United States
Environmental Protection
Agency
Office of Health and
Environmental Assessment
Washington DC 20460
EPA-600/8-82-005B
December 1983
External Review Draft
Research and Development
vvEPA
Health Assessment
Document for
Tetrachloroethylene
(Perch loroethylene)
Review
Draft
(Do Not
Cite or Quote)
NOTICE
This document is a preliminary draft. It has not been formally
released by EPA and should not at this stage be construed to
represent Agency policy. It is being circulated for comment on its
technical accuracy and policy implications.
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EPA-600/8-82-005B
Jan. 1984
External Review Draft
Health Assessment Document
for
Tetrachloroethylene
(Perchloroethylene)
NOTICE
This document is a preliminary draft. It has not been formally
released by the U.S. Environmental Protection Agency and
should not at this stage be construed to represent Agency policy.
It is being circulated for comment on its technical accuracy and
policy implications.
U.S. ENVIRONMENTAL PROTECTION AGENCY
Office of Research and Development
Office of Health and Environmental Assessment
Environmental Criteria and Assessment Office
Research Triangle Park, NC 27711
-------
DISCLAIMER
This report is an external draft for review purposes only and does not
constitute Agency policy. Mention of trade names or commercial products does
not constitute endorsement or recommendation for use.
003PE2/A 12/7/83
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Docket No. BCAP-HA-78-3
DRAFT HEALTH ASSESSMENT DOCUMENT FOR
TETRAOfljOROETHYLENE (PERCHLOROETHYLENE)
(ORD-FRL-2548-2)
AGENCY: Environmental Protection Agency
ACTION: Revision Clarifying the Carcinogenicity Conclusions of the Draft Health
Assessment Document for Tetrachloroethylene (Perchloroethylene), and
Reopening of Public Comment Period.
SUMMARY: The External Review Draft of the EPA document, Draft Health
Assessment Document for Tetrachloroethylene (Perchloroethylene),
EPA-600/8-82-005B dated December 1983 was announced on December 23, 1983
in the Federal Register as being available for public review and comment from
January 5, 1984 through March 5, 1984. On February 3, errata for pages 1-4 and
9-45 were distributed to all parties who had requested a copy of the document
from EPA and inserted in all undistributed copies.
Because of a need to further clarify the conclusion regarding the
carcinogenicity findings, the Agency is issuing a revised statement of these
conclusions.. This statement replaces the last four lines on page 1-3 and all
of page 1-4, and also replaces the Conclusions section 9.4.3 on page 9-45. The
revised statement is as follows:
Tetrachloroethylene has been demonstrated to induce malignant tumors of
the liver in both male and female mice of the B6C3F1 strain. This constitutes
a signal that tetrachloroethylene might be a carcinogen for humans.
The technical adequacy and the strong nature of the positive response in
the 1977 NCI tetrachlorethylene bioassay study makes it likely that a repeat
bioassay would also be positive. In fact, the recent National Toxicology
Program (NTP) study, currently under audit, showed similar positive results.
This bioassay study, if validated, would strengthen the evidence for carcinogen-
icity of tetrachloroethylene.
It should be recognized that there is a substantial body of opinion in
the scientific community to the effect that the mouse liver overreacts to
chlorinated organic compounds in contrast to that of the rat, and that the
induction of liver cancer in the mouse represents only a promoting action for
spontaneous liver tumors which normally occur with substantial incidence.
Furthermore, this promoting action might be related to liver damage associated
only with high level exposures to chlorinated agents such as tetrachlorethylene.
However, the evidence is inconclusive either for this restrictive position on
the mouse liver carcinogenic response to chlorinated organics or to the position
that the mouse liver is as good as any other mammalian indicator of carcinogenicity
for these compounds.
-------
According to a literal interpretation of the criteria of the International
Agency for Research on Cancer (IARC), the animal data supporting the carcinogenicity
of tetrachloroethylene might be classified as limited. Also, since existing
human epidemiologic data for tetrachloroethylene are inconclusive, its overall
IARC ranking might be classsified as Group 3, meaning, according to IARC language,
that perchloroethylene cannot be classified as to its human carcinogenicity.
It should be recognized that Group 3 covers a broad range of evidence:
from inadequate to almost sufficient animal data. Because of the stength of the
mouse liver cancer response, tetrachloroethylene is at the upper end of this
range. Hence, the classification of the carcinogenicity of tetrachloroethylene
under the IARC criteria for animal evidence could be limited or almost sufficient
depending on the nature of the bioassay evidence as it exists today and on the
differing current scientific views about the induction of liver tumors in mice
by chlorinated organic compounds. Therefore, the overall IARC ranking of
tetrachloroethylene is Group 3 but close to Group 2B, i.e. the more conservative
scientific view would regard tetrachloroethylene as being close to a probable
human carcinogen, but there is considerable scientific sentiment for regarding
tetrachloroethylene as an agent that cannot be classified as to its carcinogen-
icity for humans.
In consideration of the above action, the public comment period will be
reopened for 30 days beginning March 23, 1984 and ending April 23, 1984. In
addition to this notice, copies of this revision will be forwarded to those who
have already received copies of the draft health assessment document and errata
from the ORD Publications Office,. Center for Environmental Research Information
(CERI), in Cincinnati. During the 30-day public comment period, requesters may
obtain copies of these materials, as follows:
Single copies will be available from ORD Publications - CERI-FRN, U.S.
Environmental Protection Agency, 26 West St. Clair Street, Cincinnati, Ohio
45268. Tel. (513) 684-7562.
These documents also will be available for public inspection and copying
at the EPA library at Waterside Mall, 401 M Street, S.W., Washington, D.C. 20460.
Comments on the revision should be submitted in writing by close of business
on April 23, 1984 to: Project Officer for Tetrachloroethylene (Perchloroethylene),
Environmental Criteria and Asssessment Office (MD-52), U.S. Environmental
Protection Agency, Research Triangle Park, North Carolina 27711.
FOR FURTHER INFORMATION CONTACT: Ms. Diane Chappell, 919/541-3637.
March 15, 1984 /s/
Date Bernard D. Goldstein
Assistant Administrator
for Research and Development
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"twoi*5
UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, O.C. 20460
February 3, 1984
FROM:
TO:
OFFICE OF
RESEARCH AND DEVELOPMENT
SUBJECT: ERRATA
El IzabethTTfArj^erson
Director, Office of Health and
Environmental Assessment (RD-689)
Recipients of the Second External Review Draft of the Health
Assessment Document for Tetrachloroethylene (Perchloroethylene),
December 1983, EPA-600/8-82-005B
Because of errors on two pages 1n the Tetrachloroethylene document (pages
1-4 and 9-45), we are providing you with replacement pages for Insertion Into
the document. The shaded areas on the replacement pages Indicate the material
that has been revised.
We apologize for any Inconvenience these changes may have caused the
recipients.
Attachments-2
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UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
FPCM:
OFFICE OF
RESEARCH AND DEVELOPMENT
SUBJECT: Dra£t^fjear±*h Assessment Document for Tetrachloroethylene
rson
Director, Office of Health and
Environmental Assessment
TO:
Addressees
The attached second external review draft of the Health Assessment
Document for Tetrachloroethylene is provided for your information.
The draft document was made available for public review and comment on
January 5, 1984 (December 23, 1983, 48 FR 56847) and the Agency is accepting
public comments until March 5, 1984. If you would like to cement on this
document, please address your conments to:
Project Officer for Tetrachloroethylene
Environmental Criteria and Assessment Office (MD-52)
U.S. Environmental Protection Agency
Research Triangle Park, North Caroline 27711
Conments should be in writing and should be submitted by close-of-business
March 5, 1984.
A limited supply of copies will be available from the ORE publications
office in Cincinnati (FTS 684-7562).
After receipt of all comments, the EPA Science Advisory Board will review
the subject dco.:rsent in ~i public meeting. This meeting will be announced in a
subsequent Federal Register notice.
Attachment
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PREFACE
The Office of Health and Environmental Assessment has prepared this health
assessment to serve as a "source document" for EPA use. Originally the health
assessment was developed for use by the Office of Air Quality Planning and
Standards to support decision-making regarding possible regulation of PCE
as a hazardous air pollutant. However, at the request of the Agency's Work
Group on Solvents the assessment scope was expanded to address multimedia
aspects.
In the development of the assessment document, the scientific literature
has been inventoried, key studies have been evaluated, and summary/conclusions
have been prepared so that the chemical's toxicity and related characteristics
are qualitatively identified. Observed-effect levels and dose-response
relationships are discussed, where appropriate, so that the nature of the
adverse health responses is placed in perspective with observed environmental
levels.
003PE2/A 12/1/83
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CONTENTS
Page
1. EXECUTIVE SUMMARY 1-1
2. INTRODUCTION 2-1
3. GENERAL BACKGROUND INFORMATION 3-1
3.1 PHYSICAL AND CHEMICAL PROPERTIES 3-1
3.2 PRODUCTION 3-1
3. 3 USE 3-3
3.4 EMISSIONS 3-4
3.5 ENVIRONMENTAL FATE AND TRANSPORT 3-5
3.5.1 Ambient Air 3-5
3.5.2 Water 3-9
3.6 LEVELS OF EXPOSURE 3-10
3.6.1 Mixing Ratios 3-10
3. 7 ANALYTICAL METHOD 3-19
3.7.1 Ambient Air 3-19
3.7.2 Water 3-23
3.7.3 Biological Media 3-25
3.7.4 Calibration 3-25
3.7.5 Storage and Stability of PCE 3-25
3.8 REFERENCES 3-27
4. ECOSYSTEM CONSIDERATIONS 4-1
4.1 EFFECTS ON AQUATIC ORGANISMS AND PLANTS 4-1
4.1.1 Effects on Freshwater Species 4-1
4.1.2 Effects, on Aquatic Plants 4-2
4.1.3 Effects on Saltwater Species 4-2
4. 2 BIOCONCENTRATION AND BIOACCUMULATION 4-3
4.2.1 Levels of PCE in Tissues of Aquatic Species 4-4
4.3 BEHAVIOR IN WATER AND SOIL 4-13
4.4 SUMMARY 4-15
4. 5 REFERENCES 4-16
5. COMPOUND DISTRIBUTION AND RELATED PHARMACOKINETICS 5-1
5.1 HUMAN AND ANIMAL STUDIES 5-1
5.1.1 Absorption 5-1
5.1.2 Distribution 5-4
5.1.3 Metabolism 5-6
5.1.4 Excretion and Elimination 5-13
5.1.5 Estimates of Biological Half-life 5-24
5.1.6 Interaction of PCE with Other Compounds 5-24
5.1. 7 Summary 5-25
5.2 REFERENCES 5-28
IV
003PE2/A 12/1/83
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CONTENTS (continued)
6. TOXIC EFFECTS 6-1
6.1 HUMANS 6-1
6.1.1 Effects on the Liver 6-1
6.1.2 Effects on Kidneys 6-5
6.1.3 Effects on Other Organs/Tissues 6-5
6.1.4 Behavioral and Neurological Effects 6-6
6.2 ANIMALS 6-10
6.2.1 Effects on the Nervous System 6-10
6.2.2 Effects on the Liver and Kidney 6-19
6.2.3 Effects on on the Heart 6-24
6.2.4 Effects on the Skin and Eye 6-24
6.3 ADVERSE EFFECTS OF SECONDARY POLLUTANTS 6-24
6.4 SUMMARY OF ADVERSE HEALTH EFFECTS AND ASSOCIATED LOWEST
OBSERVABLE EFFECT CONCENTRATIONS 6-24
6.4.1 Inhalation Exposure 6-24
6.4.2 Oral Exposure 6-26
6.4.3 Dermal Exposure 6-26
6.5 REFERENCES 6-27
7. TERATOGENICITY, EMBRYOTOXICITY, AND REPRODUCTIVE EFFECTS 7-1
7.1 ANIMAL STUDIES 7-2
7.1.1 Mice 7-2
7.1.2 Rats 7-3
7.1. 3 Rabbits 7-6
7.2 SUMMARY 7-7
7. 3 REFERENCES 7-8
8. MUTAGEN IC ITY 8-1
8.1 GENE MUTATION TESTS 8-1
8.1.1 Bacteria 8-1
8.1.2 Drosophila 8-12
8.2 CHROMOSOME ABERRATION TESTS 8-13
8.2.1 Whole-Mammal Bone Marrow Cells 8-13
8.2.2 Human Peripheral Lymphocytes 8-14
8.2.3 Drosophila 8-15
8. 3 OTHER TESTS INDICATIVE OF DNA DAMAGE 8-15
8. 3.1 DNA Repair 8-15
8.3.2 Mitotic Recombination 8-18
8.4 DNA BINDING STUDIES 8-21
8.5 STUDIES INDICATIVE OF MUTAGENICITY IN GERM CELLS 8-21
8. 6 MUTAGENICITY OF METABOLITES 8-22
8. 7 SUMMARY AND CONCLUSIONS 8-24
8. 8 REFERENCES 8-26
003PE2/A
12/1/83
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CONTENTS (continued)
9. CARCINOGENICITY 9-1
9.1 ANIMAL STUDIES 9-1
9.1.1 National Cancer Institute Bioassay (1977a) 9-1
9.1.2 Dow Chemical Company Inhalation Study (Rampy et al.,
1978) 9-10
9.1.3 Intraperitoneal Administration Study (Theiss et al.
1977) 9-12
9.1.4 Skin Painting Study (Van Duuren et al., 1979) 9-13
9.2 EPIDEMIOLOGIC STUDIES 9-15
9.2.1 Kaplan (1980) 9-15
9.2.2 Blair et al. (1979) 9-18
9.2.3 Katz and Jowett (1981) 9-18
9.2.4 Lin and Kessler (1981) 9-20
9.2.5 Asal (personal communication, 1983) 9-20
9.3 QUANTITATIVE ESTIMATION 9-21
9.3.1 Procedures for the Determination of Unit Risk 9-23
9.3.2 Unit Risk Estimates 9-31
9.3.3 Comparison of Potency with Other Compounds 9-39
9.4 SUMMARY AND CONCLUSIONS 9-39
9.4.1 Qualitative 9-39
9.4.2 Quantitative 9-44
9.4. 3 Conclusions 9-45
9. 5 REFERENCES 9-46
APPENDIX A. Comparison Among Different Extrapolation
Models A-l
APPENDIX B. Time-To-Event Data and Calculations of
Potency B-l
VI
003PE2/A 12/1/83
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LIST OF TABLES
Table Page
3-1 Major U. S. producers of PCE 3-3
3-2 Ambient air mixing ratios of PCE measured at sites around
the world 3-11
4-1 Levels of PCE in tissues of marine organisms, birds, and
mamma Is 4-5
4-2 Accumulation of PCE by dabs 4-9
4-3 Concentration of PCE and trichloroethylene in mollusks and
fish near the Isle of Man 4-12
5-1 Estimated uptake of six individuals exposed to tetrachloro-
ethylene while at rest and after rest/exercise 5-3
5-2 Alcohol and diazepam effects upon tetrachloroethylene (PCE)
in blood and breath concentrations, 5.5-hour exposures 5-26
6-1 Summary of the effects of tetrachloroethylene on animals 6-11
6-2 Toxic dose data 6-17
8-1 Summary of mutagenicity testing of PCE 8-2
8-2 Results of bacterial tests of different purities and
sources of PCE 8-3
9-1 Incidence of hepatocellular carcinomas in B6C3F1 mice fed
PCE 9-3
9-2 Cumulative survival of Sprague-Dawley rats exposed to PCE
for 12 months 9-12
9-3 Pulmonary tumor response to PCE 9-13
9-4 Incidence rate of hepatocellular carcinomas in female
B6C3F1 mice 9-32
9-5 Upper-bound (point) estimates of risk at various dose levels
based on data of hepatocellular carcinomas in female mice
(NCI) 9-36
9-6 Expired PCE unchanged as percentage of recovered radio-
activity 9-37
9-7 Relative carcinogenic potencies among 53 chemicals evaluated
by the Carcinogen Assessment Group as suspect human
carcinogens 9-41
A-l Maximum likelihood estimate of the parameters for each of
the four extrapolation models based on hepatocellular
carcinomas in female mice A-2
B-l Time-to-event data and calculations of potency B-l
VI 1
003PE2/A 12/1/83
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LIST OF FIGURES
Figure Page
3-1 Locations of U.S. PCE production facilities producing more
than 100 mi 11 i on pounds 3-4
4-1 Accumulation and loss of PCE by dabs 4-10
4-2 Relation between flesh and liver concentrations of PCE
in dabs 4-11
5-1 Predicted post-exposure alveolar air concentrations of PCE
at various times against duration of exposure 5-5
5-2 Mean and range of breath concentrations of PCE after exposure
of individuals to single or repeated exposures 5-15
5-3 Mean and range of breath concentrations of five individuals
during post-exposure after five separate exposures to 96,
109, 104, 98 and 99 ppm : 5-16
5-4 PCE in blood and exhaled air following exposure to PCE for
4 hours 5-18
5-5 Trichloroacetic acid in blood following exposure to PCE
for 4 hours 5-21
5-6 Urinary excretion of trichloroacetic acid following exposure
to PCE for 4 hours 5-22
8-1 Dose response curves of Perchlor 200 and Perchlor 230 using
Salmonella tester strains TA100 and TA1535 8-8
8-2 Induction of mitotic recombination by PCE in Saccharomyces
cerevisiae D7 8-19
9-1 Growth curves for male and female mice in the PCE chronic
study (NCI) 9-5
9-2 Survival comparisons of male and female mice in the PCE
chronic study (NCI) 9-6
9-3 Growth curves for male and female rats in the PCE chronic
study (NCI) 9-7
9-4 Survival comparisons of male and female rats in the PCE
chronic study (NCI) 9-8
9-5 Histogram representing the frequency distribution of the
potency indices of 53 suspect carcinogens evaluated
by the Carcinogen Assessment Group 9-40
VI 1 1
003PE2/A 12/1/83
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AUTHORS AND REVIEWERS
The principal authors of this document are:
Chao W. Chen, Carcinogen Assessment Group, U.S. Environmental Protection
Agency, Washington, D.C.
