymus weight in the Barnes et al. (1985) study.
Decreased thymus weight can be a good indicator of immunotoxicity and, when accompanied by
decreased AFC response in the absence of general toxicity, serves as a predictor of
immunotoxicity (Luster et al., 1992). In the case of trans-l,2-DCE, it should be noted, however,
that decreased thymus weight was observed in female mice, whereas decreased AFC response
was observed in male mice. Confidence in the critical effect for the trans-1,2-DCE RfD would
be increased if the positive immune response reported by Shopp et al. (1985) was corroborated
by similar findings in a second study.
Dose-response modeling. BMD modeling was used to estimate the POD for the cis- and
trans-1,2-DCE RfDs. BMD modeling has advantages over a POD based on a NOAEL/LOAEL
approach because the latter is a reflection of the particular doses (and dose spacing) selected in
the principal study. The NOAEL/LOAEL approach lacks characterization of the dose-response
curve and for this reason is less informative than a POD obtained from BMD modeling. The
selected models used to derive the cis- and trans-l,2-DCE PODs provided the best mathematical
fits to the experimental data sets, but do not represent all possible models one might fit. Other
models could be selected to yield more extreme results, both higher and lower than those used to
derive the cis- and trans-isomer RfDs in the current assessment.
Animal to human extrapolation. Extrapolating dose-response data from animals to humans is
another source of uncertainty. The effect and magnitude at the POD in rodents are extrapolated
to human response. Uncertainty in interspecies extrapolation can be separated into two general
areas—toxicokinetic and toxicodynamic. In the absence of information to quantitatively assess
either toxicokinetic or toxicodynamic differences between animals and humans, a 10-fold UF
was used to account for uncertainty in extrapolating from laboratory animals to humans in the
derivation of the RfDs for cis- and trans-1,2-DCE. Toxicokinetic and toxicodynamic
information for the isomers of 1,2-DCE is not available to inform the potential magnitude of
over- or underestimation of this UF. A PBPK model adequately parameterized for both animals
and humans could reduce uncertainty in the pharmacokinetic portion of interspecies
extrapolation; however, such a model is not available for cis- or trans-1,2-DCE.
Intrahuman variability. Heterogeneity among humans is another source of uncertainty. In the
absence of cis- and trans- 1,2-DCE-specific data on human variation in response to exposure to
these isomers, a default UF of 10 was used to account for uncertainty associated with human
variation in the derivation of the RfDs. Human variations may be larger or smaller; however,
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1,2-DCE-specific data to examine the potential magnitude of over- or underestimation is
unavailable.
Subchronic to chronic exposure extrapolation. Because no chronic toxicity studies for the cis-
or trans-isomers of 1,2-DCE are available, a factor was applied to extrapolate data obtained from
studies of subchronic exposure to chronic exposure. This factor is based on the assumption that
an effect seen at a shorter duration will also be seen after a lifetime of exposure, but with greater
severity or at a lower exposure level. In the absence of information to inform this extrapolation,
a default UF of 10 was applied. The magnitude of uncertainty associated with this extrapolation
and UF cannot be quantified.
Data gaps. The cis- and trans-1,2-DCE database lacks a multigenerational study of reproductive
toxicity by any route of exposure, and the cis-1,2-DCE database lacks studies of developmental
toxicity. The absence of these studies introduces uncertainty in the RfDs. Uncertainty resulting
from gaps in developmental toxicity data specific to the cis- and trans-1,2-DCE isomers was
reduced by developmental toxicity studies of mixed 1,2-DCE isomers. Additionally,
histopathology data from subchronic studies have shown that organs of the reproductive system
are unlikely targets for 1,2-DCE toxicity. The magnitude of the uncertainty associated with
database deficiencies for these chemicals cannot be quantified. However, a database uncertainty
factor of 3 was used to account for the lack of reproductive and developmental toxicity studies.
5.4. CANCER ASSESSMENT
Epidemiologic studies of 1,2-DCE are not available, and chronic bioassays in
experimental animals have not been performed. Thus, the toxicological databases for both
isomers provide "inadequate information to assess their carcinogenic potential."
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6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HAZARD
AND DOSE RESPONSE
6.1. HUMAN HAZARD POTENTIAL
1,2-DCE exists as two isomers, cis- and the trans-forms, with a molecular mass of 96.95.
Both are colorless, flammable liquids that are heavier than water, with a chloroform-like, sweet,
pungent smell. With boiling points between 48 and 60°C, they are volatile. At approximately 5
g/L, both are moderately water soluble. Their oil:water partition coefficients, at around 100,
suggest that these chemicals will preferentially partition into lipophilic media. The two isomers
may be used in their pure forms or as a mixture of varying isomer composition, typically a 60:40
cis-/trans-mixture.
The trans-isomer is the most commonly used form of 1,2-DCE, and currently is the only
isomer commercially available in the United States. DCE was used historically as a solvent for
polymers and rubber; these uses are no longer in practice. Currently, trans-l,2-DCE is used as
an effective degreasing agent and as a component of formulated products used for precision
cleaning of electronic components. It can also used as a blowing agent for speciality foams.
Little information is available regarding the potential toxicity of cis- or trans-1,2-DCE in
humans by either the oral or the inhalation route of exposure. Acute effects described for inhaled
trans-1,2-DCE in humans include eye irritation, drowsiness, nausea, vertigo, narcosis, and death.
No long-term effects are known. There are no chronic exposure studies in animals. Several
subchronic oral exposure studies in animals have been conducted, including a 90-day gavage
study of the cis-isomer in rats (McCauley et al., 1995, 1990), 90-day drinking water studies of
the trans-isomer in rats (Hayes et al., 1987) and mice (Barnes et al., 1985), and a 90-day feed
study in rats and mice (NTP, 2002). Studies of inhalation exposure consist of two 90-day studies
of the trans-isomer (DuPont, 1998; Freundt et al., 1977) and one study of a mixture of cis- and
trans-isomers in rats (Dow, 1962). Changes in liver and kidney weight were the most frequently
observed effects following exposure to 1,2-DCE; however, there is limited evidence for any
specific pathological event (NTP, 2002; DuPont, 1998; McCauley et al., 1995, 1990; Hayes et
al., 1987; Freundt et al., 1977).
Only one subchronic oral study (McCauley et al. 1995, 1990) was conducted with
cis-l,2-DCE. Statistically significant increases in liver weight in both male and female rats,
increases in kidney weight in male rats, and inconsistent hematological responses were noted in
this study (McCauley et al., 1995, 1990). No histopathologic changes of the liver and kidney
were observed, and clinical chemistry findings were of questionable biological significance.
The subchronic oral toxicity of trans-1,2-DCE has been investigated by NTP (2002),
Hayes et al. (1987), Barnes et al. (1985), and Shopp et al. (1985). Hayes et al. (1987) found no
significant differences in body weight or body weight gain in either male or female rats or effects
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on any of the hematological, serological, or urinary parameters evaluated. In addition, there
were no significant changes in organ weights or relative organ weights in males, and only a
significant elevation in absolute kidney weights and kidney weights relative to brain weights in
females, with no evidence of microscopic histopathological changes in the female kidney. In the
Barnes et al. (1985) study, changes in relative organ weights were few, sporadic, and not
believed by the authors to be treatment-related. Few changes in hematological parameters were
seen in this study, and slight changes in several clinical chemistry parameters were observed.
Although some values were significantly different from those of the controls, there were no
consistent trends or deviations from historical control values. A statistically significant increase
in relative liver weight at the mid-dose (175 mg/kg-day), but not at the highest dose, was seen in
male mice in the Barnes et al. (1985) study, and statistically significant changes in liver function
enzymes, including LDH, AST (SGOT), and ALP activities, in male mice occurred. Significant
increases of 62 and 33% in serum ALP levels were reported at the 175 and 387 mg/kg-day doses,
respectively. These increases showed no dose-response relationship, were within the normal
range for the CD-I mouse strain, and were not observed in female mice. The findings of Barnes
et al. (1985) provide no evidence that trans-l,2-DCE induces hepatotoxicity in mice at doses up
to 387 mg/kg-day in males and 452 mg/kg-day in females.
In NTP (2002), the final mean body weight and body weight gain of male rats exposed to
trans- 1,2-DCE were reduced by about 6% below controls (a statistically significant decrease). In
general, no exposure-related alterations in clinical chemistry parameters in rats were observed.
NTP (2002) reported statistically significant changes in absolute and relative liver weights in rats
for the females only. The relative liver weights of female rats exposed to >395 mg/kg-day were
significantly higher (—6-10%) than the control. No gross or histological lesions were observed
in rats that were attributed to exposure to trans-l,2-DCE. In mice, statistically significant, dose-
dependent increases in relative liver weights in both sexes were observed in the NTP (2002)
study. The maximum changes in liver weights in mice were increases of 15 and 11% at the
highest dose for males and females, respectively. No gross or histological lesions were observed
in mice. The liver weight changes may be early indicators or precursors of liver toxicity, and it
is not possible to determine whether overt liver damage would occur at higher doses or in studies
of longer exposure duration (i.e., chronic studies).
Shopp et al. (1985) reported a dose-related suppression of the humoral immune status in
male mice treated with trans-1,2-DCE as indicated by a reduction in sRBC-specific AFCs in the
spleen. When expressed as AFCs per 106 spleen cells, the number of AFCs were reduced by
26% in male mice at doses of 175 and 387 mg/kg-day (significantly different atp< 0.05 from
control mice). Suppression of T-cell-dependent antibody response as determined by the AFC
response to sRBCs is a well validated endpoint that is one of the most predictive assays for
chemical immunotoxicity (Herzyk and Holsapple, 2007; Luster et al., 1992). EPA concluded
that the reduced number of sRBC-specific AFCs per 106 spleen cells of male mice in Shopp et al.
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(1985) is a biologically significant measure indicating suppressed antibody response associated
with oral exposure to trans-1,2-DCE.
For the inhalation route of exposure, there are no human, chronic, or subchronic studies
for the cis-isomer and three 90-day or longer duration studies using trans-1,2-DCE (DuPont,
1998; Freundt et al., 1977) or the isomer mixture (Dow, 1962). The unpublished DuPont (1998)
report in male and female rats demonstrated small increases (1-8%) in liver and kidney weights
and some hematological effects of uncertain toxicological significance. The Dow (1962) report
is an unpublished study and reported increased liver and kidney weights in rats and increased
liver weights in rabbits. The Freundt et al. (1977) subchronic inhalation study is a one-
concentration study in which animals were exposed for periods of 1 to 16 weeks. This study
reported liver effects (fatty accumulation in liver lobules and Kupffer cells) that were also seen
in some of the controls. For each of the exposure durations, the incidence and severity of fat
accumulation increased with increasing exposure duration; however, these increases were not
statistically significantly different from the controls. Similar histopathologic findings in the liver
were not observed in DuPont (1998) at exposure concentrations 20-fold higher than the
concentration used in Freundt et al. (1977), nor in the NTP 90-day oral study at concentrations in
the diet up to 50,000 ppm.
Under the Guidelines for Carcinogen Risk Assessment (U .S. EPA, 2005a), there is
"inadequate information to assess the carcinogenic potential" of cis- or trans-1,2-DCE. This
descriptor reflects the lack of human epidemiological investigations or chronic animal bioassays.
6.2. DOSE RESPONSE
6.2.1. Noncancer — Oral Exposure
6.2.1.1. cis-l,2-DCE
McCauley et al. (1995, 1990) conducted the only available subchronic study of
cis-1,2-DCE. This 90-day gavage study was used as the basis for the oral RfD. An increase in
relative kidney weight in male rats was selected as the critical effect for derivation of the RfD.