I.W.F. Davidson, Department of Physiology and Pharmacology, Bowman Gray
School of Medicine, Winston-Salem, N.C.
Vicki Vaughan-Oellarco, Reproductive Effects Assessment Group, U.S. Environ-
mental Protection Agency, Washington, D.C.
Herman Gibb, Carcinogen Assessment Group, U.S. Environmental Protection Agency,
Washington, D.C.
Mark Greenberg, Environmental Criteria and Assessment Office, U.S. Environmental
Protection Agency, Research Triangle Park, N.C.
Charalingayya B. Hiremath, Carcinogen Assessment Group, U.S. Environmental
Protection Agency, Washington, D.C.
Jean C. Parker, Office of Solid Waste, U.S. Environmental Protection Agency,
Washington, D.C.
IX
003PE2/A 12/1/83
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The following individuals were asked to review an early draft of this document
and submit comments:
Dr. Joseph Borzelleca
Dept. of Pharmacology
The Medical College of Virginia
Virginia Commonwealth University
Richmond, VA 23298
Dr. Benjamin Van Duuren
Institute of Environmental Medicine
New York University Medical Center
New York, NY 10016
Dr. Herbert Cornish
Dept. of Environmental and Industrial Health
University of Michigan
Ypsilanti, MI 48197
Dr. I. W. F. Davidson
Dept. of Physiology/Pharmacology
The Bowman Gray School of Medicine
Winston-Salem, NC 27103
Dr. Lawrence Fishbein
National Center for Toxicological Research
Jefferson, AR 72079
Dr. John G. Keller
P. 0. Box 10763
Research Triangle Park, NC 27709
Dr. John L. Laseter
Director, Environmental Affairs, Inc.
New Orleans, LA 70122
Al1 Members of the
Interagency Regulatory Liaison Group
Subcommittee on Organic Solvents
003PE2/A 12/1/83
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Participating Members of The Carcinogen Assessment Group
Roy E. Albert, M.D. (Chairman)
Elizabeth L. Anderson, Ph.D.
Larry D. Anderson, Ph.D.
Steven Bayard, Ph.D.
David L. Bayliss, M.S.
Chao W. Chen, Ph.D.
Margaret M. L. Chu, Ph.D.
Bernard H. Haberman, D.V.M., M.S.
Charalingayya B. Hiremath, Ph.D.
Robert E. McGaughy, Ph.D.
Dharm W. Singh, D.V.M., Ph.D.
Todd W. Thorslund, Sc.D.
Participating Members of the Reproductive Effects Assessment Group
Peter E. Voytek, Ph.D. (Director)
John R. Fowle III, Ph.D.
Carol Sakai, Ph.D.
Ernest Jackson, M.D.
K.S. Lavappa, Ph.D.
Sheila Rosenthal, Ph.D.
Vicki Vaughan-Dellarco, Ph.D.
Members of the Agency Work Group on Solvents
Elizabeth L. Anderson
Charles H. Ris
Jean Parker
Mark Greenberg
Cynthia Sonich
Steve Lutkenhoff
James A. Stewart
Paul Price
William Lappenbush
Hugh Spitzer
David R. Patrick
Lois Jacob
Arnold Edelman
Josephine Brecher
Mike Ruggiero
Jan Jablonski
Charles Delos
Richard Johnson
Priscilla Holtzclaw
Office of Health and Environmental Assessment
Office of Health and Environmental Assessment
Office of Health and Environmental Assessment
Office of Health and Environmental Assessment
Office of Health and Environmental Assessment
Office of Health and Environmental Assessment
Office of Toxic Substances
Office of Toxic Substances
Office of Drinking Water
Consumer Product Safety Commission
Office of Air Quality Planning and Standards
Office of General Enforcement
Office of Toxic Integration
Office of Water Regulations and Standards
Office of Water Regulations and Standards
Office of Solid Waste
Office Water Regulations and Standards
Office of Pesticide Programs
Office of Emergency and Remedial Response
003PE2/A
XI
12/7/83
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The following individuals attended a review workshop to discuss draft EPA
documents on organic compounds which included an early draft of this
document:
Dr. Mildred Christian
Argus Laboratories
Perkasie, PA 18944
Dr. Rudolf Jaeger
Institute of Environmental Medicine
New York, NY 10016
Dr. Benjamin Van Duuren
Institute of Environmental Medicine
New York University Medical Center
New York, NY 10016
Dr. Herbert Cornish
School of Public Health
University of Michigan
Ann Arbor, MI 48197
Dr. I. W. F. Davidson
Dept. of Physiology/Pharmacology
The Bowman Gray School of Medicine
Winston-Sal em, NC' 27103
Dr. John Egle
Dept. of Pharmacology
Virginia Commonwealth University
Richmond, VA 23298
Dr. John G. Keller
P. 0. Box 10763
Research Triangle Park, NC 27709
Dr. Norman Trieff
Dept. of Preventive Medicine
University of Texas Medical Branch
Galveston, TX 77550
Dr. Thomas Haley
National Center for Toxicology Research
Jefferson, AK 72079
Dr. James Withey
Food Directorate
Bureau of Food Chemistry
Ottawa, Canada
xi i
003PE2/A . 12/1/83
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1. EXECUTIVE SUMMARY
Tetrachloroethylene (PCE) is a moderately volatile chlorinated hydrocarbon
which has important commercial applications in the dry cleaning of fabrics and
in the degreasing of fabricated metal parts. It is estimated that approximately
300,000 metric tons were produced in the United States in 1980. Approximately
90 percent of production is estimated to be released eventually to the atmos-
phere. Because PCE is relatively insoluble in water (150 mg/1) and has a
vapor pressure of 0.19 pascals at 20°C, PCE in natural waters would be con-
veyed to the atmosphere rapidly, through evaporation. There are no known or
expected natural sources of emissions.
PCE has been detected in the ambient (natural environment) air of a variety
of urban and nonurban areas of the United States and other regions of the
world. Levels can range from trace amounts in rural areas to as much as 10 parts
per billion (ppb) or 0.068 mg/m3 in some large urban centers. The global
average background level is estimated at about 25 parts per trillion (ppt) or
0.175 x 10-3 mg/m3. It is detected less frequently in water, because it is not
appreciably soluble; but PCE has been monitored in some surface and drinking
waters, generally at levels of 1 ppb or less. In certain instances involving
contamination of groundwater, much higher levels have been reported. Although
there is very limited information on the behavior of PCE in soil, PCE can be
expected to leach through soils of low (< 0.1 percent) organic carbon content.
The amount of PCE adsorbed to soils is dependent on the partition coefficient,
the organic carbon content, and the concentration of PCE in the liquid phase.
In the troposphere, a region of the atmosphere extending to between 8
and 16 kilometers above the earth's surface, PCE undergoes photochemical deg-
radation to the extent that its estimated lifetime is appreciably less than
1 year. Little PCE is expected to be conveyed to the stratosphere. Recent
studies have shown that, in real atmospheres, neither atomic chlorine- nor
hydroxy radical-induced photooxidation of PCE generates substantial concen-
trations of ozone or other oxidants; thus, PCE is not believed to be a signi-
ficant factor in production of photochemically-induced pollution often experi-
enced near large urban centers. Because of the reduced solar flux in winter
and seasonal variations in hydroxy radical concentration, PCE levels in ambient
air are expected to be higher in winter than in summer. On a daily basis, PCE
levels fluctuate considerably.
003PE4/F 1-1 11/22/83
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Inhalation is the principal route by which PCE enters the body. Ingestion
of drinking water contaminated by PCE is a secondary source of exposure. During
inhalation, PCE is absorbed by the blood and distributed throughout the body.
Controlled studies with human volunteers (at 100 ppm) have shown that pulmonary
steady-state conditions are reached between 3 and 4 hours. Physical activity
increases uptake. At levels approaching those found or expected in ambient
air, the time required to reach steady-state conditions may be considerably
longer. Once body equilibrium or steady-state has been reached, no further
uptake is possible. There is evidence that PCE will partition selectively
into lipid-rich tissues with chronic or long-term exposure to even low
ambient air concentrations until steady-state is attained. Because of its
lipophilic nature, PCE is expected to cross membrane barriers in the body.
Most inhaled PCE is excreted via the lungs in unchanged form. Both con-
trolled and occupational exposures of humans indicate that the principal
urinary excretion product of PCE metabolism is trichloroacetic acid. Con-
trolled studies with humans have demonstrated that PCE metabolism (urinary
trichloroacetic acid) represents 2 percent or less of the amount of PCE ab-
sorbed. Studies in rats and mice suggest that metabolism of PCE is a saturable
process. In humans, saturation would not be expected until exposure levels
approximate 100 ppm (678 mg/m3).
Toxicity testing in experimental animals, coupled with limited human data
derived principally from overexposure situations, suggests that long-term expo-
sure of humans to environmental levels of PCE is not likely to present a serious
health concern.
Decrements in task performance and coordination are the first gross signs
of central nervous system (CNS) depression and behavioral alterations observed
in controlled human studies in which individuals were exposed to about 100 ppm
(678 mg/m3) for up to 7 hours. More sensitive tests, however, would have to
be performed to determine if PCE affects the nervous system at even lower
concentrations.
Evidence in rodent species suggests that PCE has the potential to cause
liver damage with acute or prolonged exposure at levels that, in humans, would
cause only slight CNS depression. However, there are insufficient data to
estimate the lowest levels of PCE that are associated with adverse effects
upon the liver in humans.
The lowest observed-adverse-effect level (LOAEL), based on CNS dysfunc-
tion, is about 100 ppm (678 mg/m3). However, the LOAEL may not be sufficiently
003PE4/F 1-2 11/22/83
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protective of human health when one considers that intermittent or prolonged
exposure of animals to PCE has been observed to result in liver and kidney
damage at levels exceeding 200 ppm (1,356 mg/m3). It should be noted that
liver damage in humans is generally associated with short-term exposures
greatly in excess of 100 ppm (678 mg/m3). The LOAEL is defined as the lowest
exposure level in a study or group of studies which produces statistically
significant increases in frequency or severity of adverse effects between the
exposed population and its appropriate control.
The mammalian animal tests performed to date do not indicate any signifi-
cant teratogenic potential of PCE in the species tested. On this basis, there
is no evidence to suggest that the conceptus is uniquely susceptible to the
effects of PCE. The anatomical effects observed reflect delayed development
and can be considered reversible. It is important to note, however, that the
reversible nature of an embryonic/fetal effect in one species might, in another
species, be manifested in a more serious and irreversible manner. At the
current time, the teratogenic potential of PCE for humans must be considered
unknown.
Tetrachloroethylene has been evaluated for its ability to cause gene
mutation, chromosomal aberrations, unscheduled DMA synthesis, and mitotic
recombination. These tests were conducted using bacteria, Drosophila, yeast,
cultured mammalian cells, whole mammal systems, and cytogenetic analyses of
exposed humans. Certain technical and commercial samples of PCE elicited
increased responses in the Ames bacterial test, a yeast mitotic recombination
assay, a host-mediated assay using Salmonella, and DNA repair tests. Exogenous
metabolic activation was not required for detection of these increased effects.
In general, the responses were weak and observed at high concentrations that
were cytotoxic; dose-dependent relationships were not established. The positive
findings may be the result of mutagenic contaminants and/or added stabilizers.
There have been several other tests of commercial and technical samples of PCE
which have been reported as negative. The epoxide of PCE, which is thought to
be the active biological intermediate, was found to be positive in bacterial,
studies.
Tetrachloroethylene has been demonstrated to induce malignant tumors of
the liver in both male and female mice of the B6C3F1 strain. This constitutes
limited evidence that PCE may be carcinogenic in humans. It should be recog-
nized that a substantial part of the scientific community believes that the
003PE4/F 1-3 11/22/83
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mouse liver overreacts to chlorinated organic compounds, in contrast to that
of the rat, and that the induction of liver cancer in the mouse represents
only a promoting action for spontaneous liver tumors which normally occur with
substantial incidence.
According to the criteria of the International Agency for Research on
Cancer (IARC), the data supporting the carcinogenicity of PCE must be classi-
fied as "limited." Because existing human epidemiological data for PCE is
inconclusive, its overall IARC ranking should be Group 3, corresponding to the
conservative scientific view that PCE is probably carcinogenic in humans.
003PE4/F 1-4 11/22/83
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Replacement Page 1-4 for EPA-600/8-82-005B, "Health Assessment Document
for Tetrachloroethylene (Perchloroethylene)
mouse liver overreacts to chlorinated organic compounds, in contrast to that
of the rat, and that the Induction of liver cancer in the mouse represents
only a promoting action for spontaneous liver tumors which normally occur with
substantial Incidence.
According to the criteria of the International Agency for Research on
Cancer (IARC), the data supporting the carcinogenlcity of PCE must be classi-
fied as "limited." Because existing human epidemiological data for PCE is
::^^^^^x•^:^•^xw.^x.^x•^x««v:::^::
inconclusive, its overall IARC ranking should be Group 3,;!meaning that there :•$•:
*&<<<<<<'X<<
-------
2. INTRODUCTION
Tetrachloroethylene (PCE) is one member of a family of unsaturated
chlorinated aliphatic compounds. Other common names/acronyms are perchloro-
ethylene, Perk, PER, and PERC. Its synonyms include carbon dichloride,
tetrachloroethene, and 1,1,2,2-tetrachloroethylene.
Tetrachloroethylene, though a water and solid waste contaminant, is pri-
marily of interest in ambient air exposure situations. It is released into
ambient air as a result of evaporative losses during production, storage,
and/or use. It is not known to be generated from natural sources. It has
negligible photochemical reactivity in the troposphere and is removed by
scavenging mechanisms, principally via hydroxyl radicals.
The scientific data base is limited with reference to the effects of PCE
on humans. Effects on humans have generally been ascertained from studies
involving individuals occupationally or accidentally exposed. During such
exposures, the concentrations associated with adverse effects on human health
were either unknown or far in excess of concentrations found or expected in
ambient air. Controlled PCE exposure studies have been directed toward eluci-
dating the effects on the central nervous system, effects on clinical chemistries,
and pharmacokinetic parameters of PCE exposure.
Since epidemiological studies have not been able to assess adequately the
overall impact of PCE on human health, it has been necessary to rely greatly
on animal studies to derive indications of potential harmful effects. Although
animal data cannot always be extrapolated to humans, indications of probable
or likely effects among animal species increase confidence that similar effects
may occur in humans.
This document is intended to provide an evaluation of the scientific data
base concerning PCE. It is believed that the literature has been comprehen-
sively reviewed and critically evaluated through September 1983. The publica-
tions cited in this document represent a majority of the known scientific
references to PCE. Reports which had little or no bearing upon the issues
discussed were not cited.
003PE1/B 2-1 11/22/83
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3. GENERAL BACKGROUND INFORMATION
3.1 PHYSICAL AND CHEMICAL PROPERTIES
Tetrachloroethylene, also called PCE (1,1,2,2-tetrachloroethylene or per-
chloroethylene), is a colorless, heavy liquid with a chloroform-like odor. It
is used as a solvent for organic substances and is commercially important as a
solvent in the dry cleaning of fabrics and in the degreasing of metals. It
has a molecular weight of 165.85 and is relatively insoluble in water (150 mg/1)
(Handbook of Chemistry and Physics, 1976; Chemical and Process Technology Ency-
clopedia, 1974). Its CAS registry number is 127184. In air, at standard tem-
perature and pressure, 1 part per million (ppm) is equivalent to 6.78 mg/m3.
Tetrachloroethylene has negligible photochemical reactivity (Dimitriades
et al., 1983) and, in the troposphere, is decomposed via free radical mechanisms.
When in contact with water for prolonged periods, PCE slowly decomposes to yield
trichloroacetic and hydrochloric acids. Upon prolonged storage in light, it
was reported to decompose slowly to trichloroacetyl chloride and phosgene by
autooxidation (Hardie, 1966). At 700°C, it decomposes, when in contact with
activated charcoal, to hexachloroethane and hexachlorobenzene (Gonikberg, 1956).
Tetrachloroethylene has a boiling point of 121.1°C at 760 mm Hg and a vapor pres-
sure of 14 torr at 20°C. MacKay et al. (1982) have calculated a vapor pressure
of 19 torr at 25°C.
The chemical reactivity of PCE has been discussed by Bonse and Henschler
(1976) in terms of the electron-inductive effect of the chlorine atoms, which
reduce electron density about the ethylene bond. This effect, in combination
with a steric protective effect afforded by the chlorine atoms, provides in-
creased stability against electrophilic attack. This is exemplified in the
reaction of PCE with ozone. Compared to ethylene and less-substituted chlorina-
tion hydrocarbons, PCE has a low rate of reaction (Williamson and Cvetanovik,
1968).
3.2 PRODUCTION
Tetrachloroethylene may be produced by several processes:
1. Chlorination of trichloroethylene:
CHC1 = CC12 + C12 |j^ CHCI2 CC13
003PE4/B 3-1 11/22/83
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2CHC12CC13 + Ca(OH)2 2iLi C12C = CC12 + CaCl2 + 2H20
2. Dehydrochlorination of S-tetrachloroethane:
CHC12-CHC12 + C12 > CC12 = CC12 + 2HC1
3. Oxygenation of S-tetrachloroethane:
2CHClaCHCl2 + 02 > 2CC12 = CC12 + 2H20
4. Chlorination of acetylene:
?nn°r
CC12 = CC12 + C12 -^-^ CC13-CC13
ru - ru * o rn rn 20Q-4000C 4CC12 = CC12 + 2HC1
LH _ LH + J LLI3LLI3 catalyst
5. Chlorination of hydrocarbons:
C3H8 + 8 C12 » CC12 = CC12 + CC14 + 8HC1
(propane)
2CC14 » CC12 = CC12 + 2C12
6. Oxychlorination of 1,2-dichloroethane:
2C2H4C12 + 5C12 > C2H2C14 + C2HC15 + 5HC1
C2H2C14 + C2HC15 » C2HC13 + 2HC1 + CC12 = CC12
7HC1 + 1.75 02 » 3.5 H20 + 3.5 C12
2C2H4C12 + 1.5 C12 + 1.75 02 » C2HC13 + CC12 = CC12 + 3.5 H20
The majority of PCE producted in the United States is derived from the
oxychlorination of 1,2-dichloroethane or via Chlorination of hydrocarbons
(Lowenheim and Moran, 1975).
In 1980, 347,859 metric tons of PCE were produced (U.S. International
Trade Commission, 1980). The major producers and production capacities are
shown in Table 3-1. Locations of U.S. production facilities are shown in Fig-
ure 3-1.
Imports of PCE may be sizeable, although they are partially offset by ex-
ports. Most of the PCE imported is produced in Belgium, Italy, France, and
Canada (Chemical Marketing Reporter, 1979a).
003PE4/B 3-2 ' 11/22/83
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TABLE 3-1. MAJOR U.S. PRODUCERS OF PCE3
Organization Yearly capacity, MTa
Dow Chemical 320
PPG 200
Vulcan 200
Diamond Shamrock 165
Ethyl Corporation
E. I. du Pont de Nemours 160
a ^
Adapted from Chemical Economics Handbook (SRI, 1982). MT = Metric tons.
Recently announced withdrawal from production (Halogenated Solvents
Industry Alliance, 1983)
C0utput for captive use only.
3.3 USE
Tetrachloroethylene has the following uses (Gosselin et al., 1976; Fishbein,
1977): (1) dry cleaning solvent; (2) textile scouring solvent; (3) dried vege-
table fumigant; (4) rug and upholstery cleaner; (5) stain, spot, lipstick, and
rust remover; (6) paint remover; (7) printing ink ingredient; (8) heat transfer
media ingredient; (9) chemical intermediate in the production of other organic
compounds; and (10) metal degreaser.
Testimony provided by the International Trade Commission to the Secretary
of the Treasury in March 1979 stated that dry cleaning consumes approximately
75 percent of U.S. PCE production and imports (Chemical Marketing Reporter,
1979a). About 72 percent of commercial dry cleaning plants are estimated to
use PCE (U.S. EPA, 1979).
3.4 EMISSIONS
Emissions of PCE arise during its production, from its use as a chemical
intermediate in industrial processes, from storage containers, during disposal,
and from its use as a solvent. Because emissions are almost exclusively to
the atmosphere, the information presented in this section focuses on air.
Few data are available concerning discharges to water. The section dealing
003PE4/B 3-3 11/22/83
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100 mllliv»n pounds
Figure 3-1. Locations of U.S. PCE production facilities producing more
than 100 million pounds.
Source: Fuller, 1976.
003PE4/B
3-4
11/22/83
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with monitoring indicates that discharges to water probably occur. Emissions
estimates reflect a diversity of sources throughout the country. Dry cleaning
operations are located primarily in urban areas. Approximately 26,000 estab-
lishments are estimated to exist, according to Bureau of Census data (U.S. EPA,
1979).
In 1977, global emissions were estimated at 570,000 ± 285,000 metric tons
(Singh et al., 1979). It was also estimated that emissions accounted for
approximately 90 percent of the amount of PCE produced (300,000 metric tons)
in the United States.
3.5 ENVIRONMENTAL FATE AND TRANSPORT
The potential for ambient air and water mixing ratios of PCE to pose a
hazard to human health is influenced by many processes. Such factors include
transformation into secondary pollutants of concern and degradation rates in
air and water.
3.5.1 Ambient Air
3.5.1.1 Tropospheric Reactivity-- Reaction with the hydroxyl radical (-OH) is
the principal process by which many organic compounds, including PCE, are scav-
enged from the troposphere. Hydroxyl radicals are produced when 0$ is irradi-
ated, resulting in excited atomic oxygen, which then reacts with water vapor.
The tropospheric lifetime of a compound is related to the -OH mixing ratio
according to the expression:
_
where k is the rate constant of reaction.
Singh et al. (1979, 1981) calculated a tropospheric residence of PCE of
about 68 days. This calculation was based on an average 24-hour -OH abundance
of 10^ molecules cm •* in the boundary layer of a polluted atmosphere. Justi-
fication for this -OH mixing ratio stems from the field studies of Calvert (1976)
and from Singh and coworkers (1979a). Because this -OH mixing ratio is more
typical of summer months, Singh et al. (1981) suggested that a seasonally ad-
justed mixing ratio would result in a longer chemical residence time. If a
seasonally-averaged -OH mixing ratio of 4 x 105 molecules cm 3 (a level supported
003PE4/B 3-5 11/22/83
-------
by the field measurements of Campbell et al. , 1979) and a weighted global aver-
age temperature (265°K), an average residence time of PCE would be calculated
to be 292 days or 0.8 years (Singh et al., 1979).
Estimations of a residence time of PCE of one year or less have been
reported by a number of investigators (Altshuller, 1980; Singh et al., 1978a;
Singh, 1977; Crutzen et al., 1978; Lillian et al., 1975; Yung et al., 1975;
Pearson and McConnell, 1975).
Dimitriades et al. (1983) calculated a very low tropospheric reactivity
for PCE based on observations that ambient levels of PCE are constant. Atmos-
pheric consumption of PCE is 0.02 percent per daylight hour.
Higher levels of -OH have been reported for the southern hemisphere com-
pared to those in the northern hemisphere (Singh, 1978). This gradient probably
is due to the fact that carbon monoxide levels are much higher in the northern
hemisphere, thus reducing -OH levels (Singh, 1978). Measurements of PCE and
other reactive halocarbons indicate that mixing ratios are higher in the nor-
thern hemisphere where the -OH mixing ratio is low and where most of the PCE
is released (Singh et al., 1978b).
Chamber studies indicate that PCE, on irradiation in the presence of other
atmospheric constituents, can be transformed into secondary products. This
area of study has been recently reviewed and further explored by Dimitriades
et al. (1983). These investigators have confirmed that PCE, under smog chamber
conditions, reacts to form Og and ozone precursors by means of a Cl-initiated
photooxidation mechanism. However, such photooxidation is not expected to occur
in the real atmosphere at a rate high enough for substantial 03 production.
It is the authors' contention that Cl atoms are effectively scavenged by the
hydrocarbons normally present in the atmosphere; thus, PCE was judged to con-
tribute less to 03 production than equal concentrations of ethane. Ethane is
regarded by the authors to be a boundary species separating the reactive vola-
tile organics from the unreactive ones.
Studies on PCE reactions with Og, 0, and «OH have indicated that rate
constants are lower than with Cl (Dimitriades et al., 1983).
Gay et al. (1976) had determined that trichloroacetyl chloride was a
photooxidation product of PCE in smog chamber studies. Dimitriades et al.
(1983) found that, on irradiation, the only product observed was CO.
Phosgene was not detected. When 2 ppm (13.6 mg/nr*) PCE and 20 ppb trichloro-
acetyl chloride were irradiated together, the phosgene level reached 0.1 ppm.
003PE4/B 3-6 11/22/83
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Phosgene production from the photochemical oxidation of PCE in the presence
of other substances has been reported by others (Lillian et al., 1975; Gay et
al., 1976). The extent to which phosgene may be formed in real atmospheres,
based on smog chamber results, would also be expected to be minimal.
3.5.1.2 Tropospheric Removal Mechanisms for PCE--The reaction sequence by
which PCE may be scavenged from the troposphere is as follows (Graedel , 1978):
C2C14 + HO- - > HOC(C1)2C(C1)2-
HOC(C1)2C(C1)2- + Og - > HOC(C1)2C(C1)202-
HOC(C1)2C(C1)202- - oxygen - » HOC(C1)2C(C1)20-
abstraction
HOC(C1)2C(C1)20- - > HOC(C1)2- + COClj,
HOC(C1)2- + 02 - » COC12 + H02-
Howard (1976) suggested that the reaction path for the atmospheric oxida
tion of PCE may follow the scheme below, leading to the production of oxalyl
chloride:
C2C14 + -OH - » C2C12OH-
- > CC12CC10H- + Cl- (3-3)
CCl2CCljjOH- + 02 - »• 02CC12CC12OH
02CC12CC12OH + NO - » COC1CC12OH- + N02 + Cl •
•OH + COC1CC12OH - > COC1COC1 + H20 + Cl •
Compared to other ethylene compounds studies, Howard (1976) reported that PCE
exhibits unusually low reactivity toward hydroxyl radicals.
003PE4/B 3-7 11/22/83
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Snelson et al. (1978) suggested that trichloroacetyl chloride and phos-
gene would hydrolyze to the corresponding chloroacetic acids and hydrogen
chloride via homogeneous gas phase hydrolysis. The acids then would pre-
sumably be washed out of the atmosphere.
The environmental significance of the production of phosgene from PCE has
been discussed by Singh and coworkers (Singh et al., 1975; Singh, 1976). As
PCE emissions are likely to be higher in urban areas, the reactivity of this
halocarbon may result in concentrations of phosgene in the low ppb range during
adverse meteorological conditions in and around urban centers (Singh, 1976).
It should be noted that overt adverse health damage would not be expected at
these phosgene levels. Considering the smog chamber results of Dimitriades et
al. (1983), in which PCE was found to have negligible reactivity, it appears
unlikely that phosgene would be produced at other than trace levels. The
results of Singh and coworkers indicate that negligible tropospheric loss via
gas phase hydrolysis would be expected, because phosgene is very stable in the
gas phase. The two important sinks cited were heterogeneous decomposition and
slow liquid phase hydrolysis. Singh (1976) concluded that phosgene is removed
slowly from the atmosphere. Rainfall appears to lower atmospheric levels of
phosgene (Singh et al., 1977a). On the other hand, phosgene has been reported
to hydrolyze to C02 and HC1 rapidly in liquid water. Manogue (1958) reported
a half-life of 0.1 second at 25°C. Thus, it is not expected to persist in the
troposphere because of rainout and hydrolysis in aqueous aerosols.
The observed diurnal variations in PCE levels suggest that PCE has higher
mixing ratios in the morning and evening hours than at other times (Lillian et
al., 1975; Singh et al., 1977a). Ohta et al. (1977) reported that mixing ratios
tended to be highest on cloudy days and lowest on rainy days.
Solar flux is a major factor in the rate at which PCE is removed from the
atmosphere. Singh et al. (1977a) suggested that the reduced solar flux in winter
months would permit a much longer transport of PCE because of reduced reactivity.
The effect of solar flux was calculated by Altshuller (1980) who estimated that
a 1 percent consumption of PCE by reaction with -OH would take 14 days during
the month of January as opposed to one day in July.
While the studies of Gay et al. (1976) indicate that trichloroacetyl chlo-
ride may be formed through chlorine atom migration in an epoxide intermediate,
evaluation of -OH and oxygen atom rate constants indicate that less than 1 per-
cent of PCE in ambient air reacts with atomic oxygen and, of the activated epox-
ides formed, only a small percentage undergo rearrangement (Graedel, 1978).
003PE4/B . 3-8 11/22/83
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3.5.2 Water
Jensen and Rosenberg (1975) investigated the degradability of PCE (~ 0.1
to 1.0 ppm) in both open and closed systems, with sea water and fresh water.
In open aquaria, with 20 1 of sea water, PCE levels decreased 50 percent within
200 hours (daylight). In a closed system, levels decreased approximately 25 per-
cent over the same interval (daylight). It was reported that PCE levels in
boiled, deionized water in a closed system did not exhibit any significant de-
crease after 8 days. Analysis was by headspace collection, followed by gas
chromatography-electron capture detection (GC-ECD) quantification. Accuracy
and limits of detection were not reported. Variation in detector sensitivity
was checked daily by an injection of PCE in hexane.
Dilling et al. (1975) reported that PCE in water slowly decomposes to form
trichloroacetic and hydrochloric acids. The evaporation rate was determined
by dissolving 1 ppm (w/w) PCE in 200 ml of water. Solution temperature was
approximately 25°C. The solution was stirred magnetically in a sealed system.
Quantification was made by mass spectroscopy. The evaporation rate of PCE,
determined from measurements over a 2-week period, was characterized by a 50
percent decrease in the initial mixing ratio in 24 to 28 minutes. The stir-
ring speed had a marked effect on the evaporation rate. With no stirring ex-
cept for 15 seconds every 5 minutes, the time required for 50 percent depletion
ranged from 72 to 90 minutes. The evaporative half-life was 27 ± 3 minutes.
Addition of dry, granular bentonite clay (500 ppm) appeared to increase the
rate of disappearance by 33 percent at 20 minutes. However, when the clay was
allowed to remain in contact with purified water for several days and then added
to the solution, there was no change in the rate compared with control. In
closed system investigations, Dilling et al. (1975) used dry, powdered dolomi-
tic limestone, bentonite, and peat moss to determine the adsorption rate for
PCE. With 500 ppm bentonite, there was a 22 percent absorption after 30 minutes.
There was no further absorption. Addition of limestone resulted in a 50 percent
depletion in 20 ± 2 minutes. Addition of silica sand had no effect on the dis-
appearance rate. When 500 ppm peat moss was added, up to 0.4 ppm PCE was ab-
sorbed after 10 minutes. At longer times, no further absorption was observed.
It was concluded that evaporation probably is the major pathway by which PCE
i s lost from water.
In reactivity studies, Dilling et al. (1975) found that sunlight had the
greatest effect on the rate of PCE disappearance. The PCE level in water was
003PE4/B 3-9 11/22/83
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6 |jM. Over a 12-month period, in which PCE solutions were stored in the dark
as well as in the light, PCE levels decreased from 1 ppm (0 time), to 0.63 ppm
(6 months), to 0.35 ppm (12 months) in samples stored in the dark. In the light-
exposed solution, the level decreased to 0.52 ppm (6 months), and 0.24 ppm (12
months).
Schwarzenbach et al. (1979), in measurements of PCE levels in Lake Zurich,
reported findings compatible with those of Dilling et al. (1975) in that evapora-
tion is the dominant elimination process from surface waters. The annual release
to the atmosphere was estimated by applying a steady-state mass balance model.
Based on vertical concentration profiles from the lake, about 240 kg PCE was
released from the central basin annually.
3.6 LEVELS OF EXPOSURE
3.6.1. Mixing Ratios
3.6.1.1 Ambient Aii—A wide variety of halogenated aliphatic hydrocarbons,
including PCE, have been detected in ambient air. Ambient measurements of PCE
have been conducted in both the United States and other areas of the world.
These determinations provide a basis for assessing the levels to which human
populations may be exposed.
Measured ambient air concentrations differ widely and undoubtedly reflect
the influences of a variety of factors, e.g., meteorological conditions, tropo-
spheric reactivity, diurnal variations, sampling times, and source emissions.
Table 3-2 provides summary information regarding background and urban con-
centrations of PCE. It should be noted that, in general, these values reflect
short sampling times.
Measurements of ground-level samples by Singh et al. (1978b), in both the
northern and southern hemispheres, gave average background levels of 0.040 ±
0.012 ppb (2.7 x 10"4 ± 0.08 x 10"4 mg/m3) and 0.012 ± 0.003 ppb (0.081 x lo"3
±0.02 x 10 3 mg/m3), respectively. Globally, the average background level of
PCE was 0.026 ± 0.007.7 ppb (1.7 x lo"4 ± 0.47 x 10"4 mg/m3); the coefficient
of variation was 27 percent. The average urban level of PCE was found to be
about 30 times the background level.
Evidence for considerable variability in ambient air levels of PCE was
shown by Lillian et al. (1975). The authors attributed the variability of PCE
to its tropospheric reactivity (reaction with hydroxyl radicals).