Relative kidney weights were increased by up to 27% in males and 23% in females. There were
no histopathological changes in the kidney. A 10% change in relative kidney weight compared
with the control was selected as the BMR level for this endpoint. BMD modeling was used to
calculate the POD by estimating the effective dose at a specified level of response (BMDio) and
its 95% lower confidence limit (BMDLio). All of the continuous models in U.S. EPA's BMDS
(version 1.4.1)(U.S. EPA, 2007) were fit to the relative kidney weight data. For both male and
female rat relative kidney weight data, the Hill model (restricted) provided the best fit of the
data, yielding a BMDio and BMDLio of 19.8 and 5.1 mg/kg-day, respectively, in males, and
55.2 and 10.4 mg/kg-day, respectively, in females. The POD for the RfD for cis-l,2-DCE was
chosen as 5.1 mg/kg-day, the lower of the male and female BMDLio values. Applying a
composite UF of 3,000 to the POD of 5.1 mg/kg-day yields an RfD of 0.002 mg/kg-day. The
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composite UF of 3,000 includes factors of 10 to protect sensitive individuals, 10 to extrapolate
from animals to humans, 10 for use of a study of subchronic duration, and 3 to account for
database deficiencies. Information was unavailable to quantitatively assess toxicokinetic or
toxicodynamic differences between experimental animals and humans (applied a factor of 10) or
the potential variability in human susceptibility (applied a factor of 10) to cis-l,2-DCE. In the
absence of any chronic toxicity studies, a UF of 10 was used to account for extrapolating from a
subchronic study to estimate chronic exposure conditions. An UF of 3 was used to account for
deficiencies in the database, including lack of reproductive and developmental toxicity data for
the cis-isomer. The potential for developmental toxicity of cis-l,2-DCE, however, is informed
by a series of oral range-finding studies of the developmental toxicity of a mixture of 1,2-DCE
isomers (composition of isomers unknown) (NTP, 1991a, b, c). No evidence of developmental
toxicity was observed in mice or rats based on the parameters evaluated in these range-finding
studies (gravid uterus weight, fetal body weight, and number of fetuses [live/dead], implantation
sites, and resorptions).
Confidence in the principal study (McCauley et al. 1995, 1990) is medium. The 90-day
oral gavage study (McCauley et al. 1995, 1990) used four dose groups plus a control and
measured multiple parameters, including body weight, liver weight, kidney weight, clinical
chemistry, and hematology parameters. There are no oral studies of chronic, reproductive, or
developmental toxicity of cis-l,2-DCE. The McCauley et al. (1995, 1990) study is the only
available subchronic study of cis-1,2-DCE and was used as the basis for the oral RfD. However,
the developmental toxicity potential is informed by several range-finding studies for a mixture of
cis-l,2-DCE isomers (NTP, 1991a, b, c) that showed no evidence of developmental toxicity.
Thus, the confidence in the database is low to medium. The overall confidence in the RfD is
low.
6.2.1.2. trans-1,2- DCE
The Shopp et al. (1985) study was chosen as the principal study. Shopp et al. (1985)
reported a statistically significant dose-related suppression of sRBC-specific AFCs in the spleen
in male mice exposed to trans-1,2-DCE in drinking water for 90 days. The authors of this study
reported marked suppression in humoral immune status in male mice as indicated by the
significantly decreased number of AFCs. As described in more detail in Section 4.6.1.2, EPA
concluded that the 26% suppression in the number of sRBC-specific AFCs per 106 spleen cells
of male mice in Shopp et al. (1985) is a biologically significant measure indicating suppressed
antibody response associated with oral exposure to trans-1,2-DCE.
BMD modeling methods were used to calculate the POD by estimating the effective dose
at a specified level of response (BMDx) and its 95% lower confidence limit (BMDLx). A BMR
of 1 standard deviation from the control mean number of AFCs per 106 spleen cells was used.
All of the continuous models in U.S. EPA's BMDS (version 1.4. lc) (U.S. EPA, 2007) were fit to
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the data on numbers of AFCs (i.e., AFCs per 106 spleen cells) observed in male CD-I mice. A
second-degree polynomial model provided the best fit to these data, yielding a BMDLisd of
65.0 mg/kg-day. Thus, the POD for the RfD for trans- 1,2-DCE was the BMDLisd of
65.0 mg/kg-day. Applying a composite UF of 3,000 to the POD of 65 mg/kg-day yields an RfD
of 0.02 mg/kg-day. The composite UF of 3,000 includes a UF of 10 for intraspecies variability,
a UF of 10 for interspecies variability, a UF of 10 for extrapolation from a subchronic to a
chronic study, and a UF of 3 for database uncertainties. Information was unavailable to
quantitatively assess toxicokinetic or toxicodynamic differences between experimental animals
and humans (applied a factor of 10) or the potential variability in human susceptibility (applied a
factor of 10) to trans-1,2-DCE. In the absence of any chronic toxicity studies, a UF of 10 was
used to account for extrapolating from a subchronic study to estimate chronic exposure
conditions. An UF of 3 was used to account for deficiencies in the database, including lack of a
multigeneration reproductive toxicity study.
Confidence in the principal study (Shopp et al., 1985) is medium. This 90-day
immunotoxicity study of oral exposure of male and female CD-I mice to trans-1,2-DCE
(administered in drinking water) is a well-conducted, peer reviewed study. The Shopp et al.
(1985) study included three dose groups as well as a vehicle control group. Animals were
evaluated for humoral immune status as measured by the ability of spleen cells from these mice
to produce splenic IgM AFCs against sRBC, hemagglutination titers to sRBC, and by spleen cell
response to LPS. Confidence in the oral database is low to medium. Four subchronic studies
were considered in the evaluation of oral exposure to trans-l,2-DCE (NTP, 2002; Hayes, et al.,
1987; Barnes et al., 1985; Shopp et al., 1985). These studies evaluated a wide range of toxicity
endpoints, including hematology, urinalysis, clinical chemistry, histopathology, and immune
system function. Developmental toxicity potential for trans-1,2-DCE is informed by a series of
oral range-finding studies of a mixture of 1,2-DCE isomers (NTP, 1991a, b, c) that showed no
evidence of developmental toxicity.. There are no chronic studies of trans- 1,2-DCE toxicity.
The overall confidence in the RfD is low.
6.2.2. Noncancer — Inhalation Exposure
6.2.2.1. cis-l,2-DCE
There are no human studies, nor chronic or subchronic inhalation studies in animals for
cis-1,2-DCE. In the absence of a long-term inhalation study, no RfC was derived.
6.2.2.2. trans-l,2-DCE
The available inhalation data for trans-1,2-DCE were considered insufficient to support
reference value derivation. An RfC for trans-1,2-DCE was not derived.
6.2.3. Cancer
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Epidemiologic studies of 1,2-DCE are not available, and chronic bioassays in
experimental animals have not been performed. Thus, the toxicological databases for both
isomers provide "inadequate information to assess their carcinogenic potential."
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nceapublication. cfm ? ActT ype=PublicationT opics&detype=DOCUMENT&subj ect=
BENCHMARK+DOSE&subjtype=TITLE&excCol=Archive.
U.S. EPA. (2000c) Supplementary guidance for conducting health risk assessment of chemical mixtures. Risk
Assessment Forum, Washington, DC; EPA/630/R-00/002. Available online at
http://cfpub.epa.gov/ncea/raf/chem_mix.cfm.
U.S. EPA. (2002a) Toxicological review of 1,1-dichloroethylene (CAS No. 75-35-4) in support of summary
information on the Integrated Risk Information System (IRIS). Integrated Risk Information System (IRIS). National
Center for Environmental Assessment, Washington, DC. Available online at http://www.epa.gov/iris.
U. S. EPA. (2002b) A review of the reference dose concentration and reference concentration processess. Risk
Assessment Forum, Washington, DC; EPA/630/P-02/002F. Available online at
http://cfpub. epa.gov/ncea/raf/raf_pubtitles. cfm?detype=document&excCol=archive.
U.S. EPA. (2003) Toxicological review for dichloroacetic acid. Integrated Risk Information System (IRIS),
National Center for Environmental Assessment, Washington, DC. Available online at
http://www.epa.gov/iris/toxreviews/0654-tr.pdf.
U.S. EPA. (2005a) Guidelines for carcinogen risk assessment. Federal Register 70(66): 17765—18717. Available
online at http://www.epa.gov/cancerguidelines.
U.S. EPA. (2005b) Supplemental guidance for assessing susceptibility from early-life exposure to carcinogens. Risk
Assessment Forum, Washington, DC; EPA/630/R-03/003F. Available online at
http://cfpub. epa.gov/ncea/cfm/recordisplay. cfm?deid=l 60003.
U.S. EPA. (2006a) Science policy council handbook: peer review. 3rd edition. Office of Science Policy, Office of
Research and Development, Washington, DC; EPA/100/B-06/002. Available online at
http://www.epa.gov/OSA/spc/2peerrev.htm.
U.S. EPA. (2006b) A Framework for Assessing Health Risk of Environmental Exposures to Children. National
Center for Environmental Assessment, Washington, DC, EPA/600/R-05/093F. Available from:
.
U.S. EPA. (2007) Benchmark dose software (BMDS) version 1.4.1. (last modified February 2007). Available from:
.
U. S. EPA). (2007) Benchmark dose software (BMDS) version 2.1 (last modified November 2009). Available from:
.
Vogel, EW; Nivard, MJ. (1993) Performance of 181 chemicals in a drosophila assay predominantly monitoring
interchromosomal mitotic recombination. Mutagenesis 8:57-81.
Wan, J; Ernstgard, L; Song, BJ; et al. (2006) Chlorzoxazone metabolism is increased in fasted Sprague-Dawley rats.
JPharm Pharmacol 58:51-61.
WHO (World Health Organization). (1996) Principles and methods for assessing direct immunotoxicity associated
with exposure to chemicals. A report of the International Programme on Chemical Safety (Environmental Health
Criteria; 180, World Health Organization, Geneva.
Yannai, S. (1983). Adrenocortical response to single and repeated doses of chloroform in rats. Arch Toxicol
54:145-156.
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Zeiger, E; Anderson, B; Haworth, S; et al. (1988) Salmonella mutagenicity tests. 4. Results from the testing of 300
chemicals. Environ Mol Mutagen 11:1-158.
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APPENDIX A: SUMMARY OF EXTERNAL PEER REVIEW, PUBLIC COMMENTS,
AND DISPOSITION
The Toxicological Review of cis- and trans-1,2-Dichloroethylene (August, 2009) has
undergone a formal external peer review performed by scientists in accordance with EPA
guidance on peer review (U.S. EPA, 2006a, 2000a). An external peer review meeting was held
December 17, 2009. The external peer reviewers were tasked with providing written answers to
general questions on the overall assessment and on chemical-specific questions in areas of
scientific controversy or uncertainty. A summary of significant comments made by the external
reviewers and EPA's responses to these comments arranged by charge question follow. In many
cases the comments of the individual reviewers have been synthesized and paraphrased in
development of Appendix A. EPA also received scientific comments from the public. These
comments and EPA's responses are included in a separate section of this appendix.
EXTERNAL PEER REVIEW PANEL COMMENTS
The reviewers made several editorial suggestions to clarify specific portions of the text.
These changes were incorporated in the document as appropriate and are not discussed further.
I. General Charge Questions and Comments
1. Is the Toxicological Review logical, clear and concise? Has EPA accurately, clearly and
objectively represented and synthesized the scientific evidence for noncancer and cancer hazard?