003PE4/B 3-10 11/22/83
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TABLE 3-2. AMBIENT AIR MIXING RATIOS OF PCE MEASURED AT SITES AROUND THE WORLD
Location
Alabama3
Birmingham
Arizona
Grand Canyon
Phoenix
California
San Bernardino Mtns.
oo Badger Pass
i
i— • Point Arena
Stanford Hills
Point Reyes
Dominguez
El Cajon
La Jolla
Los Angeles
Menlo Park
Mill Valley
Mt. Cuyamaca
Date of Reported Concentration, ppra (mg/mj)
Measurement Max i urn Minimum Average
April 12-22, 1977 0.008 0 0.001 ± 0.003
Nov. 28-Oec. 5, 1977 0 0 0
Apr. 23-May 6, 1979b 3.696 (0.025) 0.129 0.9938 ± 0.7155
(0.0008)
Fall. 1972^ 0.09 (6.1 x 10"4)
May 12-16, 1976 0.03 (2 x 10~4)
1976 0.03 (2 x 10~4)
Nov. 24-30, 1975 0.04 (2.6 x 10"4)
Dec. 2-12. 1975 0.043 (2.9 x lo"4)
May 14, 1976 2.9
April 9, 1975 0.31
Apr 9, 1974-Jan 6, 1976 2.3 0 0.53 ± 0.63
Sept. 22. 1972- 2.2 0.067 1.1 ± 0.45
April 19, 1979
Nov. 24-30, 1975 0.20 ± 0.21
Jan. 1-27, 1977 0.065 ± 0.075
Mar. 15, 1975 0.22
Reference
Pellizzari. 1979
Pellizzari, 1979
Singh et al., 1981
(0.0067 ± 0.0048)
Simmonds et al. , 1974
Singh et al. , 1977a
Singh et al. , 1978a
Singh et al. , 1977b
Singh et al. , 1977b
Pellizzari. 1977
Su and Goldberg, 1976
Su and Goldberg, 1976
Simmonds et al., 1974;
Singh, 1976;
Singh et al. , 1977a;
Su and Goldberg, 1976
Singh et al. , 1977a
Singh et al. , 1979
Su and Goldberg, 1976
-------
TABLE 3-2. (continued)
ro
Location
California
Oakland
Palm Springs
Riverside
San Jose
Santa Monica
Upland
Colorado
Denver
Delaware
Delaware City
Louisiana
Baton Rouge, Gel soar,
and Plaquemine
Maryland
Baltimore
Michigan
Detroit
New Jersey
Bayonne
Date of Reported Concentration, ppm (mg/m3)
Measurement Max i urn Minimum Average
June 30-July 8, 1979 0.64 0.12 0.31 * 0.17
May 5-11. 1976 1.1 0.12 0.28 ± 0.084
April 25, 1977- 0.98 0.37 0.49 ± 0.13
July 12, 1980
Aug. 21-27, 1978 . 1.1 ± 0.036
April 6, 1974 2.3
Aug. 13-Sept. 23, 1977 1.1 0.01 0.19 ± 0.35
Jun 16-26, 1980 0.47 0.24 0.39 ± 0.077
July 8-10. 1974 0.51 (0.0034) <0.02 (<0.0001) 0.24 (0.0016)
Fall. 1978 0.18 (0.001) .001 (0.007 x lo"3) 0.017 (0.118 x 10"3)
July 11-12, 1974 0.29 (0.0019) <0.02 (<0.0001) 0.18 (0.0012)
Oct 27-Nov 5. 1978b 2.16 (0.004) <0.1 (<0.001) 0.35 (0.002)
March, 1973-Dec. , 1973 8.2 (0.0055) 0.30 (0.0020) 1.63 (0.0110)
Reference
Singh et al. , 1979;
Singh et al. , 1981
Singh et al. , 1978a
Singh et al. , 1979;
Singh et al. . 1979
1980
Su and Goldberg, 1976
Pellizzarl, 1979
Singh et al. , 1980
Lillian et al. . 1975
Pellizzari et al. . 1979b
Lillian et al., 1975
Evans et al . , 1979
Lillian et al. , 1975
-------
TABLE 3-2. (continued)
Location
Date of
Measurement
Reported Concentration, ppra (mg/mj)
Maximum Minimum
Average
Reference
New Jersey
New Brunswick
New Brunswick
0.5 (0.003)
0.12 (0.0081)
Seagirt
Sandy Hook
Boundbrook, Rahway,
Edison and Passaic
Batsto3
Bridgeport3
Burlington3
Camden3
Carlstadt3
Edison3
Elizabeth3
Fords3
Middlesex3
Newark3
June 18-19, 1974
July 2, 1974
Sept 18-22, 1978
Feb. 26-Oec. 29. 1979
Sept. 22, 1977
Sept. 19, 1977
April 3-Oct. 24, 1979
Sept. 28-30, 1978
March 24, 1976-
Sept. 24, 1978
Sept. 15, 1978-
Oec. 29, 1979
Mar. 26, 1976-
Sept. 27. 1978
July 23-28, 1978
Mar. 23, 1976-
Dec. 29, 1979
0.88 (0.059)
1.4 (95 x 10"«)
58 (0.394)c
0.53
0.041
30
5.5
8.7
14
5.6
0.21
32
0.10 (0.067)
0.15 (10 x 10~«)
trace
0
0
0
1.1
0.11
0
0
0
0
0.32 (0.0022
30.9 (0.210)
0.034
0.020
0.027
1.8 ± 6.2
3.5 ± 2.3
2.8 ± 3.1
2.0 ± 3.1
2.8 ± 2.7
0.068 ± 0.091
1.3 ± 3.1
Lillian et al., 1976
Lillian and Singh, 1974
Lillian et al., 1975
Lillian et al., 1975
Pellizari et al., 1979
Bozzelli et al., 1980
Pellizzari and Bunch, 1979
Pellizzari and Bunch, 1979
Bozzelli et al., 1980
Pellizzari et al. , 1979
Pellizzari et al., 1979; Pelliz-
zari, 1978; Bunn et al., 1975
Bozzelli et al. , 1980;
Pellizzari, 1979
Pellizzari et al. , 1979;
Pellizzari, 1977
0.090 Bozzelli and Kebbekus, 1979
Bozzelli and Kebbekus, 1979;
Bozzelli et al., 1980;
Pellizzari, 1977
-------
TABLE 3-2. (continued)
Location
New Jersey
Rahway
Rutherford3
3
Somerset
South Amboy
New York
New York City
OJ
i Niagara Falls and
£ Buffalo
Whiteface Htn.
Ohio3
Wilmington
Texas
Houston
Aldine3
Deer Park3
El Pasoa
Freeport3
Houston
Date of
Measurement
Sept. 20-22. 1978
May 1. 19 78- Dec 29, 1979
July 18-26. 1978
Jan. 27-Dec. 29, 1979
June 27-28. 1974
Aug 18-27. 1978°
Fall 1978
Sept. 17. 1974
July 16-26, 1974
Sept 16-25, 1978b
June 22-Oct. 20, 1977
July 29-30, 1976
April 5-May 1, 1978
Aug. 9, 1976
July 27, 1976-
May 24, 1980
Reported Concentration,
Maxium
5.0
9.2
0.068
2.2
9.75 (0.0661)
10.61 (0.0721)
2.0 (0.014)
0.19 (12.8 x 10~4)
4.52 (0.030)
0.037
0.15
0.39
1.3
ppm (oig/in3)
Minimum
2.7
0
0
0
1.0 (0.006)
0.16 (0.001)
0.02 (0.122 x 10"3)
0.02 (0.13 x 10"3)
<0.1 (<0.001)
0
0.01
0.11
0
Average
3.8 ± 1.2
0.89 ± 0.14
0.036 ± 0.26
0.21 ± 0.53
4.5 (0.030)
1.00 (0.006)
1.0 (0.0068)
0.15 ± 0.015
0.11 (0.001)
0.012
0.07
0.15
0.12
0.33
Reference
Bozzelli and Kebbekus, 1979;
Bozzelli et al. . 1980
Bozzelli et al. , 1980
Bozzelli and Kebbekus, 1979
Bozzelli et al. . 1980
Lillian et al. , 1975
Evans et al. , 1979
Pellizzari et al. , 1979b
Lillian et al. , 1975
Lillian et al. , 1975
Evans et al. , 1979
Pellizzari et al. , 1979
Pellizzari et al. , 1979
Pellizzari, 1979
Pellizzari et al., 1979
Pellizzari et al. , 1979;
Singh et al.. 1980
-------
TABLE 3-2. (continued)
Location
Texas
LaPorte3
Pasadena
Utah3
oo Hagna
^ a
-------
Some of the highest air levels of PCE reported have been associated with
waste disposal sites. Pellizzari (1978) reported levels to ranges at sites in
New Jersey that ranged from trace amounts to a maximum of 58 ppb (0.394 mg/m3)
in a 14-minute sampling period. PCE was adsorbed using Tenax cartridges.
Coefficients of variation for most of the recent studies reported by Singh
et al. have been less than 30 percent.
Howie (1980) reported ambient air levels of PCE in the vicinity of laun-
dries to be as high as 32 ppb (0.22 mg/m3). In this study of indoor PCE con-
centrations, outdoor samples provided background data. Measurements were made
by adsorbing PCE onto charcoal filters, followed by desorption with carbon
disulfide and quantification by GC-ECD and GC-MS. Outdoor samples were collec-
ted for 24 hours. Of 124 measured samples, 56 had 24-hour levels of less than
1 ppb. Replicate sample analyses were reported to give an overall precision
of better than 20 percent for both indoor and outdoor samples.
3.6.1.2 Water—Various studies have shown that PCE is found in both natural
and municipal waters. A review by Deinzer et al. (1978) has summarized many
of the findings. Love and Eilers (1982), in their review, reported that halo-
genated solvents such as PCE are seldom detected in concentrations greater
than a few micrograms per liter in surface waters.
3.6.1.2.1 Natural waters. Surface waters, such as rivers and lakes, are the
most important sources of drinking water in the United States. Attempts have
been made to show an epidemiological link between the presence of halogenated
organic compounds in drinking water and cancer (Harris and Epstein, 1976) but
a cause-effect relationship has not been established.
Dowty et al. (1975a,b) detected PCE by GC-MS techniques in untreated Miss-
issippi River water as well as in treated water. An approximate six-fold re-
duction in concentration occurred after sedimentation and chlorination.
Tetrachloroethylene in water from a commercial deionizing charcoal filtering
unit showed a marked increase over the amount found in finished water from
treatment facilities or commercial sources of bottled water. The value of
charcoal filtering to remove organics from water requires further study.
Suffet et al. (1977) reported detection of PCE in river waters supplying
drinking water to Philadelphia, Pennsylvania. The Belmont Water Treatment
Plant, with an average capacity of 78 million gallons per day, obtains influent
from the Schuykill River.
In a study designed to detect pollutants in surface water at different
U.S. sites, Ewing et al. (1977) identified PCE among the pollutants. Detection
003PE4/B 3-16 11/22/83
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limits were not reported. The highest level (45 ug/1) was reported for surface
water in Ashtabula, Ohio, located along the south edge of Lake Erie. In the
vicinity of Philadelphia, 3 ug/1 was detected. Less than 1 ug/1 was reported
along the Ohio River and at the confluence of its tributaries. Similarly,
1 ug/1 or less was reported for all sites sampled in the Great Lakes and at
six sites along the Tennessee River. An exception was Chattanooga where 13
ug/1 was reported. In the greater Chicago area, the highest level found was 5
ug/1. In all samples taken in California, Oregon, and Washington, PCE was
either not detected or was found at a concentration of 1 ug/1 or less. Sam-
pling sites included those in the vicinity of Los Angeles Harbor, Santa Monica
Bay, and San Francisco Bay, at three sites along the Willamette River in Oregon,
and two in the Puget Sound area.
Tetrachloroethylene was among a number of halogenated organics found in
community drinking water supply wells in Nassau County, New York. Because con-
tamination, 16 of these wells were closed by the New York State Health Depart-
ment (Ewing et al., 1977). The maximum detected level of PCE in the contami-
nated wells was 375 ug/1. Since PCE is generally not used as a cesspool
cleaning agent, previous industrial dumping may be the source of contamination.
Pearson and McConnell (1975) found an average PCE concentration of 0.12
ppb in Liverpool Bay sea water; the maximum concentration found was 2.6 ppb.
Sediments from Liverpool Bay were found to contain 4.8 ppb (w/w). No direct
correlation was found between PCE concentration in sediments and in the waters
above. Rainwater collected near an organochlorine manufacturing site was found
to contain 0.15 ppb (w/w) PCE (Pearson and McConnell, 1975); it was not detected
in well waters. Upland waters of two rivers in Wales were found to contain
approximately 0.15 ppb PCE; similar levels of trichloroethylene were found
(Pearson and McConnell, 1975).
Lbchner (1976) found that levels of PCE in Bavarian lake waters ranged
from 0.015 to 3 ppb (0.015 x 10~^ to 2.7 x 10~3 mg/1). European surface
waters were reported to have uniform PCE concentrations ranging from 0.2 x 10 3
to 0.002 mg/1. Analyses of river, canal, and sea water, all containing effluent
from production and user sites in four countries, revealed PCE concentrations
ranging from 0.01 to 46 ppb (0.01 to 46 ug/liter) (Correia et al. , 1977).
3.6.1.2.2 Municipal waters. Bellar et al. (1974) measured the concentration
of PCE in water obtained from sewage treatment plants in several cities. Before
treatment, the average PCE concentration was 6.2 ug/1. The treated water
003PE4/B 3-17 11/22/83
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before chlorination contained 3.9 pg/l PCE. After chlorination, the effluent
contained 4.2 (jg/1 PCE.
Tetrachloroethylene has been detected in the drinking water of a number
of U.S. cities. These include Evansville, Indiana (Keith et al., 1977); Kirk-
wood, Missouri (Keith et al., 1977); New Orleans, Louisiana (Dowty et al.,
1975); Jefferson Parish, Louisiana (Dowty et al., 1975b); Cincinnati, Ohio
(Keith et al., 1977); Miami, Florida (Keith et al., 1977); Grand Forks, North
Dakota (Keith et al., 1977); Lawrence, Kansas (Keith et al., 1977); New York
City (Keith et al., 1977); and Tucson, Arizona (Keith et al., 1977).
Concentrations recorded for the above cities were less than 1 ug/1. An
exception was Jefferson Parish, which had a measured concentration of 5 ppb
(5 ug/1). Keith et al. (1977) did not detect PCE in the drinking water of
Philadelphia. Tetrachloroethylene was found in Evansville tap water from July
1971 to December 1972. The Ohio River Basin, a heavily industralized area, is
upstream from Evansville and serves as a major source of drinking water for
that community.
Dowty et al. (1975b) determined levels of PCE in the drinking water for
New Orleans. Considerable variation in the relative concentrations of the
various halogenated compounds was observed from day to day.
Contamination of drinking water by PCE was recently investigated by Wake-
ham et al. (1980). It was reported that elevated concentrations of PCE were
found in drinking water transported in vinyl-coated asbestos-cement pipes in
areas of the town of Falmouth, Massachusetts. Tetrachloroethylene is used as
a solvent during the application of the vinyl coating to the pipe during manu-
facturing. It was suggested that residual solvent leaches into the water
carried in these pipes.
Using a charcoal trap with flame ionization detection, Wakeham and co-
workers (1980) detected levels ranging from 140 to 18,000 ppb in unflushed
pipes. In other parts of the distribution system, levels were less than 2 ppb.
The authors reported that vinyl-coated asbestos-cement pipe has been used in
parts of the northeastern United States over the past decade in response to
concerns that water carried in uncoated pipes could contain asbestos fibers.
In municipal waters supplying the cities of Liverpool, Chester, and
Manchester, England, 0.38 ppm (w/w) PCE was found (Pearson and McConnell, 1975).
Munich (Germany) drinking water was analyzed by Lbchner (1976). Samples
taken at various sampling points and times gave a range of 1.1 x 10 ? to 2.4 x
003PE4/B 3-18 11/22/83
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10 3 mg/1. Raw sewage at Munich contained 0.088 mg/1 PCE. On mechanical clari-
fication, the 24-hour average concentration of PCE was 0.0068 mg/1.
3.7 ANALYTICAL METHODOLOGY
Tetrachloroethylene has been analyzed in air and in water, as well as in
biological fluids, by a variety of methods. Separation of PCE from other com-
pounds is usually carried out by gas chromatography (GC). Quantification is
usually made by either electron capture detection (ECD) or mass spectroscopy
(MS). These analytical methods, GC/ECD or GC/MS, have a lower limit of detec-
tion of a few ppt.
3.7.1 Ambient Air
Because PCE levels in air are typically in the sub-ppb range, the sampling
and analysis techniques have been designed to detect trace gas levels.
3.7.1.1 Sampling and Sources of Error—Because of the low levels occurring in
ambient air, sampling techniques have focused on adsorption onto solids such
as charcoal (Evans et al., 1979) or on concentration methods that increase the
amount of PCE to above detection limits (Rasmussen et al., 1977). In the upper
troposphere, PCE has been sampled by pumping air into stainless steel or glass
containers until there is a positive pressure relative to the surrounding atmos-
phere (Singh et al., 1979).
Evans et al. (1979) sampled PCE using a method based on adsorption onto
activated charcoal, followed by desorption by carbon disulfide/methanol. The
precision of the analytical method, expressed as a coefficient of variation
for the total measurement system (including sample collection, handling, and
preparation) was reported as 16 percent. When 49 quality control samples were
analyzed, the overall percent recovery from the charcoal tubes was 70.2 ±1.7
percent. When the total measurement system was independently checked by using
Tenax GC, another solid adsorbent, the paired data were correlated with a
coefficient of 0.82, with the average Tenax result exceeding the average char-
coal result by 21 percent. Samples were in the sub- to low-ppb range. Evans
et al. (1979) reported that PCE is stable on charcoal tubes for at least one
month at 0°C. The lower limit of detection for the total measurement method
(to include GC/ECD) was estimated at 0.68 ug/m3 (0.1 ppb).
Pellizzari and Bunch (1979) reported the use of Tenax GC, a porous poly-
mer based on 2,6-diphenyl-p-phenylene oxide, to adsorb PCE from ambient air.
Recovery was made by thermal desorption and helium purging into a freezeout
003PE4/B 3-19 11/22/83
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trap. The estimated detection limit, when high resolution GC/MS is used, is
0.38 ppt (2.5 x 10 6 mg/m?). Accuracy of analysis was reported at ± 30 percent.
Included among the inherent analytical errors were (1) the ability to accurate-
ly determine the breakthrough volume, (2) the percent recovery from the sam-
pling cartridge after a period of storage, and (3) the reproducibility of ther-
mal desorption from the cartridge and its introduction into the analytical
system. To minimize loss of sample, cartridge samplers should be enclosed in
cartridge holders and placed in a second container that can be sealed, protec-
ted from light, and stored at 0°C. The advantages reported for Tenax include
(1) low water retention, (2) high thermal stability, and (3) low background
levels (Pellizzari, 1974, 1975, 1977; Pellizzari et al., 1976). Singh et al.
(1982) have cautioned that Tenax suffers from serious artifact problems.
Krost et al. (1982) reported an estimated detection limit of 0.3 ppt for
PCE using high resolution GC/MS. The detection limit was calculated on the
basis of the breakthrough volume for a known amount of Tenax GC at 10°, 21°,
and 32°C. Field sampling and analysis precision of the Tenax method was found
to range from ± 10 to ± 40 percent relative standard deviation for different
substances when replicate field sampling cartridges were examined.
Knoll et al. (1979) reported resolution of PCE from other chlorinated hy-
drocarbons with Carbopak C-HT, a graphitized thermal carbon black treated with
hydrogen at 1000°C. Carbowax 20M, reacted with nitroterephthalic acid, was
reported not to give good separation. Porapak T, a porous polymer based on
ethylene glycol dimethacrylate, was reported to give good separation.
A freezeout concentration has been developed by Rasmussen and coworkers
(1977) to determine trace levels of PCE in the presence of other compounds.
The detection limit was reported at 0.2 ppt (1.36 x 10 6 mg/m?) for 500-ml
aliquots of ambient air samples measured by GC coupled with EDC. When freeze-
out is complete, PCE remains behind, and such gases as oxygen and nitrogen are
passed through as the freezeout loop is heated. The carrier gas sweeps the
contents onto the column.
Singh et al. (1979, 1982) have employed the cryogenic trapping of air
containing trace levels of PCE and other compounds of interest. During sam-
pling, traps are maintained at liquid oxygen temperature. Traps were made of
stainless steel packed with a 4-inch bed of glass beads or glass wool. Ali-
quots are thermally desorbed and injected directly into the gas chromatograph.
Both electric heating and hot water desorption techniques were found to be
satisfactory.
003PE4/B 3-20 11/22/83
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Makide et al. (1980) employed stainless steel canisters for sampling of
air containing PCE at levels of about 20 ppt. Canisters were polished electro-
chemically. Canisters were evaluated to 10 4 Pa at 200°C before sampling.
The composition of the samples was reported to have remained unchanged for over
a year.
The monitoring method recommended by the National Institute for Occupa-
tional Safety and Health (1977) utilizes adsorption onto charcoal followed by
desorption with carbon disulfide. This method is recommended for the range 96
to 405 ppm (655 to 2749 mg/m3). The coefficient of variation for the analyti-
cal and sampling method is 0.052. A reported disadvantage is that the char-
coal may be overloaded, thus limiting the amount of sample that can be collec-
ted. This may be obviated by using more than one adsorber.
Budde and Eichelberger (1979) reported that carbon adsorption methods
generally have more disadvantages than those methods using porous polymers.
The advantage of porous polymers coupled with thermal desorption, as contras-
ted with solvent desorption, is higher sensitivity, because the total sample
is measured and there is no background from the solvent. However, because the
total sample is measured, multiple samples must be collected to insure against
accident and loss of sample, and to obtain information on the precision of the
method. Tenax GC was reported to be superior to other polymers for organics
analysis. Samples are taken by pulling air through glass tubes packed with
Tenax GC, 60/80 mesh and supported by plugs of glass wool. After a suitable
sampling period (about 2 to 4 hours in urban areas), tubes are capped and
stored. Samples are thermally desorbed (250 to 270°C) for 3 minutes under a
10-ml helium flow.
Criteria for evaluating methods using solid sorbents to collect organic
compounds from air have been discussed by Melcher et al. (1978). Among the
factors to be considered are effects of (1) size of collection tube (2) break-
through volume, (3) humidity, (4) temperature, (5) migration, (6) desorption
efficiency, and (7) concentration.
3.7.1.2 Analysis—The sampling methods which use solid adsorbents or cryogenic
techniques have the trap connected to the gas chromatograph by multiple-port
gas sampling valves. With solid traps, the collected organics are quickly
heated and the desorbed organics are passed through capillary columns. A
number of coating materials in the capillary columns have been successfully
used for separating PCE. These materials include (1) SF-96 on 100/120 mesh
003PE4/B 3-21 11/22/83
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Chromosorb W (Cronn et al.,1977); (2) SP-2100 on 80/100 mesh Supelcoport (Singh
et al., 1979); (3) 80/100 mesh Carbopak C-HT, Porapak T, and SP-2100/0.1 percent
Carbowax 1500 on 100/ 120 mesh Supelcoport (Knoll et al., 1979). In the method
used by Evans et al. (1979) in field studies, a 1.8-m glass column with a
2-mm i.d. , packed with 0.1 percent SP-1000 on Carbopack C, 80/100 mesh was used
to separate PCE from other organics in ambient air samples. Twenty-four-hour
samples were adsorbed onto charcoal. After desorption with carbon disulfide/
methanol, a 1.0-ul aliquot was injected into the gas chromatograph. The separa-
tion conditions included an oven temperature of 125°C (all transfer lines at
least 170°C), and the carrier gas was 5 percent methane in argon. Quantifica-
tion was made by ECD (Nickel 63; ECD temperature of 218°C) and a standing
current of 0.5 amperes. To insure that the cell was not contaminated, the
sensitivity of the detector was evaluated by comparing the standing current
with the pulse frequency curve.
Electron capture detection is a method of choice used by a number of inves-
tigators (Singh et al., 1977, 1979, 1982; Rasmussen et al., 1977). Singh et
al. (1979) maintained the ECO at a higher temperature (325°C) than did Evans et
al. (1979), because it was found that the ECD response increased with an increase
in temperature. The identity of PCE was confirmed by determining its ionization
efficiency as well as the EC thermal response.
More recently, Singh et al. (1982) maintained the ECD at 275°C with a
carrier flow rate of 40 ml/min to analyze for PCE and 11 other halogenated
organics in ambient air. Identity of PCE and other compounds was established
from retention times on multiple columns, the ECD thermal response, and the
ECD ionization efficiency. Lillian and Singh (1974) reported that the accuracy
associated with GC-ECD measurements of compounds having ionization efficiencies
exceeding 50 percent is 75 percent or greater. Using two ECDs in series, PCE
was found to have an ionization efficiency of 70 percent. In a comparison of
GC-ECD with GC-MS, Cronn et al. (1976) judged GC-ECD to be superior in repro-
ducibility for quantitating halocarbons. Of four halocarbon standards (PCE
not among them) measured by GC-ECD, the coefficients of variation ranged from
1.4 to 4.3 percent, compared to a range of 4 to 19 percent when 11 halocarbon
standards were measured by GC-MS. A close agreement between the levels of PCE
and other halocarbons determined by GC-ECD and GC-MS on the same ambient air
samples was obtained by Russell and Shadoff (1977).
GC-ECD was used by Pellizzari et al. (1979) to measure PCE in ambient air
samples. Samples were adsorbed onto charcoal and desorbed with a mixture of
003PE4/B 3-22 11/22/83
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methanol and carbon disulfide, and aliquots were separated on a 2.5-mm (i.d.)
Pyrex column containing 0.2 percent Carbowax 1500 on Carbopack C. The esti-
mated detection limit was 2.5 x 10 6 mg/m3 (0.38 ppt).
Makide et al. (1980) separated trace levels of PCE from other halogenated
organics on a silicone OV-101 column (10 percent by weight coated on Chromosorb
W-HP, 80-100 mesh) of 5-mm i.d. and 3 m long. Samples were transferred to the
column cooled at -40°C during preconcentration. Separation was carried out by
raising column temperature 5°C per minute up to 70°C. Methane was added to
the carrier gas (nitrogen) to improve the signal-to-noise ratio and to stabilize
the baseline. Quantification was made by a constant-current ECD. The detection
limit for PCE was reported as <0.05 ppt. Precision was reported to be within
2 percent.
To measure PCE at levels expected to occur in occupational air, flame ion-
ization detection has been used. The analytical method S335, suggested by the
National Institute for Occupational Safety and Health (1977) for organic sol-
vents in air, utilizes adsorption onto charcoal, followed by desorption with
carbon disulfide. PCE is separated by GC. The method is recommended for the
range 96 to 405 ppm (655 to 2749 mg/m3). The coefficient of variation for the
analytical sampling method is 5.2 percent. With the method, interferences are
minimal, and those that do occur can be eliminated by altering chromatographic
conditions.
3.7.2 Water
3.7.2.1 Sampling--A variety of techniques and methods are commonly used to
sample trace levels of PCE and other halogenated organics in water samples.
Coleman et al. (1981) have reported that the Grob closed-loop-stripping tech-
nique is an excellent tool to monitor organics in water at the ppt level. It
was reported that a million-fold concentration of most low and intermediate
molecular weight organics can be achieved. Quantisation is performed by spik-
ing the initial water sample with a series of internal standards, stripping at
30°C for two hours, and by chromatographing the CS2 extract on a wall-coated
open-tubular capillary.
3.7.2.1.1 Gas purging and trapping. In this method, finely divided gas bubbles
are passed through the sample, transferring the organic compounds to the gas
phase. The gas is then passed through a solid adsorbent in a trap. Compounds
are desorbed at elevated temperature by backflushing with a carrier gas into
003PE4/B 3-23 11/22/83
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the gas chromatograph (Budde and Eichelberger, 1979). Since the boiling point
of PCE is 121°C, Tenax GC would be an effective absorbent.
The purge and trap procedure is widely used, as is the purging device
developed by Bellar and Lichtenberg (1974). For most organic compounds, detec-
tion limits as low as 1 ^9/1 can be obtained when GC/MS is used for analysis.
Mieure (1980) reported that adding salt to the sample or increasing its tempera-
ture dramatically improves the removal of most organic compounds.
Otson and Williams (1982) have described a modified purge and trap tech-
nique for evaluation of volatile organic pollutants in water. The detection
limit reported for PCE was 0.1 ug/1 with ECD, and 1 ug/1 with flame ionization
detection (FID). Tenax GC was used as packing for the combined trap/chromatog-
raphic column.
3.7.2.1.2 Headspace analysis. This method describes static sampling of the
vapor phase that is in equilibrium with the aqueous sample. The concentration
in the headspace is proportional to the concentration in the water (Kepner et
al., 1964). In this procedure, trace organics in the range of 10 to 100 ug/1
can be sampled (Mieure, 1980). Mieure (1980) reported a detection limit for
PCE (analyzed by ECD) of 0.01 ug/1. With flame ionization detection, the limit
was 32 ug/1. Typically, a 1- to 2-ml sample of the headspace is removed and
injected into the gas chromatograph. Headspace extraction coupled with mixed
column separation and ECD analysis was reported by Castello et al. (1982) to
be suitable for rapid screening of drinking water supplies.
3.7.2.1.3 Liquid/liquid extraction. Mieure (1980) reported that recovery of
PCE from water spiked with 2.3 to 90 ug/1 ranges from 100 to 113 percent. The
precision ranges from 10 to 12 (RSD). These results were obtained from a
round-robin study, by the American Society for Testing and Materials (ASTM)
Committee D19 on Water, using liquid/liquid extraction. The extractant was
not identified.
Budde and Eichelberger (1979) cautioned that a disadvantage to this method
is that very volatile compounds may be lost during extract concentration or
during solvent elution from the gas chromatograph. Methylene chloride was
recommended as the only general-purpose solvent.
Sheldon and Hites (1978) used methylene chloride in a sampling procedure
applied to the identification of PCE and 98 other organic compounds in river
water. Grab samples were collected in amber glass bottles and samples for
solvent extraction were immediately preserved by acidifying to pH 2 with
hydrochloric acid and by adding 250 ml of methylene chloride. The analytical
003PE4/B 3-24 11/22/83
-------
techniques used were those reported by Jungclaus et al. (1978). Solvent extrac-
tion efficiencies were not determined. While PCE was previously detected in
vapor stripping analysis of prior samples, it was not detected in the water
samples cited in their report.
3.7.2.2 Analysis—Schwarzenbach et al. (1979) used ECD to measure PCE levels
in water samples. Volatile organics were purged and adsorbed onto charcoal.
Desorption was by carbon disulfide. Quantification was made by FID and ECD.
Dowty et al. (1975a) used Tenax GC in trapping purgeable organics from
water samples. The polymer containing the trapped organics was placed in the
GC injection port maintained at 200°C. Final separation was made on a glass
capillary column coated with Pluronics 121. The effluent of the column was
split to allow for FID and ECD.
3.7.3 Biological Media
Ramsey and Flanagan (1982) have described a gas chromatographic method
reported to be suitable for analysis of PCE and other organics present in blood.
Detection was both by flame ionization and ECD. Approximately 200 ul of blood
or 200 mg tissue is required for analysis.
3.7.4 Calibration
Singh et al. (1982) found that primary standards of PCE in the low-ppb
range could be satisfactorily calibrated using permeation tubes maintained
either at 30° or 70°C. Permeation tubes were standard FEP or TFE Teflon. All
permeation tubes were conditioned for two weeks or longer. Errors in the per-
meation rate were ± 10 percent. Singh et al. (1982) found that long-term (2
years) stability of primary standards (10 ppm) of PCE in aluminum containers
was excellent.
3.7.5 Storage and Stability of PCE
Sampling of exhaled breath commonly is accomplished by use of Saran bags
or glass pipettes. Temperature and storage time of the samples before analysis
are factors to be considered in obtaining accurate data.
3.7.5.1 Glass Sampling Tubes—Evaluation of glass sampling tubes was made by
Pasquini (1978). Serial alveolar breath samples were collected in the tubes
and the concentrations of PCE were analyzed by a gas chromatograph equipped with
a flame ionization detector. Analysis of vapor retention over 169 hours indi-
cated that glass tubes can be acceptable containers for breath samples if
003PE4/B 3-25 11/22/83
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precautions are taken. Moisture, temperature, and tube surface and condition
can greatly alter vapor retention.
In tubes filled with breath samples taken at room temperature and also
stored at room temperature, the mean percent loss of PCE was 64.8 ± 9.4. Par-
titioning of PCE between the vapor and liquid states appears to be a reasonable
explanation for vapor retention loss. It was shown for trichloroethylene that
if storage tubes were maintained at 37°C, vapor retention was greater. It was
also greater if si 1 iconized tubes were used.
3.7.5.2 Saran, Teflon and Tedlar Containers—Saran bags as storage containers
for PCE vapors have been evaluated by Desbaumes and Imhoff (1971). Although
it was concluded that Saran can be an acceptable container, the diffusion rate
was appreciable over a 24-hour storage period. Storage temperature was not
reported.
Teflon containers were judged by Drasche et al. (1972) to be more suit-
able than Saran even though losses of PCE due to adherence to Teflon surfaces
were appreciable. Within the first 30 minutes after introduction of a mixture
(relative humidity = 45 percent) of benzene, trichloroethylene, and PCE into a
Teflon bag, vapor concentrations of each dropped 40 to 60 percent. However,
when the bag was heated to 100°C for 30 minutes after the mixture had been stored
for 44 hours at 25°C, concentrations rose to the initial values.
Knoll et al. (1979) reported that PCE, when stored at ambient tempera-
tures for 10 days or less in Tedlar bags, was stable. When the vapor mixture
is heated to 70°C, PCE is stable for no more than 5 hours.
003PE4/B 3-26 11/22/83
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003PE4/B 3-34 11/22/83
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4. ECOSYSTEM CONSIDERATIONS
4.1 EFFECTS ON AQUATIC ORGANISMS AND PLANTS
Tetrachloroethylene (PCE) has been tested for acute toxicity in a limited
number of aquatic species. The information presented in this chapter presents
observed levels reported to result in adverse effects under laboratory condi-
tions. It is recognized that such parameters of toxicity are not easily extra-
polated to environmental situations. Test populations themselves may not be
representative of the entire species, in which susceptibility of various life-
stages to the test substance may vary considerably. Guidelines for the utili-
zation of these data in the development of criteria levels for PCE in water
are discussed elsewhere (U.S. EPA, 1979).
The toxicity of PCE to fish and other aquatic organisms has been gauged
principally by flow-through and static testing methods (Committee on Methods
for Toxicity Tests with Aquatic Organisms, 1975). The flow-through method
exposes the organism(s) continuously to a constant concentration of PCE while
oxygen is continuously replenished and waste products are removed. A static
test, on the other hand, exposes the organism(s) to the added initial concen-
tration only. Both types of tests are commonly used as initial indicators of
the potential of substances to cause adverse effects.
4.1.1 Effects on Freshwater Species
Alexander et al. (1978) used both flow-through (measured) and static
(unmeasured) methods to investigate the acute toxicity of four chlorinated
solvents, including PCE, to adult fathead minnows (Pimephales promelas).
Studies were conducted in accordance with test methods described by the
Committee on Methods for Toxicity Tests with Aquatic Organisms (1975).