Comment: In general, the reviewers observed that the Toxicological Review was logical, clear,
and concise, and that EPA accurately and clearly represented and synthesized the scientific
information. One reviewer stated that the Toxicological Review was not concise and that format
requirements resulted in redundancy. One reviewer suggested that a new subsection entitled
"Interactions with Other Chemicals" be added under Toxicokinetics to emphasize that DCEs
may be protective against cytotoxic and mutagenic/carcinogenic actions of VOCs and other
chemicals that undergo CYP2E1-catalyzed metabolic activation. Another reviewer suggested
that a summary paragraph on 1,2-DCE metabolism be added that includes a discussion of the
likely toxic metabolite and the suicide inhibition effect.
Response: The Toxicological Review was revised to eliminate redundant text wherever possible.
A new section, Section 3.3.3, CYP2E1 Inactivation by 1,2-DCE, was added to the
Toxicokinetics section to describe the potential role that DCEs may play in the metabolism of
other chemicals that undergo CYP2E1-catalyzed metabolic activation. Section 3.3, Metabolism,
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was revised to include a summary paragraph on 1,2-DCE metabolism.
Comment: One reviewer suggested that the document indicate that, based on the toxicity data
available from the general literature, the relative potency of cis- and trans-1,2-DCE is not high.
This reviewer also suggested adding to the Toxicological Review that the toxicity of 1,2-DCE
has not been intentionally understudied, but rather that preliminary testing that demonstrated that
the compounds are not particularly toxic or genotoxic provided little impetus to conduct studies
of chronic and reproductive toxicity.
Response: The relatively low toxicity of 1,2-DCE is discussed in Section 4.6. The observation
that demonstration of limited toxicity in the available animal studies provided little impetus to
conduct more extensive testing of these chemicals, while possibly correct, is speculative and was
not added to the Toxicological Review.
Comment: One reviewer stated that the synthesis of the evidence for hazard was superficial with
respect to concerns that corn oil may exacerbate hepatotoxicity of chloroalkenes and concerns
about whether the database supports deriving an RfD for the trans isomer.
Response: Based on peer reviewer comments on the cis-1,2-DCE RfD, the critical effect was
changed from increased liver weight to increased kidney weight. The RfD for trans-1,2-DCE is
based on immunotoxicity. While effects on the liver were considered as candidate critical
effects, oral toxicity values for 1,2-DCE isomers were not based on hepatotoxicity. Therefore,
consideration of uncertainty in the RfD associated with potential influence of corn oil on the
hepatotoxicity of 1,2-DCE was no longer relevant.
2. Please identify any additional studies that should be considered in the assessment of the
noncancer and cancer health effects of cis- and trans-l,2-dicholoroethylene.
Comment: The following additional papers were identified by the peer reviewers for
consideration in the assessment:
Ahmed, U; Redgrave, TG; Oates, PS. (2009) Effect of dietary fat to produce non-
alcoholic fatty liver in the rat. Journal of Gastroenterology and Hepatology, 24(8): 1463-
1471.
Caldwell, JC; Keshava, N. (2006) Key issues in the modes of action and effects of
trichloroethylene metabolites for liver and kidney tumorigenesis. Environmental Health
Perspectives 114(9): 1457-63.
Chetty, KN; et al. (2006) Cholesterol-induced alteration in liver mineral concentrations in
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corn oil and olive oil fed rats. Pathophysiology: The Official Journal of the International
Society for Pathophysiology 13(l):35-37.
Condie, LW. (1985) Target organ toxicology of halocarbons commonly found
contaminating drinking water. The Science of the Total Environment 47:433-442.
Huber, WW; et al. (1997) Inhibition instead of enhancement of lipid peroxidation by
pretreatment with the carcinogenic peroxisome proliferator nafenopin in rat liver exposed
to a high single dose of corn oil. Archives of Toxicology 71(9):575-581.
Raymond, P; Plaa, GL. (1997) Effect of dosing vehicle on the hepatotoxicity of CCI4 and
nephrotoxicity of CHCI3 in rats. Journal of Toxicology and Environmental Health
51(5):463-476.
Rivera, CA; et al. (2006) Feeding a corn oil/sucrose-enriched diet enhances
steatohepatitis in sedentary rats. American Journal of Physiology. Gastrointestinal and
Liver Physiology 290(2):G386-393.
Response: Relevant information from these papers was incorporated into appropriate sections of
the Toxicological Review.
II. Chemical-Specific Charge Questions and Comments
(A) Oral Reference Dose (RfD) for cis-l,2-DCE
1. The McCauley et al. (1990, 1995) subchronic gavage study in rats was selected as the basis for
the derivation of the RfD for cis-1,2-DCE. Please comment on whether the selection of this
study as the principal study is scientifically justified. Please identify and provide the rationale
for any other study that should be selected as the principal study.
Comment: All five reviewers agreed that the McCauley et al. (1995) study is the best available
study and should be used as the principal study for the derivation of the RfD for cis-l,2-DCE.
Response: The study has been retained as the principal study for the derivation of the RfD for
cis-l,2-DCE.
Comment: One reviewer recommended that an RfD for cis-l,2-DCE not be derived in light of
the discrepancies noted between the published and unpublished versions of the McCauley et al.
study and use of corn oil as the vehicle in this gavage study. This reviewer noted that corn oil by
itself can enhance hepatic lipid peroxidation and thereby the toxicity of the compounds and that
such interactions have been noted between corn oil and chloroalkenes. A second reviewer
suggested that documentation of the inconsistencies between the published and unpublished
McCauley et al. studies be provided.
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Response: Text was added to Sections 4.2.1.1.1 and 4.2.1.2.1 to more fully describe the
discrepancies between the published and unpublished McCauley et al. studies. These
discrepancies were not considered to compromise the integrity of the data since the
inconsistencies were more likely an issue of the quality of the report writing than an issue with
the findings themselves.
As noted in response to general charge question 1, the critical effect for the cis-l,2-DCE
RfD was changed from increased relative liver weight to increased relative kidney weight.
Because hepatotoxicity is not the basis for the cis-l,2-DCE RfD, the potential for corn oil to
enhance hepatic toxicity of 1,2-DCE is not an issue.
Comment: One reviewer suggested the historical ranges for liver and kidney weights and
absolute organ weight be provided in Table 4-1.
Response: No historical ranges for liver and kidney weight from this laboratory were available.
Because absolute liver and kidney weight data were not reported in the published McCauley et
al. study, these values were not added to Table 4-1; however, absolute liver and kidney weights
as reported in the unpublished McCauley et al. report were added to the text of Section 4.2.1.2.1
2. Increased relative liver weight in male rats (McCauley et al., 1990, 1995) was selected as the
critical effect for the RfD for cis-1,2-DCE. Please comment on whether the selection of this
critical effect is scientifically justified. Please identify and provide the rationale for any other
endpoint that should be considered in the selection of the critical effect.
Comment: One reviewer questioned whether absolute liver weights were significantly elevated
and stated that an increase in absolute liver weight would be more toxicologically significant
than the increase in relative liver weight.
Response: Absolute liver weights as reported in the unpublished McCauley et al. study and
results of statistical significance testing were added to Section 4.2.1.2.1. In general, relative
weight is considered more toxicologically relevant as a measure of effect on organ weight than
absolute organ weight because relative organ weight adjusts for any effect of the chemical on
body weight. Thus, relative organ weight changes were used over absolute organ weight
changes as candidate critical effects.
Comment: One reviewer concurred with the selection of liver weight over kidney weight as the
critical effect but asked for a more explicit discussion of the quantitative effect of this choice in
light of the fact that a significant effect on kidney weight in male rats was observed at a lower
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dose than the liver effect. Another reviewer believed that increased liver weight represented the
critical effect and that a convincing argument had been presented that the liver weight effect was
chemical-related, but observed that the effect was relatively small and not associated with
histopathology or serum liver enzyme changes. This reviewer recommended that the
Toxicological Review provide a more scientifically-based argument for increase liver weight as
an adverse effect or precursor to an adverse effect. Three reviewers suggested that it may be
worthwhile to look at the kidney as a potential critical target organ. One reviewer suggested that
hypercalcinemia be considered at a potential critical effect.
Response: The data for increased relatively kidney weight in the McCauley et al. (1995) study
were evaluated. A POD based on increased relative kidney weight was derived using BMDS
modeling methods. Tables B-3 and B-4 present the goodness-of-fit statistics and BMD and
BMDL estimates for all continuous models fit to these data for male and female rats,
respectively. BMDS modeling of relative kidney weight data for the male and female rat yielded
BMDLio values of 5.1 and 10.4 mg/kg-day, respectively. Both relative liver and kidney weight
increases were presented as potential candidates for the critical effect. Because the BMDLio of
5.1 mg/kg-day based on male rat data is more sensitive than the BMDLio values based on liver
weight data, increased relative kidney weight in male rats was selected as the critical effect.
Sections 5.1.1.1-5.1.1.3, 6.1, 6.2, and Appendix B were revised to reflect this change.
Section 4.6.1.1 discusses the fact that the observed increases in liver and kidney weight
could represent early indicators or precursors of liver and kidney toxicity, but that the available
toxicity literature is not sufficient to predict whether liver or kidney damage would or would not
occur at higher concentrations or in studies of longer exposure duration (i.e., chronic studies).
Hypercalcinemia was also evaluated as a potential critical effect. Calcium levels in the
90-day McCauley et al. (1995) study were statistically significantly increased as compared to
controls at the two low doses in male rats only; no increases in calcium levels were seen in
female rats. In the 14-day study there were statistically significant increases at the two high-
doses in male rats (10 and 12%, respectively) but no significant increases were noted in female
rats. Since there was no evidence of a dose-response relationship for hypercalcinemia in the 90-
day study despite suggestion of an increase in the 14-day study, the increased levels of calcium
were considered to be transient and not biologically significant.
3. BMD modeling methods were applied to liver weight data to derive the POD for the RfD. Has
the BMD modeling been appropriately conducted? Is the BMR selected for use in deriving the
POD (i.e., a 10% change in relative liver weight) scientifically justified? Please identify and
provide the rationale for any alternative approaches (including the selection of the BMR, model,
etc.) for the determination of the POD and discuss whether such approaches are preferred to
EPA's approach.
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Comment: Of the reviewers who commented on BMD modeling methods as they were applied to
liver weight data to derive the POD for the RfD, all agreed that BMD modeling is a reasonable
approach, was appropriately conducted, and was scientifically justified.
Response: No response is required.
Comment: One reviewer recommended that BMD modeling be applied to kidney weight data.
Response: As described in response to a comment on charge question A2 for the cis-l,2-DCE
RfD, relative kidney weight data for male and female rats were modeled to produce candidate
PODs. Increased relative kidney weight in the male rat was selected as the basis for the POD for
the RfD.
4. Please comment on the rationale for the selection of the uncertainty factors (UFs) applied to
the POD for the derivation of the RfD. If changes to the selected UFs are proposed, please
identify and provide a rationale(s).
Comment: Three reviewers considered the intraspecies UF of 10 to be justified in the absence of
information to suggest a smaller value. One reviewer recommended that the intraspecies UF of
10 be reduced to 3. This reviewer noted that the effects of 1,2-DCE are generally believed to be
due to their metabolites (via CYP2E1), that quantities of CYP2E1 and other constitutive CYP
isozymes in all persons are far in excess of the amounts necessary to metabolize all low levels of
1,2-DCE, and therefore that elevated CYP2E1 activity in some individuals is inconsequential to
the total amount of bioactive metabolite formed (i.e., eliminating the need for the toxicokinetic
component of the intraspecies UF).
Response: No information specific to variation in response to 1,2-DCE within the human
population is available to support a decrease from the intraspecies UF of 10. Therefore, the
intraspecies UF of 10 was retained.