The static and flow-through results for the 96-hour experiments indicated
that PCE was the most toxic of the solvents tested. The lethal concentration
(96-hour LC50) necessary to kill 50 percent of the fathead minnows in the
flow-through test was 18.4 mg/1 (18.4 ppm); the 95 percent confidence limits
were 14.8 to 21.3 mg/1 (14.8 to 21.3 ppm). In comparison, the static experi-
ments gave a 96-hour LC50 of 21.4 mg/1 (21.4 ppm); the 95 percent confidence
limits were 16.5 to 26.4 mg/1 (16.5 to 26.4 ppm). Fish affected during expo-
sure were transferred to static freshwater aquaria at the end of exposure.
Only those fish severely affected by high concentrations did not recover.
003PE1/D 4-1 11/22/83
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When the minnows were exposed to sublethal levels for short exposure
intervals, only reversible effects were observed. Endpoints evaluated were
loss of equilibrium, melanization, narcosis, and swollen, hemorrhaging gills.
The effective flow-through concentration (EC50) of PCE that produced one or
more of these reversible effects was 14.4 mg/1 (14.4 ppm).
The 96-hour LC50) in a static test with the bluegill (Lepomis macrochirus),
was reported as 12.9 mg/1 (12.9 ppm) (U.S. EPA, 1978, 1980). The most sensitive
species tested is the rainbow trout (Salmo gairdneri). The LC5o determined by
a flow-through measured procedure was 5.28 mg/1 (5.28 ppm) (U.S. EPA, 1980).
With embryo-larval test procedures, a chronic value of 0.840 mg/1 (0.840 ppm)
was obtained by the U.S. EPA (1980) for the fathead minnow.
With the freshwater invertebrate Daphnia magna, a 48-hour EC5(j value of
17.7 mg/1 (17.7 ppm) was obtained (U.S. EPA, 1980). The midge Tanytarsus
dissimi1 is was more resistant, with a 48-hour LC5o value of 30.84 mg/1 (30-84
ppm) determined under static, measured conditions.
4.1.2 Effects on Aquatic Plants
As cited in the U.S. EPA Ambient Water Quality Criteria Document (U.S.
EPA, 1980), no adverse effects on chlorophyll a or cell numbers of the fresh-
water alga Selenastrum capricornutum were observed at exposure concentrations
as high as 816 mg/1 (816 ppm).
For the saltwater species Skeletonema costatum a 96-hour EC50 of about
500 mg/1 (500 ppm) was determined for effects on chlorophyll a and cell number.
This alga species is more resistant than the alga Phaeodactylum tricornutum for
which the EC50 value was determined to be 10.5 mg/1 (10 ppm) (Pearson and
McConnell, 1975).
4.1.3 Effects on Saltwater Species
Pearson and McConnell (1975) investigated the acute toxicity of PCE on
the dab (Limanda 1imanda), barnacle larvae (Barnacle nauplii), and on unicell-
ular algae (Phaeodactylum tricornutum). The LC50 was 5 mg/1 (5 ppm) for the
dab. The 48-hour LC5o for barnacle larvae was 3.5 mg/1 (3.5 ppm).
Toxicity to the unicellular alga was assessed by measuring alterations in
the uptake of carbon from atmospheric carbon dioxide during photosynthesis.
14
Uptake of carbon dioxide was measured by the use of sodium- C-carbonate. The
EC50 was 10.5 mg/1 (10.5 ppm).
003PE1/D 4-2 11/22/83
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Data collected by the U.S. Environmental Protection Agency (1980) indi-
cate that, for mysid shrimp (Mysidopsis bahia), the LC5u was 10.2 mg/1 (10.2
ppm) in a 96-hour static, unmeasured procedure. Chronic testing over the life
cycle of the mysid shrimp resulted in a chronic value of 0.450 mg/1 (0.45 ppm)
(U.S. EPA, 1980). The chronic value is 0.044 times the 96-hour LC5u.
4.2 BIOCONCENTRATION AND BIOACCUMULATION
An indicator of the potential for a substance to result in cumulative or
chronic toxic effects in aquatic species is the bioconcentration factor (BCF).
Bioconcentration refers to the increased concentration of a substance within
an organism (e.g., fish) relative to the ambient water concentration under
steady-state conditions. As defined by Veith et al. (1979), the bioconcentration
factor is a constant of proportionality between the concentration of the chemical
in fish and in the water. This can be more clearly expressed as
Cf Id
C~ = K~ = KRCF at steady state. (4-1)
w ?
Bioaccumulation, a term often erroneously used in place of bioconcentration, can
be defined as that process which includes bioconcentration and any uptake of
toxic substances through consumption of one organism by another. The BCF alone,
however, may not be the most useful measure of the overall fate of a substance
in water or of its potential for producing toxic effects, for all chemicals.
In the absence of direct measurement, a measure commonly used to assess
the degree to which a compound may be bioconcentrated is the octanol-water par-
tition coefficient. Estimates for the octanol/water partition coefficient range
from 339 to 871 (Neely et al., 1974; U.S. EPA, 1980; Chiou et al., 1977). The
partition coefficient has been shown to be directly related to bioconcentration
potential in fish (Neely et al., 1974). In guidelines recently set forth by
the American Society for Testing and Materials (ASTM), a log partition coeffi-
cient exceeding a value of three was considered an indication of a high proba-
bility of measurable bioaccumulation in aquatic species (ASTM, 1978). Compounds
that exhibit a large log coefficient generally are those with low water solu-
bility and high solubility in organic solvents. Although a compound may demon-
strate a high BCF or log partition coefficient, other environmental factors that
act to reduce this potential often exist. The compound may be rapidly hydrolyzed
003PE1/D 4-3 11/22/83
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or degraded by other mechanisms. Measurable uptake by the organism may be pre-
cluded if the tissue depuration rate for the substance is great.
With regard to PCE, the BCF was calculated to be 34 and 49 in two fish
species (U.S. EPA, 1980; Neely et al., 1974). Neely et al. (1974) found that
the BCF for PCE and other chemicals was linearly related to the respective parti'
tion coefficients. For PCE, the log partition coefficient was 2.88, and the
BCF, determined in trout (rainbow) muscle, was 39.6 ± 5.5. The trout were
exposed to two undefined levels of PCE for an undefined period of time. The
extent to which the levels approached the acute LC5U level for this species or
whether a steady-state was achieved was not reported. Preliminary data in the
U.S. Environmental Protection Agency study (1980) with bluegill indicated a
BCF of 49. The log partition coefficient was 2.53. The depuration rate was
rapid, with a half-life of less than one day.
Although these studies suggest that PCE does have bioconcentration poten-
tial, the extent to which this potential can be manifested in the form of
adverse effects can be gauged only from the results of toxicological studies.
4.2.1 Levels of PCE in Tissues of Aquatic Species
Pearson and McConnell (1975) suggested that chronic and sublethal effects
of PCE may result from exposure to low concentrations of PCE, if the halo-
carbon can be bioaccumulated. As a first step in addressing the question of
bioaccumulation, these investigators determined levels of PCE in a variety of
invertebrate and vertebrate species (Tables 4-1 and 4-2).
Among marine invertebrates, wet tissue concentrations of PCE were found
to range from 0.001 to 0.009 ppm. The highest concentration (0.008 to 0.009
ppm) found was in the crab (Cancer pagurus). Higher levels were found in
marine algae (0.013 to 0.022 ppm). In tissues of fish, a range of 0.0003 to
0.041 ppm was found. Concentrations in the livers of three species of fish
were found to greatly exceed those found in the flesh. Tissue levels from all
species are shown in Table 4-1. Concentrations reported for fish-eating birds
and marine mammals were for selected tissues such as fish liver, sea bird
eggs, and seal blubber. If the reported tissue concentrations for birds and
mammals are converted to a whole-body weight basis, concentrations are much
lower and closer to concentrations measured in seawater, indicating little
or no bioconcentration and biomagnification (U.S. EPA, 1981).
003PE1/D 4-4 11/22/83
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TABLE 4-1. LEVELS OF PCE IN TISSUES OF MARINE ORGANISMS, BIRDS, AND MAMMALS
en
Species
Invertebrates
Plankton
Plankton
Ragworm (Nereis diversicolor)
Mussel (Mytilus edulis)
Cockle (Cerastoderma edule)
Oyster (Ostrea edulis)
Whelk (Buccinum undatum)
Slipper limpet (Crepidula
fornicata)
Crab (Cancer pagurus)
Shorecrab (Carcinus maenus)
Hermit crab (Eupagurus
bernhardus)
Source
Liverpool Bay
Torbay
Mersey Estuary
Liverpool Bay
Firth of Forth
Thames Estuary
Liverpool Bay
Thames Estuary
Thames Estuary
Thames Estuary
Tees Bay
Liverpoool Bay
Firth of Forth
Firth of Forth
Firth of Forth
Thames Estuary
Trichlorg-
ethylene
Tissue (ppm x 103)
0.05 - 0.4
0.9
Not detected
4 - 11.9
9
8
6 - 11
2
Not detected
9
2.6
10 - 12
15
12
15
5
PCE
(ppm x 103)
0.05 - 0.5
2.3
2.9
1.3 - 6.4
9
1
2-3
0.5
1
2
2.3
8-9
7
6
15
2
Shrimp (Crangon crangon)
Firth of Forth
16
-------
TABLE 4-1. (continued)
Species
Starfish (Asterias rubens)
Sunstar (Solaster sp. )
Sea Urchin (Echinus esculentus)
Marine Algae
Enteromorpha compressa
Ulva lactuca
Fucus vesiculosus
Fucus serratus
Fucus spiral is
Fish
Ray (Raja clavata)
Plaice (Pleuronectes platessa)
Flounder (Platyethys f lesus)
Dab (Limanda limanda)
Source
Thames Estuary
Thames Estuary
Thames Estuary
Mersey Estuary
Mersey Estuary
Mersey Estuary
Mersey Estuary
Mersey Estuary
Liverpool Bay
Liverpool Bay
Liverpool Bay
Liverpool Bay
Trichloro-
ethylene
Tissue (ppm x 103)
5
2
1
19 - 20
23
17 - 18
22
16
flesh 0.8 - 5
liver 5-56
flesh 0.8 - 8
liver 16 - 20
flesh 3
liver 2
flesh 3-5
liver 12 - 21
PCE
(ppm x 103)
1
2
1
14 - 14.5
22
13 - 20
15
13
0.3 - 8
14 - 41
4-8
11 - 28
2
1
1.5 - 11
15 - 30
-------
TABLE 4-1. (continued)
Species
Mackerel (Scomber scombrus)
Dab (Limanda limanda)
Plaice (Pleuronectes platessa)
Sole (Solea solea)
Red gurnard (Aspitrigla
cuculus)
Scad (Trachurus trachurus)
Pout (Trisopterus luscus)
Spurdog (Squalus acanthi as)
Mackerel (Scomber scombrus)
Clupea sprattus
Cod (Gadus morrhua)
Sea and Freshwater Birds
Gannet (Sula bassana)
Source
Liverpool Bay
Redcar, Yorks
Thames Estuary
Thames Estuary
Thames Estuary
Thames Estuary
Thames Estuary
Thames Estuary
Thames Estuary
Torbay, Devon
Torbay, Devon
Torbay, Devon
Irish Sea
Tissue
flesh
1 iver
flesh
flesh
flesh
flesh
guts
flesh
guts
flesh
flesh
flesh
flesh
flesh
flesh
Air
bladder
1 iver
eggs
Trichloro-
ethylene
(ppm x 103)
5
8
4.6
2
3
2
11
11
6
2
2
3
2.1
3.4
0.8
<0.1
4.5 - 6
9-17
PCE
(ppm x 103)
1
not
5.1
3
3
4
1
1
2
4
2
1
Not
1.6
<0.
3.6
1.5
4.5
detected
detected
1
- 3.2
- 26
Shag (Phalacrocerax aristotelis) Irish Sea
eggs
2.4
1.4
-------
TABLE 4-1. (continued)
00
Species
Razorbill (Alca torda)
Kittiwake (Rissa tridactyla)
Swan (Cygnus olor)
Moorhen (Gallinula chloropus)
Mallard (Anas platyrynchos)
Mammals
Grey Seal (Halichoerus grypus
Common Shrew (Sorex araneus)
Source
Irish Sea
North Sea
Frodsham Marsh
Merseyside
Merseyside
Fame Island
Frodsham Marsh
Tissue
eggs
eggs
1 iver
kidney
liver
muscle
eggs
eggs
blubber
1 iver
-
Trichloro-
ethylene
(ppm x 103)
28 - 29
33
2.1
14
6
2.5
6.2 - 7.8
9.8 - 16
2.5 - 7.2
3 - 6.2
2.6 - 7.8
PCE
(ppm x 103)
32-39
25
1.9
6.4
3.1
0.7
1.3 - 2.5
1.9 - 4.5
0.6 - 19
0 - 3.2
1
Levels for trichloroethylene included for comparative purposes.
Source: Pearson and McConnell, 1975.
-------
TABLE 4-2. ACCUMULATION OF PCE BY DABS
Period of
Tissue Exposure (days)
flesh 3-35
liver 3-35
flesh 3 - 35
liver 3-35
flesh 10
liver 10
Mean Exposure
Concentration (ppm)
0.3
0.3
0.03
0.03
0.2
0.2
Mean
Concentration in
Tissue (ppm x 100)
2,800a (13)
113,000 (14)
60 (9)
7,400b (9)
1,300 (7)
69,000 (7)
Accumu-
lation
Factor
x 9
x 400
x 5
x 200
x 6
x 350
Numbers in parentheses are number of specimens analyzed.
aOne fish had a flesh concentration of 29.7 ppm and was omitted from calculations.
One fish had flesh concentration of 50.3 ppm and was omitted from calculations.
Source: Pearson and McConnell, 1975.
The average concentration of PCE in seawater taken from Liverpool Bay, an
area where many species of organisms were collected, was 0.00012 ppm. A
comparison of this value with those presented in Table 4-1 indicates an uptake
of as much as 75-fold. It was the authors' contention that, based on their ob-
servations, there is little indication that bioaccumulation occurs in the food
chain.
As shown in Table 4-2, dabs (Limanda 1 imanda) exposed to 0.3 ppm for 3
to 35 days were found to have a BCF (liver) for PCE of 400. It was not
reported whether this period of exposure approximated a steady-state for PCE.
After dabs were returned to clean seawater, the level of PCE dropped to 1/100
of the original level in 4 days and to 1/1000 of the initial level after 11
days (Figure 4-1). The ratio between liver and flesh concentrations is approx-
imately 100 to 1. The relationship between flesh and liver concentrations
in the dab is shown in Figure 4-2.
Dickson and Riley (1976) detected PCE in three species of mollusks and
in five species of fish collected near Port Erin, Isle of Man. Levels of PCE
in various tissues are shown in Table 4-3. Relative to the PCE concentration
in seawater, there was only a slight enrichment in the tissues (< 25 times).
Tetrachloroethylene had one of the lowest mean bioconcentration factors.
003PE1/D 4-9 11/22/83
-------
100
-fie
Ul
X.
01 -
O 1
0.1
d\ B
B^ -
O LIVER ACCUMULATION
0 LIVER LOSS
0 16 32
EXPOSURE TIME, diyt
Figure 4-1. Accumulation and loss of PCE by dabs.
Source: Alexander et al., 1978.
003PE1/D
4-10
11/22/83
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100
Q.
a
I
in
z
ui
UJ
O
oc
O
oc
I-
III
t-
10
0.1
0.01
I
EXPOSURE LEVELS, ppm
O 0.3
Q 0.2
A 0.02
AA
I
10 100
TETRACHLOROETHYLENE IN LIVER, ppm
1000
Figure 4-2. Relation between flesh and liver concentration of PCE in dabs.
Source: Alexander et al., 1978.
003PE1/D
4-11
11/22/83
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TABLE 4-3. CONCENTRATION OF PCE AND TRICHLOROETHYLENE IN MOLLUSKS AND
FISH NEAR THE ISLE OF MAN
Species
Eel (Conger conger)
brain
gin
gut
1 iver
muscle
Cod (Gadus morhua)
brain
gill
heart
1 iver
muscle
PCE
(mg x
6
2
3
43
1
3
3
8
2
Trichloroethylene
106/g dry weight tissue)
62
29
29
43
70
56
21
11
66
8
skeletal tissue
stomach
Coalfish (Pollachius birens)
alimentary canal
brain
gill
heart
1 iver
muscle
Dogfish (Scyl1iorhinus canicula)
brain
gill
gut
heart
liver
muscle
spleen
Bib (Trisopterus luscus)
brain
gut
1 iver
muscle
skeletal tissue
6
2
12
13
4
0.3
306
71
70
8
40
176
41
274
479
41
307
143
187
185
003PE1/D
4-12
11/22/83
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TABLE 4-3. (continued)
PCE
Species
Trichloroethylene
(mg x
dry weight tissue)
Baccinum undatum
digestive gland
muscle
Modi ol us modi ol us
33
39
digestive tissue
mantle
muscle
Pecten maximus
gill
mantle
muscle
ovary
testis
-
63
16
88
40
24
-
176
56
250
33
detected
-
-
-
Source: Dickson and Riley, 1976.
4.3 BEHAVIOR IN WATER AND SOIL
The potential of any substance for bioconcentration is influenced by many
factors, including the rate at which it volatizes and its reactivity.
In the laboratory study by Oil ling et al. (1975), the measured half-life
of PCE ranged from 24 to 28 minutes in water. Factors affecting the evaporation
of PCE were surface wind speed, agitation of the water, and water and air
temperatures. Reactivity of PCE in water was measured by exposing sealed
quartz tubes containing 1 ppm PCE to sunlight for one year. At 6 months, the
level of PCE had declined to 0.52 ppm, and at 1 year, to 0.25 ppm. Oil ling et
al. (1975) reported that the presence of 3 percent NaCl (as in seawater)
caused about a 10 percent decrease in the evaporation after 40 percent had
already evaporated. The addition of 500 ppm clay appeared to increase the
rate of disappearance to 85 percent soluble depletion at 20 minutes. These
experiments were conducted to simulate the evaporation of PCE under conditions
more nearly like those found in the environment. Evaporation of PCE from the
hydrosphere is a rapid process.
003PE1/D
4-13
11/22/83
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The field studies of Zoeteman et al. (1980) suggest that PCE is more
persistent in natural water environments than is indicated by laboratory mea-
surements. In a field study of the persistence of a variety of organic chemi-
cals in different aquatic environments in the Netherlands, Zoeteman et al.
(1980) estimated the persistence of PCE in river water from 3 to 30 days
(half-life). In lakes and groundwaters, the half-life was estimated at 10-fold
higher. Estimates were derived from monitored values of samples collected
between two sites along the Rhine River, into which no discharges were expected.
PCE was analyzed by GC-MS.
Estimates of the persistence of PCE in rivers, lakes, and ponds, by
calculation according to Smith et al. (1980) are in general agreement with
the field results of Zoeteman and coworkers (1980). The half-life of PCE is
obtained from the expression:
t1/2 = 0.693. (4-2)
kv
where kc is the volatilization rate constant.
Using the data provided by Smith et al. (1980), the tl,2 (days) is as
follows: ponds, 9 to 20; lakes, < 1 to 20; rivers, < 1 to 20.
Bouwer et al. (1981) found that PCE and other halogenated organics have
the potential to leach rapidly through soil. When secondary treated municipal
wastewater containing from 1 to 10 ng/1 PCE was applied to soil columns, at
rates typical of high-rate land application systems and under conditions in
which volatilization was prevented, PCE was detected in the effluent. Leaching
of PCE through soil was suggested by Zoeteman et al. (1980) as a probable
factor in the contamination of groundwater supplies in the Netherlands.
The potential for halogenated organics, including PCE, to contaminate
groundwater supplies via leaching from surface waters was examined by Scwarzen-
bach and Westall (1981). In batch and column experiments with various types
of sorbents and organics designed to simulate field conditions, these investi-
gators found that the partition coefficient for a particular compound can be
estimated from its octanol/water partition coefficient and from the fraction
of organic carbon in the sorbent. A high degree of correlation was found
between the partition coefficient and organic carbon content when the fraction
003PE1/D 4-14 11/22/83
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of organic carbon was greater than 0.1 percent. A partition coefficient of
0.56 ± 0.09 was found for PCE, using natural aquifer material (organic carbon
= 0.15 percent) from a field site in Switzerland. It was concluded that, for
concentrations typically encountered in natural waters, sorption of PCE and
other organics of comparable 1ipophilicity by aquifer materials is reversible.
The expression
S = kpC, (4-3)
where S = concentration in solid phase
kp = partition coefficient
C = concentration in liquid phase
was found satisfactory to describe sorption equilibrium.
4.4 SUMMARY
The available data for PCE indicate that acute and chronic toxicity to
freshwater aquatic life can occur at concentrations around 5280 and 840 ug/1,
respectively. For saltwater aquatic life, the acute and chronic toxicity values
are 10,200 and 450 ug/1, respectively.
Tetrachloroethylene does not appear to biomagnify or concentrate as it
moves up the food chain. The available data suggest that the bioconcentration
potential of PCE is low, and it appears to be eliminated rapidly from aquatic
organisms.
Contamination of groundwater supplies by PCE leaching through soil could
be a concern, particularly in situations in which soils of low organic carbon
content are involved.
003PE1/D 4-15 11/22/83
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4.5 REFERENCES
Alexander, H. C., W. M. McCarty, and E. A. Bartlett. Toxicity of perchloro-
ethylene, trichloroethylene, 1,1,1-trichloroethane, and methylene chloride
to fathead minnows. Bull. Environ. Contain. Toxicol. 20:344-352, 1978.
American Society for Testing and Materials. Estimating the hazard of chemical
substances to aquatic life. J. Gavins, K. L. Dickson, and A. W. Maki,
eds. , STP 657. Committee D-19 on Water, 1978.
Bouwer, E. J., P. L. McCarty, and J. C. Lance. Trace organic behavior in soil
columns during rapid infiltration of secondary wastewater. Water Res.
15(1):151-160, 1981.
Chiou, C. T., V. H. Freed, D. W. Schmedding, and R. L. Kohnert. Partition
coefficient and bioaccumulation of selected organic chemicals. Environ.
Sci. Technol. 11:475-478, 1977.
Committee on Methods for Toxicity Tests with Aquatic Organisms: methods for
acute toxicity tests with fish, macroinvertebrates, and amphibians. Ecol.
Res. Series, EPA 600/3-75-009, 1975.
Dickson, A. G., and J. P. Riley. The distribution of short-chain halogenated
aliphatic hydrocarbons in some marine organisms. Marine Pollut. Bull.
7(9):167-169, 1976.
Dilling, W. L., N. B. Tefertiller, and G. J. Kallos. Evaporation rates and
reactivities of methylene chloride, chloroform, 1,1,1-trichloroethane,
trichloroethylene, tetrachloroethylene, and other chlorinated compounds
in dilute aqueous solutions. Environ. Sci. Technol. 9(a):833-838, 1975.
Neely, W. B., D. R. Branson, and G. E. Blau. Partition coefficient to
measure bioconcentration potential of organic chemicals in fish.
Environ. Sci. Technol. 8:1113, 1974.
Pearson, C. R. , and G. McConnell. Chlorinated Cx and C^ hydrocarbons in
the marine environment. Proc. Roy. Soc. London B 189:305332, 1975.
Schwarzenbach, R. P. and J. Westall. Transport of nonpolar organic compounds
from surface water to groundwater. Laboratory sorption studies. Environ.
Sci. Technol. 15:1360-1367, 1981.
Smith, J. H., D. C. Bomberger, Jr., and D. L. Haynes. Prediction of the
volatilization rates of high-volatility chemicals from natural water
bodies. Environ. Sci. Technol. 14(11):1332-1337, 1980.
U.S. Environmental Protection Agency. In-depth studies on health and environ-
mental impacts of selected water pollutants. Contract No. 68-01-4646,
Duluth, MN, 1978.
U.S. Environmental Protection Agency. Tetrachloroethylene: water quality
criteria. Federal Register 44(52):15966-15969, 1979.
003PE1/D 4-16 11/22/83
-------
U.S. Environmental Protection Agency. Tetrachloroethylene: ambient water
quality criteria. Office of Water Regulations and Standards, EPA
440/5-80-073, October 1980.
U.S. Environmental Protection Agency. Environmental risk assessment of tetra-
chloroethylene, draft report. Office of Toxic Substances. 14 September,
1981.
Veith, G. D., D. L. DeFoe, and B. V. Bergstedt. Measuring and estimating the
bioconcentration factor of chemicals in fish. J. Fish. Res. Board Canada
36:1040-0145, 1979.
Zoeteman, B. C. J., K. Harmsen, J. B. H. J. Linders, C. F. H. Morra, and
W. Slooff. Persistent organic pollutants in river water and ground
water of the Netherlands. Chemosphere 9:231-249, 1980.
003PE1/D 4-17 11/22/83
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5. COMPOUND DISTRIBUTION AND RELATED PHARMACOKINETICS
5.1 HUMAN AND ANIMAL STUDIES
5.1.1 Absorption
5.1.1.1 Pulmonary--Inhalation is the principal route by which tetrachloro-
ethylene (PCE) enters the body. During inhalation, PCE is absorbed by the
blood via alveolar air. The principal approaches used in calculating the
amount absorbed are those involving serial breath analysis, i.e., measuring
the amount exhaled.
The magnitude of PCE uptake into the body (dose, burden) depends primari-
ly on several parameters: inspired air concentration, pulmonary ventilation,
duration of exposure, and the rates of diffusion into and solubility in blood
and the various tissues. The concentration of PCE in alveolar air, in equili-
brium with pulmonary venous blood, approaches a minimum difference with the
concentration in the inspiratory air until a steady-state condition is reached.
After tissue and total body equilibrium is reached during exposure, uptake is
balanced by elimination through the lungs and by other routes, including
metabolism. The difference between alveolar and inspiratory air concentrations,
together with the ventilation rate (about 6 1/min at rest), provides a means
of calculating uptake during exposure:
Q = (Cins - Calv> * ' T <5-D
where Q is the quantity absorbed; C is the air concentration in milligrams per
liter; V is the alveolar ventilation rate in liters per minute; and T is the
duration of exposure in minutes. The percent retention is defined as (C.
C i )/C. x 100, and percent retention x quantity inspired (V • T • C. ) is
equal to uptake.
In serial determinations of PCE in alveolar breath and in blood, during
and following exposure (1, 3, 5.5, and 7.5 hr/.day) of males and females to 25,
50, 100, and 150 ppm (170, 339, 678, and 1017 mg/m3), Hake et al. (1976) con-
cluded that the compound was rapidly absorbed and rapidly excreted via the
lungs. The amount absorbed at a given vapor concentration was reported to be
related to the respiratory minute volume. The minute volume is defined as the
product of the tidal volume and the respiratory frequency over a 1-minute
period.
003PE1/E 5-1 11/22/83
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In a study by Monster et al. (1979), six male volunteers were exposed in
a chamber for 4 hours to 72 ± 2 ppm PCE (488 ± 13 mg/nr1) and to 144 ± 7 ppm
PCE (977 ± 47 mg/m3) while at rest. Absorption of PCE by the lungs was reported
to decrease with continuation of exposure (p <0.05); approximately 25 percent
less PCE was absorbed in the fourth hour as compared with the first hour of
exposure. The effects of a workload (bicycle ergometer) were determined in a
separate exposure of the volunteers to 142 ± 6 ppm PCE (963 ± 41 mg/ma);
individuals exercised for two 30-minute periods during the 4-hour exposure
period. A 2-week interval occurred between each exposure mode. During each
mode, the individuals inhaled vapors through a gas mask and exhalations were
®
made into a Tedlar bag.
The total uptake, shown in Table 5-1, was influenced more by lean body
mass than by respiratory minute volume or adipose tissue. Uptake is defined
as the amount of PCE absorbed per minute. The inter-individual coefficient of
variation of body burden predicted by measurements of PCE in exhaled air or in
blood was about 25 percent. The individual uptake at 144 ppm PCE (977 mg/m3)
was 2.1 times higher than that at 72 ppm (488 mg/m3) when individuals were at
rest. During exercise, total uptake increased about 40 percent; exercise had
no effect on the half-life of elimination or on the rate constant of elimina-
tion. Minute volume was three-fold higher than that observed when the indivi-
duals were at rest. An alveolar retention of approximately 60 percent was
calculated at the end of the 4-hour exposure period. Alveolar retention is
defined as the percentage difference between the amount of PCE inhaled and the
amount exhaled.
Monster (1979) suggested that the decrease in absorption (declining lung
clearance) as exposure continued reflected the small degree to which PCE was
metabolized, in contrast to trichloroethylene, which has a more constant rate
of absorption but is metabolized to a much greater degree. There would be no
reason to expect that chronic exposure to environmental levels of PCE would
not follow the patterns of absorption and lung clearance observed by Monster
and coworkers. In the linear pharmacokinetic range, these parameters would
remain proportional to the exposure concentration down the dose-response
curve.
5.1.1.2 Percutaneous--Under most circumstances of use, absorption of PCE
through the skin is of minor consideration. Stewart and Dodd (1964) observed
absorption in each of five individuals after one thumb of each was immersed in
003PE1/E 5-2 11/22/83
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TABLE 5-1. ESTIMATED UPTAKE OF SIX INDIVIDUALS EXPOSED TO
TETRACHLOROETHYLENE WHILE AT REST AND AFTER REST/EXERCISE
Uptake, according to TCE exposure
concentration
142 ppm Lean body Minute volume
72 ppm 144 ppm (rest and Body mass,, mass, at rest,
Subject (at rest) (at rest) exercise) kg kg 1/min
A
B
C
D
E
F
370
490
530
500
390
450
670
940
1000
1210
880
970
1060
1500
1400
1510
1320
1120
70
82
82
86
67
77
62
71
71
74
61
61
7.6
11.6
10.0
11.3
12.3
8.8
Source: Monster, 1979.
a beaker of PCE, which was located in a ventilated hood. By measuring concen-
trations in alveolar breath samples, the investigators concluded that there is
little likelihood that toxic amounts of PCE will be absorbed through the skin
during normal use or exposure to the compound.
Riihimaki and Pfaffli (1978) concluded that PCE concentrations found in
the work environment were not likely to result in a significant amount of PCE
being absorbed through the skin. In their experiment, three individuals who
wore full facepiece respirators to prevent pulmonary absorption and were dressed
in thin cotton pajamas and socks were exposed to 600 ppm PCE (4069 mg/m3) for
3.5 hours. During each midhour, for a period of 10 minutes, each person exer-
cised on a bicycle ergometer. The periodic exercise was employed to simulate
work conditions. Tidal and alveolar air (mixed) were collected in polyester-
lined polyethylene bags. Concentrations of PCE in blood and in exhaled air
were determined up to 20 and 50 hours, respectively. The amount absorbed
(assuming 98 percent is exhaled) was calculated at 7 ppm PCE (47.6 mg/m3).
Jakobsen et al. (1982) found that PCE was absorbed through the skin of
anesthetized guinea pigs after epicutaneous exposure. A maximum blood concen-
tration was observed after 1 hour, followed by a rapid decrease during the
003PE1/E
5-3
11/22/83
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rest of the exposure time. PCE was applied to a clipped area of the skin and
sealed under a glass ring.
Kronevi et al. (1981) found that PCE, when applied to the skin of guinea
pigs, caused considerable local skin changes within 15 minutes of topical
exposure. Jakobson et al. (1982) suggested that these skin changes might
reflect a defense reaction against penetrating agents.
5.1.2 Distribution
Tetrachloroethylene is believed to partition in body tissues having a
high lipid content (Stewart, 1969). This site probably accounts for the
prolonged retention of PCE. Such retention in adipose tissue is suggested by
the blood/air and fat/blood partition coefficients of 16 and 90, respectively
(Monster, 1979). However, the time needed to saturate adipose tissue to 50
percent of its equilibrium concentration is high, about 25 hours.
In an abstract, ,Hake et al. (1976) reported that "the solvent was rapidly
absorbed and excreted via the lungs with a small portion accumulating in the
body which was slowly excreted." Male and female volunteers were exposed for
1.3, 5.5, and 7.5 hr/day to 25, 50, 100, and 150 ppm PCE (170, 339, 678, and
1017 mg/m3).
Because of selective partitioning of PCE in adipose tissue, a long period
(about 2 weeks) is necessary to eliminate PCE from the body completely upon
cessation of exposure (Monster et al., 1979; Fernandez et al., 1976). Uptake
of PCE by adipose tissue will take place during repeated exposures until an
equilibrium is reached; further exposure will not result in further accumulation
(Schumann et al., 1980).
Guberan and Fernandez (1974) used a mathematical model to calculate
uptake and distribution of PCE in the body, and they predicted that fatty
tissues would show the slowest rate of PCE depletion because of the high
solubility of PCE in fatty tissues. Serial breath concentration decay data
obtained from 25 volunteers exposed to between 50 and 150 ppm PCE (339 and
1017 mg/m3) for up to 8 hours were used in developing the model. As shown in
Figure 5-1, theoretical curves of concentration in alveolar air (C , ) divided
by the concentration in inspired air (C. ) versus exposure time for various
post-exposure times can be used to estimate unknown concentrations to which an
individual may be exposed.
Savolainen et al. (1977) found that exposure of two strains of rats to
200 ppm PCE (13,560 mg/m3) for 6 hr/day for 4 days resulted in partitioning of
003PE1/E 5-4 11/22/83
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a
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o"
>
•3
o
345678
DURATION OF EXPOSURE, hours
Figure 5-1.
003PE1/E
Predicted post-exposure alveolar air concentrations of tetrachloro-
ethylene at various times against duration of exposure.
Source: Guberan and Fernandez, 1974.
5-5
11/22/83
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PCE in perirenal fat, brain, and liver. Statistical analyses were not performed.
When behavioral patterns were observed subsequent to the last exposure, ambula-
tion frequency of exposed rats increased (p <0.05) compared to controls.
Rearing frequency, defecation and urination frequency, and preening frequency
were not affected by exposure.
5.1.3 Metabolism (Animal Studies)
The hepatotoxic, carcinogenic, and mutagenic potentials of a number of
chlorinated ethylene compounds (U.S. Department of Health, Education, and
Welfare, 1977a,b,c; Viola et al., 1971; Maltoni and Lefemine, 1974; Creech and
Johnson, 1974; Waxweiler et al., 1976) have generated considerable interest in
the metabolic pathways of these compounds. Certain relatively inert chemicals
may be activated by biotransformation to carcinogenic intermediate metabolites
that could have the potential to induce a carcinogenic lesion. Thus, the
relationship of the metabolism of the various chlorinated ethylenes, including
PCE, to their toxicity is an important consideration.
The cytochrome P-450 dependent mixed-function oxidases of mammalian liver
microsomes have been demonstrated to oxidize the carbon-carbon double bond in
olefins to an epoxide ring (Liebman and Ortiz, 1968; Watabe and Maynert, 1968;
Liebman and Ortiz, 1970; Maynert et al., 1970). The stability of the epoxide
ring varies, depending on the configurations of the oxirane compound. Thus,
this activated intermediate metabolite may interact covalently with a variety
of cellular macromolecules. If the chemical interacts with nucleic acids
and/or proteins that are essential to cellular function, the reaction may
result in alteration of cellular metabolism, resulting in cellular necrosis,
or in carcinogenic or mutagenic lesions (Jerina and Daly, 1974).