Comment: One reviewer considered an interspecies UF of 10 to be reasonable. Three reviewers
recommended an interspecies UF of 3 instead of 10 to account for potential toxicodynamic
differences between animals and humans; it was suggested that the toxicokinetic component of
the UF be removed. One of these reviewers noted that rodents metabolize short-chain aliphatic
hydrocarbons to a greater extent than humans and that although the mode of action of cis-1,2-
DCE is unknown, it is generally accepted that oxidative metabolites (e.g., epoxides) are the most
likely candidates for proximate toxicants. Another of these reviewers added that the interspecies
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UF be reduced provided that documentation can be provided that cis-l,2-DCE is less toxic to
humans than to rats.
Response: An interspecies UF of 10 to extrapolate from rats to humans was retained due to lack
of chemical-specific data on toxicokinetic or toxicodynamic differences between rats and
humans.
Comment: One reviewer preferred an extrapolation from the LOAEL to NOAEL of 3 instead of
the uncertainty factor of 1 used in the derivation of the RfD for the cis isomer.
Response: It is current EPA practice to use an uncertainty factor of 1 for the extrapolation of
LOAEL to NOAEL when BMDS modeling is used. As noted in Section 5.1.1.3, the LOAEL to
NOAEL UF is addressed as one of the considerations in selection a BMR that represents a
minimal biologically significant change.
Comment: Three reviewers considered the subchronic to chronic UF of 10 to be appropriate.
One reviewer considered an UF of 10 for lack of chronic data to be probably too high in view of
the absence of adverse effects in inhalation experiments with cis-l,2-DCE or experiments with
mixed isomers; this reviewer did not propose an alternative UF.
Response: An UF of 10 to account for extrapolation from a subchronic to chronic exposure
duration to was retained in the absence of any chronic toxicity data to inform this extrapolation.
Comment: Three reviewers supported the database UF of 3. One reviewer did not see how
database deficiencies could be assigned a quantitative value in the form of an UF and for that
reason proposed a database uncertainty factor of 1. Another reviewer proposed a database
uncertainty factor of 10 on the basis of having only a single corn oil gavage study with a liver
endpoint that cannot be confirmed with any subchronic toxicity studies. This reviewer would
have been comfortable with a database uncertainty factor of 5 if kidney data were used to derive
the RfD since the evidence for interactions between corn oil and kidney toxicity are much
weaker and some internal data from the study support a potential effect on the kidney.
Response: The database UF of 3 was retained to address the lack of studies of reproductive
toxicity for this isomer consistent with Agency practice (U.S. EPA, 2002b). It is also Agency
practice to apply UFs of 1, 10°5 (rounded to 3), or 10 (U.S. EPA, 2002b); unless supported by
chemical-specific data, values for UFs other than 1, 3, or 10 are generally not applied in deriving
a reference value. Additionally, as noted in response to comments under charge question A2 for
cis-l,2-DCE, the kidney data were used to derive the RfD.
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(B) Oral Reference Dose (RfD) for trans-l,2-DCE
1. The 90-day immunotoxicity study by Shopp et al. (1985) was selected as the basis for the RfD
for trans- 1,2-DCE. Please comment on whether the selection of this study as the principal study
is scientifically justified. Please identify and provide the rationale for any other study that should
be selected as the principal study.
Comment: Four of the five reviewers supported selection of the immunotoxicity study by Shopp
et al. (1985) as the principal study and agree that its selection has been scientifically justified in
the document. One reviewer recommended the use of the NTP study as the principal study
rather than Shopp et al. (1985) because it is a well conducted study that did not use corn oil as
the vehicle of administration, thus avoiding any potential interaction between the test compound
and the corn oil vehicle, and because the authors of the Shopp et al. (1985) study did not find any
compelling evidence that trans-1,2-DCE was responsible for a biologically significant adverse
effect on the immune system.
Response: The Shopp et al. (1985) was retained as the principal study. The NTP study was
considered as a candidate study for RfD derivation but was not chosen as the principal study
because the candidate critical effect from this study (i.e., increased liver weight) was less
sensitive, and yielded a higher BMDLio (837 mg/kg-day) than the BMDLisd of 65 mg/kg-day
for suppression of the immune system from Shopp et al. (1985). The relevance and biological
significance of the immunotoxicity study by Shopp et al. (1985) is described in Sections 4.4.3.2
and 4.6.1.2.
Comment: One reviewer suggested that consideration be given to the Barnes et al. (1985) study
as a co-principal study unless a stronger rationale is given for excluding it because of observed
effects on absolute thymus weight. Additional comments offered by this reviewer on the
analysis of thymus weight data are addressed in charge question B2.
Response: Section 5.1.2 was revised to include Barnes et al. (1985) as a candidate principal study
in light of the reported effects on thymus weight.
2. Immune suppression, as indicated by the decrease of sheep red blood cell (sRBC)-specific
IgM antibody-forming cells (AFCs) in the spleen in male mice, was selected as the critical effect
for the RfD. Please comment on whether the selection of this critical effect is scientifically
justified. Please identify and provide the rationale for any other endpoint that should be
considered in the selection of the critical effect.
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Comment: Four of the five reviewers supported immune suppression, as indicated by the
decrease of sheep red blood cell (sRBC)-specific IgM antibody-forming cells (AFCs) in the
spleen in male mice, as the critical effect for the RfD. One reviewer did not favor use of immune
suppression as reported by Shopp et al. (1985) because of the authors' interpretation that the
AFC assay did not represent an immune response to trans-1,2-DCE (but rather reflected general
toxicity).
Response: A decrease of sheep red blood cell (sRBC)-specific IgM antibody-forming cells
(AFCs) in the spleen in male mice was retained as the critical effect for the trans-1,2-DCE RfD.
Differences in the interpretation of the Shopp et al. (1985) study findings reached by the study
investigators and by the EPA are discussed in Section 4.6.1.2 and in the consideration of
confidence in this study in Section 6.2.1.2.
Comment: One reviewer suggested that decreased absolute thymus weight in female mice in the
Barnes et al. (1985) study should be considered as a potential critical effect for developing a
candidate POD.
Response: Section 5.1.2 was revised to include decreased absolute thymus weight data from
Barnes et al. (1985) in female mice as a candidate critical effect. These data were modeled using
BMD methods. Because the BMDLisd based on AFC assay data was more sensitive than the
BMDLio based on absolute thymus weight, no change was made in the selection of the critical
effect.
Comment: One reviewer described the strengths and limitations of using decreased AFC
response as the critical effect for the derivation of the RfD for trans- 1,2-DCE and suggested that
support for this endpoint be added by pointing out that decreased thymus weights can be a good
indication for immunotoxicity and when accompanied by decreased AFC response, in the
absence of general toxicity, provides an excellent predictor of immunotoxicity (Luster et al.,
1992).
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Response: The discussion of uncertainties in the selection of decrease AFC response as the
critical effect for the trans-1,2-DCE RfD (Section 5.3) was expanded to address these strengths
and limitations.
Comment: One reviewer noted that the Toxicological Review did not provide a sufficiently
detailed scientific explanation of decrease in AFC response as adverse and biologically
significant to support selection of this endpoint as the critical effect.
Response: Text was added to Section 4.6.1.2 in support of a decrease in the AFC response as
biologically significant.
3. BMD modeling was applied to data for suppression of AFCs in the spleen in male mice in the
Shopp et al. (1985) study to derive the POD for the RfD. Has the BMD modeling been
appropriately conducted? Is the BMR selected for use in deriving the POD (i.e., a change in
response of 1 standard deviation from the control mean) scientifically justified? Please identify
and provide the rationale for any alternative approaches (including the selection of the BMR,
model, etc.) for the determination of the POD and discuss whether such approaches are preferred
to EPA's approach.
Comment: Four of the five reviewers noted that the BMD modeling was appropriately conducted
and that the BMR of a change in response of 1 standard deviation from the control mean was
scientifically justified. One reviewer, who disagreed with EPA's interpretation of the Shopp et
al. (1985) findings, suggested that the traditional NOAEL approach would be more defensible
than the BMD modeling approach.
Response: The estimation of the POD using BMD modeling methods was retained. EPA's
justification for interpretation of the Shopp et al. (1985) findings is provided in Section 4.6.1.2.
4. Please comment on the rationale for the selection of the UFs applied to the POD for the
derivation of the RfD. If changes to the selected UFs are proposed, please identify and provide a
rationale(s).
Comment: One reviewer agreed with the composite UF of 3,000 considering the uncertainty in
the toxicity of this compound, but did not provide comments on the individual UFs.
Response: No response is required.
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Comment: Three reviewers considered the intraspecies UF of 10 to be justified. As with the cis-
1,2-RfD, one reviewer recommended that the intraspecies UF of 10 be reduced to 3 because the
toxicokinetic component of the intraspecies UF was not necessary.
Response: No information specific to variation in response to 1,2-DCE within the human
population is available to support a decrease from the intraspecies UF of 10. Therefore, the
intraspecies UF of 10 was retained.
Comment: Two reviewers concurred with the interspecies UF of 10. Two reviewers
recommended an interspecies UF of 3 instead of 10 to account for potential toxicodynamic
differences between animals and humans; it was suggested that the toxicokinetic component of
the UF be removed.
Response: An interspecies UF of 10 to extrapolate from mice to humans was retained due to lack
of chemical-specific data on toxicokinetic or toxicodynamic differences between rats and
humans.
Comment: One reviewer concurred with the LOAEL to NOAEL UF of 1 (used because the
current approach is to address this factor as one of the considerations in selecting a BMR for
BMD modeling). A second reviewer noted that the value of 1 for this UF may be justified,
although a dose corresponding to a mean change of one standard deviation (SD) could be
interpreted as an excess risk of 10% above an assumed 1% background risk, which may
represent more than a minimal biologically significant change and therefore may warrant a large
UF.
Response: It is current EPA practice to use an UF of 1 for LOAEL to NOAEL extrapolation
when BMD modeling is used to derive the POD. The magnitude of change in the AFC assay
that represents a biologically significant effect has not been defined; therefore, a BMR of 1 SD
was applied consistent with current Agency practice under the assumption that it represents a
minimally biologically significant change (U.S. EPA, 2000b).
Comment: Three reviewers considered the subchronic to chronic UF of 10 to be appropriate.
The other reviewers did not offer a specific comment on this UF.
Response: No response is required.
Comment: Three reviewers supported the database UF of 3. As with the cis-1,2-DCE RfD, one
reviewer did not see how database deficiencies could be assigned a quantitative value in the form
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of an UF and for that reason proposed a database uncertainty factor of 1.
Response: The database UF of 3 was retained to account for the lack of a multigenerational
reproductive study.
(C) Inhalation Reference Concentration (RfC) for cis-l,2-DCE
1. An RfC was not derived due to the lack of available studies to characterize the health effects
associated with cis- 1,2-DCE administered via the inhalation route. Are there available data that
might support development of an RfC for cis-1,2-DCE?
Comment: Four of the reviewers were not aware of any other available study that might support
development of an RfC for cis-l,2-DCE.
Response: No response is required.
Comment: One reviewer considered all the oral data to be relevant for the estimation of an effect
level for the inhalation route since the effects of cis-l,2-DCE used for the RfD are systemic.
Response: In the absence of physiologically-based pharmacokinetic models for rats and humans,
EPA did not consider route-to-route extrapolation to be supported.
(D) Inhalation Reference Concentration (RfC) for trans-l,2-DCE
1. An RfC was not derived for trans-1,2-DCE. Has the scientific justification for not deriving an
RfC been clearly described in the document? Are there available data that might support
development of an RfC for trans-l,2-DCE?
Comment: One reviewer expressed serious reservations about using the findings of Freundt et al.
(1977) for the trans-1,2-DCE RfC, noting that the study is quite old, does not establish the purity
of the test chemical, does not clearly report the frequency or magnitude of the fatty changes that
are statistically significant, and the lack of concordance with the DuPont (1998) study which
showed an absence of adverse hepatic effects. Three reviewers proposed an evaluation of the
DuPont study (1998) as a possible basis for deriving an RfC for trans-1,2-DCE. One of these
three reviewers did not agree with the authors' interpretation of the hematology findings,
observing that the decreased WBC and lymphocyte counts may represent leucopenia and would
constitute an adverse effect. This reviewer further observed that the DuPont (1998) study
authors did not provide sufficient scientific support for their argument that the hematology
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findings were secondary to stress and lacked biological significance. Another reviewer
suggested that EPA evaluate organ weight changes reported in the DuPont (1998) study for
possible RfC derivation. This reviewer further recommended that if the organ weight data were
not suitable EPA should simply state that the database is inadequate for deriving an RfC.
Response: The RfC for trans-1,2-DCE was reevaluated in light of the submission of an
unpublished report by DuPont (1998) (previously described in the Kelly [1999] abstract) during
the public comment period. The DuPont study was reviewed and considered as a potential
principal study for RfC derivation. A summary and evaluation of the study findings were added
to Sections 4.2.2.2, 4.6.2.2, and 5.2.2. Relative liver and kidney weight changes in the DuPont
(1998) study, while likely treatment related, were relatively small, not generally statistically
significant and, therefore, were not considered to be an appropriate basis for the derivation of an
RfC. Some statistically significant hematology findings were reported in the DuPont (1998)
study (see a summary and discussion of these findings in Sections 4.6.2.2 and 5.2.2). The
toxicological significance of these effects is not clear. No effects on spleen or thymus weight
were found at 90 days, and no histopathological changes in these organs were reported. DuPont
suggested that decreases in WBC count and lymphocytes were secondary to "stress" and
elevated endogenous glucocorticoids. While the study authors did not provide support for this
hypothesis, the phenomenon of irritation/stress leading to decreased WBC or lymphocyte counts
has been observed following other chemical exposures (see discussion in Section 5.2.2.2).
Further, a similar reduction in WBC and lymphocyte counts was not observed in the 90-day NTP
(2002) dietary study of trans-1,2-DCE. For these reasons, identification of hematological
changes as a potential critical effect was not considered scientifically supported. Therefore,
while changes in some endpoints in the DuPont (1998) study were attributed to trans-l,2-DCE
exposure, none of the effects were identified as potential critical effects because of the
magnitude of the effect or questions about the biological significance of the effect.
The Freundt et al. (1977) study was reevaluated in light of the findings from the DuPont
(1998) study and determined not to be an adequate study for consideration as a principal study
for RfC derivation. Therefore, the available inhalation data were not considered sufficient to
derive an RfC for trans-1,2-DCE and Section 5.2.2 was revised accordingly.
Comment: One reviewer disagreed with the decision not to extrapolate data the oral route to
derive an inhalation RfC and the decision not to use PBPK modeling for route to route
extrapolation.
Response: It is EPA practice to perform route-to-route extrapolation for derivation of an RfD or
RfC only where such extrapolation can be conducted using PBPK modeling. As discussed in
Section 3.5, a PBPK model for cis- and trans-1,2-DCE has been developed for the rat, but this
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model was not calibrated with human data. Therefore, route-to-route extrapolation using PBPK
modeling was not considered appropriate.
Comment: One reviewer proposed an UF of 3 for subchronic to chronic extrapolation and an
intraspecies UF of 3 instead of 10, thereby allowing derivation of an RfC fortrans-l,2-DCE.
One reviewer disagreed with the decision not to derive an RfC for trans-l,2-DCE because the
combined UF would be 10,000. This reviewer proposed that uncertainty be truncated at 3,000
for estimation of a health protective level.
Response: In the absence of chemical-specific information on variation in response to trans-1,2-
DCE in the human population and the lack of information on effects following a chronic
exposure, EPA considers the use of 10-fold UFs for intraspecies and subchronic to chronic
extrapolation to be appropriate. Where the composite UF for a reference value exceeds 3,000 (or
where there is uncertainty in more than four areas of extrapolation), the RfD/RfC Technical
Panel concluded that it is unlikely that the database is sufficient to derive a reference value (U.S.
EPA, 2002b). EP A's treatment of the inhalation data for trans-1,2-DCE in the External Review
draft of the Toxicological Review was consistent with the recommendations of the Technical
Panel. Upon reevaluation of the trans-l,2-DCE inhalation database (i.e., Freundt et al., 1977 and
DuPont, 1998) EPA determined that the database was insufficient to support derivation of an
RfC fortrans-l,2-DCE. An examination of UFs was removed from Section 5.2.2.
Comment: One reviewer stated that EPA's method for calculating HECs for inhaled halocarbons
such as DCE is unsuitable and questioned whether the properties of 1,2-DCE were consistent
with the qualifier for a category 2 gas (i.e., reactive in respiratory tissue). This reviewer stated
that DCE is lipophilic and has quite limited water solubility and therefore its ability to penetrate
the mucus layer in the upper respiratory tract should be limited.
Response: EPA's evaluation ofthe HEC for trans-1,2-DCE in the External Review draft ofthe
Toxicological Review was consistent with the Methods for Derivation of Inhalation Reference
Concentrations and Application of Inhalation Dosimetry (U.S. EPA, 1994b). The approach
considers the physicochemical characteristics of the gas or vapor in question as well as the
toxicological specifics of the target tissue (respiratory versus systemic and, in the former case,
extrathoracic, thoracic, tracheobronchial, or pulmonary) and separates gases into three
categories. Trans-1,2-DCE qualifies as a category 2 gas: moderately water soluble, reactive in
respiratory tissue, and toxicologically active at remote sites (U.S. EPA, 1994b). For category 2
gases, HEC values are calculated by using methods for category 1 gases for portal-of-entry
effects and category 3 methods for systemic effects (U.S. EPA, 1994b). The candidate critical
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effects considered in the derivation of an RfC were systemic effects (e.g., effects on the liver);
thus, the methods for category 3 gases were considered appropriate.
Based on comments from peer reviewers and critical review of the unpublished DuPont
(1998) study, EPA determined that derivation of an RfC for trans-1,2-DCE was not supported.
Therefore, the HEC calculation was removed from Section 5.2.2.
(E) Carcinogenicity of cis- and trans-l,2-DCE
1. Under the EPA's 2005 Guidelines for Carcinogen Risk Assessment (www.epa.gov/iris/backgr-
d.htm), the Agency concluded that there is inadequate information to assess the carcinogenic
potential of cis- and trans-1,2-DCE. Please comment on the cancer weight of evidence
characterization. Is the cancer weight of evidence characterization scientifically justified?
Comment: All the reviewers agreed that there is inadequate information to assess the
carcinogenicity of 1,2-DCE isomers and that the cancer weight of evidence characterization
applied to these chemicals is scientifically justified.
Response: No response is required.
PUBLIC COMMENTS
General
Comment: A public commenter noted that the description of uses of DCEs was correct
historically, but that these uses did not reflect current use patterns. According to this commenter,
use of DCEs as a solvent for polymers and rubber is no longer in practice. Trans-1,2-DCE,
currently the only isomer commercially available in the U.S., is now used as a degreasing agent
and as one component of formulated products used for precision cleaning of electronic
components, and (in small amounts) as a blowing agent for specialty foams.
Response: Use information for 1,2-DCEs in Sections 2 and 6.1 was updated.
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Oral Reference Dose (RfD) for trans-l,2-DCE
Comment: A public commenter stated that there seemed to be no scientific justification to base a
proposed RfD for trans-1,2-DCE on AFC changes, an effect of unknown biological significance,
and recommended that the RfD be based on liver effects, an effect with commonly recognized
biological significance, as observed in Barnes et al. (1985).
Response: Three endpoints were considered as candidates for the critical effect, i.e., decreased
number of AFCs against sRBCs (Shopp et al., 1985), decreased absolute thymus weight (Barnes
et al., 1985) and increased liver weight (NTP, 2002). The dose-response analysis of the immune,
thymus and liver endpoints suggests that the immune system is more sensitive to the effects of
trans-1,2-DCE. Therefore, suppression of the humoral immune system, as measured by spleen
cell antibody production directed against sRBCs, was selected as the critical effect for the
trans-l,2-DCE RfD, and the Shopp et al. (1985) study was identified as the principal study.
Support for the biological significance of this effect in Section 4.6.1.2 was expanded.
Inhalation Reference Concentration (RfC) for trans-l,2-DCE
Comment: A public commenter submitted a complete study report of inhalation exposure to
trans-l,2-DCE (DuPont. 1998. trans-1,2-Dichloroethylene: 90-Day Inhalation Toxicity Study in
Rats, E.I. duPont de Nemours and Company, Laboratory Project ID: HL-1998-00952) that was
previously available as an abstract only (i.e., Kelly, 1999).
Response: A summary of the complete study report (DuPont, 1998) was added to Sections
4.2.2.2.2 and the findings were considered in Sections 4.6.2.2 and 5.2.2.
Comment: This public commenter recommended that the DuPont (1998) study serve as the
principal study for the trans-1,2-DCE RfC, and that the highest exposure concentration in this
study (identified as a NOAEL) be used as the POD. The commenter questioned the use of fatty
accumulation in the liver lobules and Kupffer cells from the Freundt et al. (1977) study as the
basis for the RfC, noting that interpretation of these findings was complicated by the finding of
similar effects in some control animals, the lack of statistical significance between control and
exposed groups, no evidence of functional changes in the livers of exposed animals, and lack of
corroboration of the liver findings in the DuPont study.
Response: Findings from the DuPont (1998) study were considered as the basis for the trans-1,2-
DCE RfC. As discussed in response to peer reviewer comments on charge question Dl, EPA
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concluded that the available inhalation toxicity database for trans-1,2-DCE, including Freundt et
al. (1977) and DuPont (1998), was insufficient for derivation of an RfC for this isomer.
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APPENDIX B: BENCHMARK DOSE MODELING RESULTS AND OUTPUTS
B.l. RfD for cis-l,2-DCE
B. 1.1. Relative Liver Weight
Relative liver weight, female rat (McCauley et al., 1995,1990)
Table B-l. BMDS modeling summary of relative liver weights in female
rats exposed to cis-l,2-DCE by gavage for 90 days
Model
Test 3
/j-Value
Test 4
/j-Yalue
AIC
BMD10
(mg/kg-d)
BMDLio
(mg/kg-d)
Linear,
Polynomial (restricted)
0.6325
0.0014
-84.9916
339.0
278.3
Power (>1)
0.6325
0.0014
-84.9916
339.0
278.3
Hill (>1)
0.6325
0.3208
-96.2572
80.5
42.3
BMR = 10% change in mean relative liver weight relative to the control mean
Constant variance
2
Only the Hill model [restricted] adequately described the data (test 4 % p > 0.1).
Hill Model. (Version: 2.12; Date: 02/20/2007)
Input Data File: G:\IRIS_CHEMICALS\DCE\BMD -- MCCAULEY\RELLIVERWTFEMALE.(d)
Gnuplot Plotting File: G:\IRIS_CHEMICALS\DCE\BMD -- MCCAULEY\RELLIVERWTFEMALE.pit
BMDS MODEL RUN
The form of the response function is:
Y[dose] = intercept + v*doseAn/(k^n + dose^n)
Dependent variable = MEAN
Independent variable = Dose(mg/kg/d)
rho is set to 0
Power parameter restricted to be greater than 1
A constant variance model is fit
Total number of dose groups = 5
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
alpha = 0.0446279
rho = 0 Specified
intercept = 2.82
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v = 0.85
n = 0.3078
k = 43 9.733
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -rho -n
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
alpha intercept v k
alpha 1 -4.5e-008 -7.8e-009 -2.9e-008
intercept -4.5e-008 1 -0.021 0.6
v -7.8e-009 -0.021 1 0.7
k -2.9e-008 0.6 0.7 1
Parameter Estimat<
95.0% Wald Confidence Interval
Variable
alpha
intercept
k
Estimate
). 0419186
2 .81563
1 . 04577
1
218.547
Std. Err.
0 . 0085566
0 . 0589669
0 .140777
NA
103 .222
Lower Conf. Limit
0 . 025148
2.70006
0.76985
16 . 235
Upper Conf. Limit
0 . 0586893
2 . 93121
1 .32169
420.858
NA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
has no standard error.
Table of Data and Estimated Values of Interest
Dose
Obs Mean
Est Mean Obs Std Dev Est Std Dev Scaled Res.
32
97
291
872
2 . 82
2 . 91
3 . 21
3 . 36
3 . 67
2 . 82
2 . 95
3 .14
3 . 41
3 . 65
19
18
22
18
27
0.205
0.205
0.205
0.205
0.205
0 . 0675
- 0.574
1. 07
- 0.817
0 .281
Model Descriptions for likelihoods calculated
Model A1: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma'*,2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma (i)A2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma*2
Model A3 uses any fixed variance parameters that
were specified by the user
Model R: Yi=Mu+e(i)
Var{e(i)} = Sigma*2
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Likelihoods of Interest
Model Log(likelihood) # Param's AIC
A1 53.265523 6 -94.531046
A2 54.549420 10 -89.098840
A3 53.265523 6 -94.531046
fitted 52.128602 4 -96.257205
R 23.670875 2 -43.341750
Test 1
Test 2
Test 3
Test 4
Explanation of Tests
and/or variances differ among Dose levels?
Do respons
(A2 vs. R)
Are Variances Homogeneous? (A1 vs A2)
Are variances adequately modeled? (A2 vs
Does the Model for the Mean Fit? (A3 vs.
A3)
fitted)
(Note: When rho=0 the results of Test 3 and Test 2 will be the same.)
Tests of Interest
Test -2*log(Likelihood Ratio) Test df p-value
Test 1 61.7571 8 <.0001
Test 2 2.56779 4 0.6325
Test 3 2.56779 4 0.6325
Test 4 2.27384 2 0.3208
The p-value for Test 1 is less than .05. There appears to be a difference between response
and/or variances among the dose levels. It seems appropriate to model the data
The p-value for Test 2 is greater than .1. A homogeneous variance model appears to be
appropriate here.
The p-value for Test 3 is greater than .1. The modeled variance appears to be appropriate here.
The p-value for Test 4 is greater than .1. The model chosen seems to adequately describe the
data.
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Relative risk
Confidence level = 0.95
BMD = 80.5212
BMDL = 42.3183
1
Standard deviation
0 . 95
53 .2
28 .7
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Hill Model with 0.95 Confidence Level
3.6
3.4
3.2
7
Bfyl.DL
0
B,MD
2.6
100
200
300
400
500
600
700
800
900
dose
15:30 08/29 2007
Figure displayed above is for a BMR = 10% change in mean relative liver weight relative to the
control mean.
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Relative liver weight, male rat (McCauley et al., 1995,1990)
Table B-2. BMDS modeling summary of relative liver weights in male rats
exposed to cis-l,2-DCE by gavage for 90 days
Model
Test 3
/j-Yalue
Test 4
p-Value
AIC
BMD10
(mg/kg-d)
BMDLio
(mg/kg-d)
Linear,
Polynomial (restricted)
0.04879
0.04268
-54.8404
379.4
281.1
Power (>1)
0.04879
0.04268
-54.8404
379.4
281.1
Hill (>1)
0.04879
0.1662
-57.4185
54.4
18.6
BMR = 10% change in mean relative liver weight relative to the control mean
Constant variance
2
Only the Hill model [restricted] adequately described the data (test 4 % p > 0.1).
Modeling of the variance (i.e., test 3 statistic in BMDS output) was not adequate (i.e., test
3 x p>0.1), but since BMR is not on a standard deviation basis, fitting a homogeneous
variance model is not essential.
Hill Model. (Version: 2.12; Date: 02/20/2007)
Input Data File: G:\IRIS_CHEMICALS\DCE\BMD -- MCCAULEY\RELLIVERWTMALE.(d)
Gnuplot Plotting File: G:\IRIS_CHEMICALS\DCE\BMD -- MCCAULEY\RELLIVERWTMALE.pit
BMDS MODEL RUN
The form of the response function is:
Y[dose] = intercept + v*doseAn/(k"n + dose'n)
Dependent variable = MEAN
Independent variable = Dose(mg/kg/d)
Power parameter restricted to be greater than 1
The variance is to be modeled as Var(i) = exp(lalpha + rho * ln(mean(i)))
Total number of dose groups = 5
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
lalpha = -2.56356
rho = 0
intercept = 2.85
v = 0.9
n = 0.109937
k = 420.333
Asymptotic Correlation Matrix of Parameter Estimate
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( *** The model parameter(s) -n
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
lalpha rho
lalpha 1 -1
rho -1 1
intercept 0.056 -0.06
v 0.25 - 0.25
k 0.23 - 0.23
intercept v k
0 - 056 0-25 0-23
-0-06 -0-25 - 0 - 23
1 -0-06 0-65
-0-06 1 0-62
0-65 0-62 1
Parameter Estimates
95.0% Wald Confidence Interval
Variable
lalpha
rho
intercept
Estimate
-5 .64025
2 .59276
2 - 88261
0 - 827803
1
101 - 77
Std. Err.
3 . 22635
2 - 74992
0 - 0845329
0 -173635
NA
82 - 8267
Lower Conf. Limit
-11.9638
-2.79698
2 - 71693
0 - 487485
-60 - 567
Upper Conf. Limit
0.683284
7 - 98251
3 - 04829
1 -16812
264 - 108
NA - Indicates that this parameter has hit a bound
irrplied by some inequality constraint and thus
has no standard error.
Table of Data and Estimated Values of Interest
Dose N Obs Mean Est Mean Obs Std Dev Est Std Dev Scaled Res.
0 9 2.85 2.88 0.26 0.235 -0.416
32 10 3.15 3.08 0.27 0.256 0.856
97 10 3.28 3.29 0.18 0.279 -0.0746
291 7 3.34 3.5 0.44 0.302 -1.37
872 6 3.75 3.62 0.2 0.316 0.976
Model Descriptions for likelihoods calculated
Model A1: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)*2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = exp(lalpha + rho*ln(Mu(i)))
Model A3 uses any fixed variance parameters that
were specified by the user
Model R: Yi=Mu+e(i)
Var{e(i)} = Sigma*2
Likelihoods of Interest
Model Log(likelihood) # Param's AIC
A1 35.496626 6 -58.993253
A2 39.438593 10 -58.877185
A3 35.503884 7 -57.007768
fitted 33.709241 5 -57.418482
R 20.057106 2 -36.114211
Explanation of Tests
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Test 1: Do responses and/or variances differ among Dose levels?
(A2 vs. R)
Test 2: Are Variances Homogeneous? (A1 vs A2)
Test 3: Are variances adequately modeled? (A2 vs. A3)
Test 4: Does the Model for the Mean Fit? (A3 vs. fitted)
(Note: When rho=0 the results of Test 3 and Test 2 will be the same.)
Tests of Interest
Test -2*log(Likelihood Ratio) Test df p-value
Test 1
Test 2
Test 3
Test 4
38 .763
7.88393
7.86942
3 .58929
<.
)1
0. 09592
0.04879
0.1662
The p-value for Test 1 is less than .05. There appears to be a difference between response
and/or variances among the dose levels. It seems appropriate to model the data.
The p-value for Test 2 is less than .1. A non-homogeneous variance model appears to be
appropriate.
The p-value for Test 3 is less than .1. You may want to consider a different variance model.
The p-value for Test 4 is greater than .1. The model chosen seems to adequately describe the
data.
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Relative risk
Confidence level = 0.95
BMD = 54.3727
1
Standard deviation
0 . 95
BMDL =
18.5549
13 . 0
Model with 0.95 Confidence Level
3.6
-------
B.1.2. Relative Kidney Weight
Relative kidney weight, male rat (McCauley et al., 1995,1990)
Table B-3. BMDS modeling summary of relative kidney weight in male
rats exposed to cis-l,2-DCE by gavage for 90 days
Model
Test 3
p-Value
Test 4
/> Value
AIC
BMD10
(mg/kg-d)
BMDLio
(mg/kg-d)
Hill (constant variance)
0.2879
0.2257
-210.4213
19.8
5.1
Polynominal - Linear
0.2879
0.0014
-199.9084
521.5
369.9
Polynominal (degree>2)
0.2879
0.0014
-199.9084
521.5
369.9
Power
0.2879
0.0014
-199.9084
521.5
369.9
BMR = 10% change in mean relative kidney weight relative to the control mean
Constant variance
Hill Model. (Version: 2.14; Date: 06/26/2008)
Input Data File: C:\USEPA\BMDS2l\Data\DCE\hil_DCEkidneym_hil-10%.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2l\Data\DCE\hil_DCEkidneym_hil-10%.pit
BMDS Model Run
The form of the response function is:
Y[dose] = intercept + v*doseAn/(k^n + dose^n)
Dependent variable = Kidney_weight_mean
Independent variable = Dose
rho is set to 0
Power parameter restricted to be greater than 1
A constant variance model is fit
Total number of dose groups = 5
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
ifault Initial Parameter Values
alpha = 0.00488
rho = 0 Specified
intercept = 0.7
v = 0.19
n = 0.362485
k = 33.6
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -rho -n
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
alpha intercept v k
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alpha 1 2.2e-007 l.2e-007 5.6e-007
intercept 2.2e-007 1 -0.7 0.41
v 1.2e-007 -0.7 1 0.2
k 5.6e-007 0.41 0.2 1
Parameter Estimates
95.0% Wald Confidence Interval
Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. Limit
alpha 0.00466147 0.000932293 0.00283421 0.00648873
intercept 0.701453 0.0218812 0.658567 0.74434
v 0.170516 0.0279749 0.115686 0.225345
n 1 NA
k 28.3985 19.5948 -10.0067 66.8036
NA - Indicates that this parameter has hit a bound Implied by some inequality constraint and thus
has no standard error.
Table of Data and Estimated Values of Interest
Dose N Obs Mean Est Mean Obs Std Dev Est Std Dev Scaled Res.
0 10 0.7 0.701 0.06 0.0683 -0.0673
32 10 0.8 0.792 0.06 0.0683 0.38
97 10 0.83 0.833 0.06 0.0683 -0.155
291 10 0.83 0.857 0.1 0.0683 -1.24
872 10 0.89 0.867 0.06 0.0683 1.08
Model Descriptions for likelihoods calculated
Model A1: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma (i)A2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma*2
Model A3 uses any fixed variance parameters that
were specified by the user
Model R: Yi=Mu+e(i)
Var{e(i)} = Sigma*2
Likelihoods of Interest
Model
A1
A2
A3
fitted
R
Log (likelihood)
110.699264
113.196292
110.699264
109.210632
94.871974
# Param's
AIC
-209.398529
-206.392585
-209.398529
-210.421264
-185.743948
Explanation of Tests
Test 1: Do responses and/or variances differ among Dose levels?
(A2 vs. R)
Test 2: Are Variances Homogeneous? (A1 vs A2)
Test 3: Are variances adequately modeled? (A2 vs. A3)
Test 4: Does the Model for the Mean Fit? (A3 vs. fitted)
(Note: When rho=0 the results of Test 3 and Test 2 will be the same.)
Tests of Interest
B-9
DRAFT - DO NOT CITE OR QUOTE
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Test -2*log(Likelihood Ratio) Test df
Test 1 36.6486 8
Test 2 4.99406 4
Test 3 4.99406 4
Test 4
2.97727
2
p-value
<.0001
0.2879
0.2879
0.2257
The p-value for Test 1 is less than .05. There appears to be a difference between response
and/or variances among the dose levels It seems appropriate to model the data
The p-value for Test 2 is greater than .1. A homogeneous variance model appears to be
appropriate here
The p-value for Test 3 is greater than .1. The modeled variance appears to be appropriate here
The p-value for Test 4 is greater than .1. The model chosen seems to adequately describe the
data
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Relative risk
Confidence level = 0.95
BMD = 19.8467
BMDL = 5.06583
1
Standard deviation
0. 95
19 . 0
5.1
Model with 0.95 Confidence Level
0.95
0.85
0.75
0.65
0 100 200 300 400 500 600 700 800 900
dose
10:23 03/05 2010
Figure displayed above is for a BMR = 10% change in mean relative kidney weight relative to the
control mean.
B-10
DRAFT - DO NOT CITE OR QUOTE
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Relative kidney weight, female rat (McCauley et al., 1995,1990)
Table B-4. BMDS modeling summary of relative kidney weight in female
rats exposed to cis-l,2-DCE by gavage for 90 days
Model
Test 3
p-Value
Test 4
/> Value
AIC
BMD10
(mg/kg-d)
BMDLio
(mg/kg-d)
Hill (constant variance)
<.0001
0.9839
-137.6262
55.2
10.4
Hill (non-constant variance)
0.0157
0.0564
-162.2055
37.4
failed
Polynominal - Linear
<0001
0.1020
-135.4199
499.3
278.6
Polynominal (degree>2)
<0001
0.4426
-137.9966
105.1
46.4
Power
<0001
0.1020
-135.4199
499.3
278.6
BMR = 10% change in mean relative kidney weight relative to the control mean
Constant variance
Hill Model. (Version: 2.14; Date: 06/26/2008)
Input Data File: C:\USEPA\BMDS2l\Data\Cai\DCE\hil_DCEkidneyf_hil-10%.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2l\Data\Cai\DCE\hil_DCEkidneyf_hil-l0%.pit
BMDS Model Run
The form of the response function is:
Y[dose] = intercept + v*doseAn/(k^n + dose^n)
Dependent variable = Kidney_weight_mean
Independent variable = Dose
rho is set to 0
Power parameter restricted to be greater than 1
A constant variance model is fit
Total number of dose groups = 5
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
alpha = 0.02134
rho = 0 Specified
intercept = 0.69
v = 0.16
n = 0.707673
k = 126.545
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -rho
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
alpha intercept v n k
alpha 1 - 9.5e-010 4e-009 -8.3e- 009 -8.2e-009
B-11 DRAFT - DO NOT CITE OR QUOTE
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intercept
-9.5e-01(
-8. 3e- 0(
-8.2e-0(
1
81
). 5
56
81
1
58
24
). 5
58
1
). 3
0 . 56
- 0 . 24
0 . 3
1
Parameter Estimates
Variable
alpha
intercept
Estimate
0. 0192062
0.690031
0 .16061
3 . 06776
60.5135
Std. Err.
0. 00384123
0 . 0438105
0 . 0554723
3.37011
39 .1149
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
).0116775
0 . 604164
0 . 051886
-3 . 53754
-16.1503
0 . 0267348
0 . 775898
0.269333
9.67306
137.177
Table of Data and Estimated Values of Interest
Doss
Obs Mean
Est Mean Obs Std Dev Est Std Dev Scaled Res.
32
97
291
872
'. 71
0 . 85
0 . 85
0 . 71
0 . 82
i. 851
05
23
21
'.13!
'.13!
'.13!
'.13!
'.13!
- 0.000717
0 . 001
-0. 00144
0 . 0148
-0 . 0136
Model Descriptions for likelihoods calculated
Model A1: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma*2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)*2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma'*,2
Model A3 uses any fixed variance parameters that
were specified by the user
Model R: Yi=Mu+e(i)
Var{e(i)} = Sigma'"2
Likelihoods of Interest
Model
A1
A2
A3
fitted
R
Log (likelihood)
73.813314
94.162787
73.813314
73.813110
68.169726
# Param's AIC
6 -135.626627
10 -168.325574
6 -135.626627
5 -137.626221
2 -132.339451
Explanation of Tests
Test 1: Do responses and/or variances differ among Dose levels?
(A2 vs. R)
Test 2: Are Variances Homogeneous? (A1 vs A2)
Test 3: Are variances adequately modeled? (A2 vs. A3)
Test 4: Does the Model for the Mean Fit? (A3 vs. fitted)
(Note: When rho=0 the results of Test 3 and Test 2 will be the same
Tests of Interest
Test -2*log (Likelihood Ratio) Test df p-value
Test 1 51.9861 8 <.0001
Test 2 40.6989 4 <.0001
B-12
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Test 3
Test 4
40.6989
0 . 000406677
4
1
<.0001
0.9839
The p-value for Test 1 is less than .05. There appears to be a
difference between response and/or variances among the dose levels
It seems appropriate to model the data
The p-value for Test 2 is less than .1. Consider running a
non-homogeneous variance model
The p-value for Test 3 is less than .1. You may want to consider a
different variance model
The p-value for Test 4 is greater than .1. The model chosen seems
to adequately describe the data
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Relative risk
Confidence level = 0.95
BMD = 55.1746
BMDL = 10.3 761
1
Standard deviation
0. 95
111. 0
failed
Hill Model with 0.95 Confidence Level
0.95
0.9
0.85
0.75
0.7
0.65
EiMDL
BMD
0
100
200
300
400
500
600
700
800
900
dose
10:22 03/12 2010
Figure displayed above is for a BMR = 10% change in mean relative kidney weight relative to the
control mean.
B-13
DRAFT - DO NOT CITE OR QUOTE
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B.2. RfD for trans-l,2-DCE
B.2.1. Decreased Antibody Directed Against sRBC (Shopp et al., 1985)
Table B-5. BMDS modeling summary of decreased antibody directed
against sRBC in male mice exposed to trans-l,2-DCE in drinking water for
90 days
Model
Test 3
Value
Test 4
/j-Value
AIC
BMD1SD
(mg/kg-d)
BMDL1sd
(mg/kg-d)
Polynomial, 2nd degree (unrestricted)
0.4558
0.7077
483.818
125.55
65.04
Polynomial, 1st degree (unrestricted)
0.4558
0.2596
484.375
309.20
195.01
Power (>1)
0.4558
0.2596
484.375
309.20
195.01
Hill (>1)
0.4558
NA
485.678
45.98
13.32
BMR = change in the mean response equal to 1 standard deviation from the control mean
Polynomial Model. (Version: 2.12; Date: 02/20/2007)
Input Data File: M:\NCEA\IRIS\DCE\BMDS RUNS\MALE_MICE_AFC_SHOPP_AG.(d)
Gnuplot Plotting File: M:\NCEA\IRIS\DCE\BMDS RUNS\MALE_MICE_AFC_SHOPP_AG.pit
BMDS MODEL RUN
The form of the response function is:
Y[dose] = beta_0 + beta_l*dose + beta_2*dose^2 + ...
Dependent variable = MEAN
Independent variable = DOSE
rho is set to 0
Signs of the polynomial coefficients are not restricted
A constant variance model is fit
Total number of dose groups = 4
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
alpha = 1
rho = 0 Specified
beta_0 = 2164.22
beta_l = -4.53587
beta 2 = 0.00808257
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -rho
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
alpha beta_0 beta_l beta_2
alpha 1 0.00028 -0.00047 0.00046
B-14
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beta_0
beta_l
beta 2
0 . 00028
-0 . 00047
0 . 00046
1
- 0 . 53
-0.53
1
-0. 97
-0.97
1
Variable
alpha
beta_0
beta_l
beta 2
Estimate
202454
2172.39
-4 - 61897
0 - 00824479
Parameter Estimates
Std. Err.
47770.9
106.683
1 - 97169
)- 00506139
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
296084
108825
1963 - 3
-8-48342
' - 00167535
2381 - 49
-0 - 754516
0 - 0181649
Table of Data and Estimated Values of Interest
Dose N Obs Mean Est Mean Obs Std Dev Est Std Dev Scaled Res.
0 12 2.2e+0 03 2.17e+003 433 450 0-213
17 8 2.05e+0 03 2.1e+003 43 0 450 - 0-3 03
175 8 1.63 e+0 03 1.62e+003 385 450 0-053
387 8 1.62e+0 03 1.62e+003 63 9 450 -0-0105
Model Descriptions for likelihoods calculated
Model A1: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma (i)A2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma*2
Model A3 uses any fixed variance parameters that
were specified by the user
Model R: Yi=Mu+e(i)
Var{e(i)} = Sigma*2
Likelihoods of Interest
Model
A1
A2
A3
fitted
R
Log (likelihood)
-237 - 839020
-236 - 534278
-237 - 839020
-237 - 909292
-243 - 158365
# Param's
5
AIC
485 - 67803 9
489 - 068557
485 - 67803 9
483 - 818585
490 - 31673 0
Explanation of Tests
Test 1: Do responses and/or variances differ among Dose levels?
(A2 vs. R)
Test 2: Are Variances Homogeneous? (A1 vs A2)
Test 3: Are variances adequately modeled? (A2 vs. A3)
Test 4: Does the Model for the Mean Fit? (A3 vs. fitted)
(Note: When rho=0 the results of Test 3 and Test 2 will be the same.)
Tests of Interest
Test -2*log(Likelihood Ratio) Test df p-value
Test 1 13.2482 6 0.03926
Test 2 2.60948 3 0.4558
Test 3 2.60948 3 0.4558
B-15
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Test 4 0.140546 1 0.7077
The p-value for Test 1 is less than .05. There appears to be a
difference between response and/or variances among the dose levels
It seems appropriate to model the data
The p-value for Test 2 is greater than .1. A homogeneous variance
model appears to be appropriate here
The p-value for Test 3 is greater than .1. The modeled variance appears
to be appropriate here
The p-value for Test 4 is greater than .1. The model chosen seems
to adequately describe the data
Benchmark Dose Computation
Specified effect = 1
Risk Type = Estimated standard deviations from the control mean
Confidence level = 0.95
BMD = 125.55
BMDL =
65.0386
Polynomial Model with 0.95 Confidence Level
2600
Polynomial
551800
BMDU
1200
1000 _
0 50
09:40 03/04 2009
100 150 200
dose
250 300 350 400
B-16
DRAFT - DO NOT CITE OR QUOTE
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B.2.2. Absolute Thymus Weight (Barnes et aL, 1985)
Table B-6. BMDS modeling summary of decreased absolute thymus
weight in female mice exposed to trans-l,2-DCE in drinking water for 90
days
Model
Test 3
Value
Test 4
/j-Value
AIC
BMD10
(mg/kg-d)
BMDLio
(mg/kg-d)
Hill (constant variance)
0.9820
0.5511
471.58
153.26
10.94
Polynominal - linear
0.9820
0.7895
469.70
196.13
138.49
Polynominal - (degree>2)
0.9820
0.5422
471.60
161.80
70.75
Power
0.9820
0.7895
469.70
196.13
138.49
BMR = 10% change in mean absolute thymus weight relative to the control mean
Polynomial Model. (Version: 2.13; Date: 04/08/2008)
Input Data File: C:\USEPA\BMDS2l\Data\DCE\Abs-Thymus-Barnes\lin_Thymus-f_lin-10%-(d)
Gnuplot Plotting File: C:\USEPA\BMDS2l\Data\DCE\Abs-Thymus-Barnes\lin_Thymus-f_lin-
10%-pit
BMDS Model Run
The form of the response function is:
Y[dose] = beta_0 + beta_l*dose + beta_2*dose^2 + ...
Dependent variable = Mean
Independent variable = Dose
rho is set to 0
Signs of the polynomial coefficients are not restricted
A constant variance model is fit
Total number of dose groups = 4
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
alpha = 242.471
rho = 0 Specified
beta_0 = 69.2918
beta 1 = -0.0345935
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -rho
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
alpha
alpha 1
beta 0 le-008
beta_0 beta_l
le-008 -1.7e-0 09
1 -0.65
B-17 DRAFT - DO NOT CITE OR QUOTE
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beta 1 -1.7e-009 -0.65
Parameter Estimates
95.0% Wald Confidence Interval
Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. Limit
alpha 230.509 38.4181 155.21 305.807
beta_0 69.622 2.36029 64.9959 74.2481
beta 1 -0.0354984 0.00991539 -0.0549322 -0.0160645
Table of Data and Estimated Values of Interest
Dose N Obs Mean Est Mean Obs Std Dev Est Std Dev Scaled Res.
0 24 71 69.6 14.7 15.2 0.445
22.6 16 67 68.8 16 15.2 -0.479
224 16 61 61.7 16 15.2 -0.177
452 16 54 53.6 16 15.2 0.112
Model Descriptions for likelihoods calculated
Model A1: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma'"2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)A2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma*2
Model A3 uses any fixed variance parameters that
were specified by the user
Model R: Yi=Mu+e(i)
Var{e(i)} = Sigma*2
Likelihoods of Interest
Model
A1
A2
A3
fitted
R
Log (likelihood)
-231.613 992
-231.527960
-231.613 992
-231.850361
-237.748366
# Param's
5
8
5
3
2
AIC
473 . 227984
479 . 055919
473.227984
469.700721
479 .496732
Test 1
Test 2
Test 3
Test 4
(Note:
Explanation of Tests
and/or variances differ among Dose levels?
Do respons
(A2 vs. R)
Are Variances Homogeneous? (A1 vs A2)
Are variances adequately modeled? (A2 vs
Does the Model for the Mean Fit? (A3 vs.
When rho=0 the results of Test 3 and Test
A3)
fitted)
2 will be the same
Tests of Interest
Test
-2*log(Likelihood Ratio) Test df
p-value
Test 1
Test 2
Test 3
Test 4
12.4408
) . 172065
) . 172065
) . 472737
0 . 05283
0 . 982
0 . 982
0.7895
The p-value for Test 1 is greater than .05. There may not be a
diffence between responses and/or variances among the dose levels
Modelling the data with a dose/response curve may not be appropriate
B-18
DRAFT - DO NOT CITE OR QUOTE
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The p-value for Test 2 is greater than .1. A homogeneous variance
model appears to be appropriate here
The p-value for Test 3 is greater than .1. The modeled variance appears
to be appropriate here
The p-value for Test 4 is greater than .1. The model chosen seems
to adequately describe the data
Benchmark Dose Computation
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
0.1
Relative risk
0 . 95
196.127
138.488
1
Standard deviation
0. 95
427.70
Linear Model with 0.95 Confidence Level
80
75
70
65
60
55
50
45
15:36 03/18 2010
Linear
BMD Lower Bound
BMDL
Figure displayed above is for a BMR = 10% change in mean absolute thymus weight relative to the control
mean.
B-19 DRAFT - DO NOT CITE OR QUOTE
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B.2.3. Relative Liver Weight (NTP, 2002)
Relative liver weight, male mouse (NTP, 2002)
Table B-7. BMDS modeling summary of relative liver weight in male mice
exposed to trans-l,2-DCE in the feed for 14 weeks
Model
Test 3
p-Value
Test 4
p-Value
AIC
BMD10
(mg/kg-d)
BMDL10
(mg/kg-d)
Linear,
Polynomial (restricted)
0.2228
0.2974
-79.0425
7,063.8
5,109.8
Power (>1)
0.2228
0.2974
-79.0425
7,063.8
5,109.8
Hill (>1)
0.2228
0.6152
-80.1465
3,241.9
867.3
BMR = 10% change in mean relative liver weight relative to the control mean
Constant variance
All models used in the evaluation of relative liver weight in male mice produced outputs
with ^ p >0.1. The Hill model, with the lowest AIC values, provided the best fit of the data.
Hill Model. (Version: 2.12; Date: 02/20/2007)
Input Data File: G:\IRIS_CHEMICALS\DCE\BMD-- NTP\REL_LIVER_WT_MOUSEM.(d)
Gnuplot Plotting File: G:\IRIS_CHEMICALS\DCE\BMD-- NTP\REL_LIVER_WT_MOUSEM.pit
BMDS MODEL RUN
The form of the response function is:
Y[dose] = intercept + v*doseAn/(k"n + dose'n)
Dependent variable = MEAN
Independent variable = Dose(mg/kg-d)
rho is set to 0
Power parameter restricted to be greater than 1
A constant variance model is fit
Total number of dose groups = 6
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
alpha = 0.0912885
rho = 0 Specified
intercept = 4.347
v = 0.632
n = 0 . 479874
k = 2442.97
Asymptotic Correlation Matrix of Parameter Estimates
B-20 DRAFT - DO NOT CITE OR QUOTE
-------
( *** The model parameter(s) -rho -n
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
alpha
intercept
alpha
1
-2.5e-009
2e-008
9.8e-009
intercept
-2 . 5e-00 9
1
- 0. 02
v
2e-0 08
-0. 02
1
0. 65
. 65
1
Parameter Estimat<
95.0% Wald Confidence Interval
Variable
alpha
intercept
k
Estimate
). 0846603
4 . 36928
0.679896
1
1802.79
Std. Err.
0 . 0154568
0. 09406
0 .178916
NA
1800.91
Lower Conf. Limit
0 . 0543656
4 .18493
0 . 329227
-1726.93
Upper Conf. Limit
0.114955
4 - 55364
1 - 03 057
5332 - 51
NA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
has no standard error.
Table of Data and Estimated Values of Interest
Doss
Obs Mean
Est Mean Obs Std Dev Est Std Dev Scaled Res.
920
1900
3850
8065
4 . 35
4 . 55
4 . 75
4 . 74
37
51
72
83
92
177
357
364
266
) .25
351
'.291
'.291
'.291
'.291
'.291
'.291
- 0.242
0 .432
-0 . 0219
0.292
-1. 05
0 .587
Model Descriptions for likelihoods calculat<
Model A1: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma'"2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)A2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma*2
Model A3 uses any fixed variance parameters that
were specified by the user
Model R: Yi=Mu+e(i)
Var{e(i)} = Sigma*2
Likelihoods of Interest
Model
A1
A2
A3
fitted
R
Log (likelihood)
44.972729
48 . 458304
44.972729
44.073247
33.552988
# Param's
7
12
7
AIC
-75 . 945458
-72 . 916608
-75.945458
-80.146494
-63.105977
B-21
DRAFT - DO NOT CITE OR QUOTE
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Explanation of Tests
Test 1: Do responses and/or variances differ among Dose levels?
(A2 vs. R)
Test 2: Are Variances Homogeneous? (A1 vs A2)
Test 3: Are variances adequately modeled? (A2 vs. A3)
Test 4: Does the Model for the Mean Fit? (A3 vs. fitted)
(Note: When rho=0 the results of Test 3 and Test 2 will be the same.)
Tests of Interest
Test -2*log(Likelihood Ratio) Test df p-value
Test 1 29.8106 10 0.0009199
Test 2 6.97115 5 0.2228
Test 3 6.97115 5 0.2228
Test 4 1.79896 3 0.6152
The p-value for Test 1 is less than .05. There appears to be a difference between response
and/or variances among the dose levels. It seems appropriate to model the data.
The p-value for Test 2 is greater than .1. A homogeneous variance model appears to be
appropriate here.
The p-value for Test 3 is greater than .1. The modeled variance appears to be appropriate here.
The p-value for Test 4 is greater than .1. The model chosen seems to adequately describe the
data.
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Relative risk
Confidence level = 0.95
BMD = 3241.95
BMDL = 867.264
1
Standard deviation
0. 95
1348.69
395.878
B-22 DRAFT - DO NOT CITE OR QUOTE
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Hill Model with 0.95 Confidence Level
5.2
4.6
4.4
4.2
BMDL
BMD
0
1000
2000
3000
4000
5000
6000
7000
8000
dose
10:27 09/11 2007
Figure displayed above is for a BMR = 10% change in mean relative liver weight relative to the
control mean.
B-23
DRAFT - DO NOT CITE OR QUOTE
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Relative liver weight, female mouse (NTP, 2002)
Table B-8. BMDS modeling summary of relative liver weight in female
mice exposed to trans-l,2-DCE in the feed for 14 weeks
Model
Test 3
/7-Value
Test 4
/7-Value
AIC
BMD10
(mg/kg-d)
RMDLio
(mg/kg-d)
Linear,
Polynomial (restricted)
0.1553
0.001521
-74.100992
8,457.39
5,488.88
Polynomial, 2nd degree
(unrestricted)
0.1553
0.007562
-77.688156
3,807.76
2,027.13
Power (>1)
0.1553
0.001521
-74.100992
8,457.39
5,488.88
Power (unrestricted)
0.1553
0.007744
-77.739382
5,224.74
1,603.23
Hill (>1)
0.1553
0.003766
-76.473258
6,158.22
BMDL
computation
failed
Hill (unrestricted)
0.1553
0.003766
-76.473259
6,158.23
BMDL
computation
failed
BMR = 10% change in mean relative liver weight relative to the control mean
Nonhomogeneous variance
None of the models in BMDS (version 1.4.1) provided an adequate fit of the data for
relative liver weight in female mice from the NTP (2002) study.
B-24
DRAFT - DO NOT CITE OR QUOTE
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Relative liver weight, male rat (NTP, 2002)
BMD methods were not applied to male rat data because relative liver weights were not
significantly elevated over controls.
Relative liver weight, female rat (NTP, 2002)
Table B-9. BMDS modeling summary of relative liver weight in female
rats exposed to trans-l,2-DCE in the feed for 14 weeks
Model
Test 3
p-Value
Test 4
p-Value
AIC
BMD10
(mg/kg-d)
RMDLio
(mg/kg-d)
Linear,
Polynomial (restricted)
0.7
0.01169
-143.4885
5,447.8
3,362.9
Polynomial, 2nd degree
(unrestricted)3
0.7
0.008974
-142.8260
1,971.2
1,027.2
Power (>1)
0.7
0.01169
-143.4885
5,447.8
3,362.9
Power (unrestricted)
0.7
0.09092
-147.9366
6,165.2
539.0
Hill (>1)
0.7
0.2317
-150.1137
10% BMR is not in the range of the
fitted model
Hill (unrestricted)
0.7
0.2317
-150.1137
10% BMR is not in the range of the
fitted model
BMR = 10% change in mean relative liver weight relative to the control mean
aBMR = 8%, maximum within range of data.
Constant variance
None of the models in BMDS (version 1.4.1) provided an adequate fit of the data for
relative liver weight in female rats from the NTP (2002) study. The Hill model (with power
parameter restricted to be greater than 1) provided an adequate statistical fit of the relative liver
weight data (i.e., % p>0.1), but the curve generated by the model was not supported biologically
(i.e., the curve was essentially a step function, with almost no transition between the dose at
which no effect was observed and the dose causing a maximum effect). This data set did not
demonstrate as sensitive a response as the others.
B-25 DRAFT - DO NOT CITE OR QUOTE
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