The formation of an epoxide intermediate for a chloroethylene compound
was originally postulated by Powell in 1945. Later, Yllner (1961) and Daniel
(1963) speculated that PCE might be oxidized to an epoxide as an intermediate
metabolite during its biotransformation. Recent interest in this hypothesis
has resulted from findings that vinyl chloride is carcinogenic in man and
animals (Viola et al., 1971; Creech and Johnson, 1974; Maltoni and Lefemine,
1974; Lee and Harry, 1974) and the observation that this carcinogenicity is
probably caused by formation of an epoxide intermediate, chloroethylene oxide.
This mechanism was proposed by van Duuren in 1975. Tetrachloroethylene epoxide,
as well as other chloroethylene epoxides, have been synthesized i_n vitro
003PE1/E 5-6 11/22/83
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(Kline et al., 1978; Kline and van Duuren, 1977; and Derkosch, 1976). These
epoxides are unstable (have short half-lives) and are highly reactive.
14
Yllner (1961) studied the metabolism of C-labeled PCE in mice exposed
for 2 hours by inhalation to doses of 1.3 mg/g. Seventy percent of the ab-
sorbed radioactivity was expired in air, 20 percent was excreted in the urine,
and less than 0.5 percent was eliminated in the feces. Of the total amount of
radioactivity excreted in the urine, 52 percent was identified as trichloroace-
tic acid, 11 percent as oxalic acid, and a trace was dichloroacetic acid. No
radiolabeled monochloroacetic acid, formic acid, or trichloroethanol was
detected. Of the remaining radioactivity in the urine, 18 percent was not
extractable with ether, even after hydrolysis.
14
Daniel (1963) fed C-labeled PCE to rats and found that excretion was
largely of unchanged compound through the lungs (half-time of expiration was 8
hours). Only 2 percent of the radioactivity was excreted in the urine, and
equimolar proportions of trichloroacetic acid and inorganic chloride were the
only metabolites detected.
Other investigators (Ogata et al., 1971; Ikeda and Ohtsuji, 1972; Ikeda,
1977; Moslen et al. , 1977; Dmitrieva, 1967; Tada and Nakaaki, 1970; Boillat,
1970; Barchet et al., 1972) have reported that trichloroacetic acid (TCA) is a
urinary metabolite of PCE in experimental animals and humans.
The most likely pathway for such product formation would be via epoxida-
tion of the double bond. The resulting chloro-oxirane compound is known to be
unstable and rearranges spontaneously. However, the stability of symmetric
oxiranes such as the one formed from PCE is greater than that of the asymmetric
oxiranes such as those formed from vinyl chloride, vinylidene chloride, and
trichloroethylene. Henschler and his colleagues (Bonse et al., 1975; Greim et
al. , 1975; Henschler et al., 1975; Bonse and Henschler, 1976) have studied the
chemical reactivity, metabolism, and mutagenicity of the chlorinated ethylene
series, including vinyl chloride, trichloroethylene, and tetrachloroethylene.
These investigators have reported a correlation between biological activity and
chemical structure: those chlorinated ethylenes that are symmetrical, such as
cis- and trans-l,2-dichloroethylene and PCE, are relatively stable and not
mutagenic. In contrast, the asymmetrical ethylenes, vinyl chloride, vinylidene
chloride, and trichloroethylene, are unstable and mutagenic. Although these
investigators recognized that oxiranes (epoxides) may be formed by all six of
the chlorinated ethylenes, they concluded that the asymmetrical oxiranes are
003PE1/E 5-7 11/22/83
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far less stable than the symmetrical ones, and therefore are more highly elec-
trophilic and may react directly with nucleophilic constituents of cells more
readily, thereby exerting mutagenic or carcinogenic effects. The results of
the mutagenic tests conducted by these investigators correlate with this
structure-activity relationship.
Evidence for the involvement of the microsomal mixed-function oxidase
system in the metabolism of PCE was shown by Moslen et al. (1977). Rats that
were pretreated with phenobarbital or Arochlor-1254 (polychlorinated biphenyls),
which are inducers of the hepatic mixed-function oxidase system, showed a
significant increase in total trichlorinated urinary metabolites and TCA
excretion following a single oral administration of 0.75 ml/kg PCE. Hepato-
toxicity of PCE was enhanced by Arochlor-1254 pretreatment, as evidenced by
doubling of serum glutamic-oxaloacetic transaminase (SCOT) levels, and by the
appearance of focal areas of vacuolar degeneration and necrosis of the liver.
However, Cornish et al. (1973) did not observe a potentiation of PCE toxicity
following intraperitoneal injection of 0.3 to 2.0 ml/kg PCE to rats pretreated
with phenobarbital. (The LD50 for the mouse is 2.9 ml/kg [Klaasen and Plaa,
1966]). However, elevation of SCOT was noted at all dose levels in this
study.
Costa and Ivanetich's recent j_n vitro studies (1980) concerning the
extent of metabolism of PCE provide further evidence of the central role of
hepatic microsomal cytochrome P-450 in the metabolism of PCE. Two distinct
forms of P-450, one induced j_n vivo in rats by phenobarbital and the other
induced by pregnenolone-16«-carbonitrile, were shown to be active in the i_n
vitro binding and metabolism of PCE. P-448, induced by B-napthoflavone, was
not active in the j_n vitro system.
Spectral binding of PCE to P-450 in microsomes was measured by the differ-
ence in absorbance between 386 nm and 418 nm. Metabolism was estimated j_n
vitro by the maximal rate of production (V ) of TCA per minute per nmole of
max
P-450. The spectral binding constant (K ) associated with P-450 from uninduced
rats was 0.4 mM. Phenobarbital-induced P-450 had no effect on K , but P-450
from pregnenolone~16«-carbonitrile-induced rats increased K .
V was significantly increased (P <0.01) increased above controls
max
(uninduced rats) when the microsomal preparation was derived from either of
the inducers. Phenobarbital induction resulted in a lowering of the Michaelis-
Menton (K ) constant, indicating an increased affinity for the associated
003PE1/E 5-8 11/22/83
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enzymes. The K was increased when the microsomes were derived from pregneno-
lone-16<*-carbonitrile induction. The P-448 inducer, p-napthoflavone, had no
effect on any of these parameters.
The rate of carbon monoxide-inhibitable NADPH oxidation by PCE was ele-
vated significantly (P< 0.01) above noninduced controls by phenobarbital-
induced microsomes. Induction by pregnenolone-16«-carbonitrile was without
effect in this system. Metyrapone, a specific inhibitor of P-450, reduced
spectral binding, NADPH oxidation, and metabolism by PCE 80, 65, and 75
percent, respectively. SKF-525A reduced spectral binding and metabolism 19
and 33 percent, respectively. The microsomal preparation used in the system
was derived from phenobarbital-treated rats. The major metabolite in the j_n
vitro system was observed to be unbound TCA. Trichloroethanol was not found
in appreciable amounts. Levels of P-450, as well as cytochrome b5, heme,
NADPH-cytochrome reductase, and glucose-6-phosphatase, were not reported to be
altered, whether the inducer was phenobarbital or pregnenolone-16<*-carbonitrile.
Kaemmerer et al. (1982) reported that feeding rats 25 ppm PCE in feed
resulted in a statistically significant induction of cytochrome P-450 in rat
liver homogenates 14 days after administration, compared to untreated controls.
When 500 ppm was used, the increase occurred after 7 days. The authors also
reported that blood coagulation was impaired in rats. The authors concluded
that PCE interferes with basic physiologic functions, particularly cell meta-
bolism.
Takano and Miyazaki (1982) reported that the PCE inhibited glutamate,
succinate, and malate oxidation in the presence of rat liver mitochondria. The
lesser inhibition of succinate oxidation suggested to the authors that PCE may
act as an uncoupler in the electron transport process between NADH and coenzyme Q.
Ogata and Hasegawa (1981) have reported similar findings for PCE and other
halocarbons in succinate oxidation by rat liver mitochondria.
The extent of metabolism of PCE and its binding to hepatic macromolecules
in mice (B6C3F1) and rats (Sprague-Dawley) was investigated by Schumann et al.
(1980). Animals were exposed for 6 hours via inhalation to 10 and 600 ppm
14C-labeled PCE (67.8 and 4068 mg/m^). Purity was determined to be greater
than 98 percent. Following inhalation of 10 ppm (67.8 mg/m3) 14C-labeled PCE
by mice, urinary metabolites were observed to account for 62.5 percent of the
total radioactivity recovered, though unchanged PCE in expired air accounted for
only 12.0 ±1.3 percent. The reverse situation was observed after a single
003PE1/E 5-9 11/22/83
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oral dose of 500 mg 14C-labeled PCE/kg body weight: 82.6 percent of the total
recovered radioactivity was in the form of unchanged PCE in expired air in the
mouse. This shift in the mouse in the major route of elimination demonstrates
a saturation of oxidative metabolism. Metabolites accounted for 10.3 percent.
When compared to the extent of metabolism in rats (Pegg et al., 1979), it was
calculated that the mouse metabolized 8.5 and 1.6 times more PCE than rats at
10 ppm (67.8 mg/m3) and 500 mg/kg, respectively.
Following exposure of rats and mice to 10 and 600 ppm 14C-labeled PCE
(67.8 and 4068 mg/m3), no radioactivity was observed to be bound to purified
hepatic DMA. The detection limit was 14.5 alkylations/106 nucleotides at the
600 ppm (4068 mg/m3) level. Lack of binding to DMA also was observed following
a single oral dose of 500 mg/kg 14C-labeled PCE. However, at times of peak
binding to hepatic macromolecules, 4.7-, 5.8-, and 4.4-fold more radioactivity
was irreversibly bound in the mouse than in the rat following exposure to 10
or 600 ppm (67.8 or 4068 mg/m3), or 500 mg/kg, respectively. The extent of
hepatic macromolecular bindings was determined subsequent to homogenization of
the liver. The authors noted that the 60-fold increase in exposure level
resulted in a 10-fold increase in the extent of binding in both rats and mice.
This result suggests that metabolism of PCE is a saturable process. Binding
was determined at intervals up to 72 hours post-exposure. The lack of binding
to DNA was suggested by the authors as evidence indicating that PCE lacks
genotoxic potential. An absence of binding to DNA does not preclude such an
event that occurs below the detection limit of alkylation. Effects of treatment
on body weight and liver are reported in Chapter 6. A briefer discussion of
the results of Schumann and coworkers was presented by Watanabe et al. (1980).
A low rate of metabolism (2.8 mmole/kg) of PCE was reported by Bolt et
al. (1982) during exposure of newborn female Wistar rats for 10 weeks at 2000
ppm (13,560 mg/m3). The reported metabolic rate was 7 umole/hr/kg, a rate
30-fold less than that reported for trichloroethylene.
In the oral and inhalation exposure of male Sprague-Dawley rats, Pegg et
al. (1979) found that by either route, PCE was eliminated in unchanged form,
predominantly in expired air. In contradiction to evidence pertaining to
humans and other species, Pegg and coworkers found oxalic acid to be the major
urinary metabolite. TCA was not detected.
The protocol was similar to that used by Schumann et al. (1980) in ex-
posures of mice and rats. Rats were exposed via inhalation (6 hours) to 10
003PE1/E 5-10 11/22/83
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and 600 ppm 14C-labeled PCE (67.8 and 4068 mg/nv*). In the gavage experiment,
1 or 500 mg 14Olabeled PCE/kg body weight was administered as a single dose.
After oral administration (500 mg/kg) or inhalation (600 ppm; 4068 mg/m^),
89 percent of the radioactivity was recovered in expired air as PCE; 1 to 2
percent remained in the carcass. The difference was accounted for by 14carbon
dioxide (C02) and urinary and fecal metabolites. Pulmonary elimination was
observed to be monophasic with a half-life of about 7 hours; it was independent
of dose or route of administration. Radioactivity remaining in the carcass 72
hours after exposure by either route was primarily distributed in liver,
kidney, and fat tissue. After oral administration of 1 mg/kg, only 72 percent
of the radioactivity was recovered in expired air as PCE. Recovery of 14CO;j
and nonvolatile metabolites increased. The authors suggested that, in rats,
metabolism of PCE is a saturable process.
In blood, PCE levels peaked 1 hour after oral administration. Disappear-
ance followed apparent first-order kinetics for up to 36 hours after exposure,
with a half-life of 6 hours and an elimination constant of 0.12/hr. After
termination of exposure to 600 ppm (4068 mg/m5), PCE levels in blood declined
similarly with a half-life of 7 hours and an elimination constant of 0.10/hr.
Whole liver radioactivity and that associated irreversibly with tissue
macromolecules was assayed 0, 6, 24, and 72 hours after termination of exposure
to 10 or 600 ppm 14C-labeled PCE (67.8 or 4068 mg/m3) and 72 hours after oral
administration. Whole-liver radioactivity in livers from inhalation-exposed
animals decreased exponentially during the 72-hour collection period as a
biphasic process. The ratio of bound to metabolized radioactivity 72 hours
after oral or inhalation exposure was not significantly different between the
high and the low dose.
Nonprotein liver sulfhydryl levels were not found to be depleted, thus
suggesting to Pegg et al. that glutathione was not involved in metabolism.
The low rate of turnover of nonextractable radioactivity in liver macromolecules
also suggested that accumulation of bound PCE metabolites could potentially
occur upon repeated exposure. Such "accumulation" (or preferably, extent of
binding) would be expected to plateau as the rate of uptake, elimination,
metabolism, and protein turnover reached equilibrium.
Vainio et al. (1976) looked at the effects of PCE on liver metabolizing
enzymes jjri vivo in the rat. Oral administration of 2.6 mmol/kg PCE was associ-
ated with a statistically significant lowering of concentrations of 3,4-benzpy-
rene hydroxylation and p-nitroanisole-o-methylation.
003PE1/E 5-11 11/22/83
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Plevova et al. (1975) showed that 6 hours of exposure to 177 ppm PCE
(12,000 mg/m3) 20 hours prior to administration of 44 mg/kg intraperitoneal.
pentobarbital sodium would lengthen pentobarbital sleeping time by 30 percent
in Wistar rats. This effect was possibly mediated through hepatic drug metabo-
lizing enzyme activity. Also, changes in spontaneous motor activity induced
by intraperitoneal injection of pentobarbital, diazepam, amphetamine, and
partly by chlorpromazine were enhanced by previous inhalation of PCE. This
response probably resulted from an effect on metabolism rates.
Although TCA has been observed by several investigators to be a urinary
metabolite of PCE, the excretion of total trichloro compounds, as measured by
the nonspecific Fujiwara colorimetric reaction after oxidation, exceeded that
of TCA. In some cases, this excretion was assumed to be trichloroethanol
(Muenzer and Heder, 1972; Ikeda et al., 1972). In other studies, that portion
that was not TCA could not be demonstrated to be trichloroethanol (Hake and
Stewart, 1977). In one report, ethylene glycol was claimed to be a prominent
metabolite in the rat (Moslen et al., 1977).
Leibman and Ortiz (1975, 1977) proposed a scheme (see below) for possible
pathways of PCE metabolism. The formation of PCE epoxide by the hepatic mixed-
function oxidase system may be followed by hydration of the epoxide to PCE glycol
Because of the symmetric arrangement of the epoxide and glycol intermediates,
rearrangement of both would yield trichloroacetyl chloride, which hydrolyzes
rapidly to TCA and, to a much lesser extent, TCE in the urine. Possibly, this
response could be caused by dehydrohalogenation to oxalic acid, which has been
reported by Pegg et al. (1979) and Yllner (1961).
C1C = CC12
C10C CC19—>• C1..C-COC1 -*• Cl-jC-COOH
c. \^ j L- 3 3
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Incubation of PCE and rat liver supernatant with a nicotinamide adenine
dinucleotide phosphate (NADPH), reduced generating system confirmed the produc-
tion of TCA. Nicotinamide adenine dinucleotide, reduced (NADH), did not
003PE1/E 5-12 11/22/83
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promote the formation of TCA. Expoxide hydrase inhibition, produced by the
addition of cyclohexane to the incubation mixture, did not have any effect
upon TCA formation. Leibman and Ortiz (1977) concluded that, if the epoxide-
diol pathway is operative, tetrachloroethylene oxide is not a substrate for
hydration by epoxide hydrase, or that the epoxide and glycol rearrange to
trichloroacetyl chloride at similar rates.
5.1.4 Excretion and Elimination
PCE is removed from the body in two principal ways: elimination as an
unchanged compound in exhaled air and through elimination of its metabolites.
During exhalation, the concentration of PCE in expired air is a function of a
number of factors: (1) duration of exposure (until steady-state is reached)
and the concentration in inhaled air, (2) rate of respiration, (3) time elapsed
following exposure, and (4) total body lipid and other tissue repositories.
The principal approach used to measure the amount and kinetics of elimination
via the lungs is serial breath analysis of alveolar air.
From controlled exposure studies, Stewart and coworkers (Stewart, 1969;
Hake et al., 1976; Stewart et al., 1970; Stewart et al., 1961a,b; Stewart et
al., 1977; Stewart et al., 1974) concluded that PCE is rapidly excreted from
the lungs and is principally excreted in unchanged form. These findings have
recently been confirmed in the controlled exposure studies of Monster et al.
(1979), who used serial breath analysis and found that between 80 and 100
percent of PCE is eliminated unchanged via the lungs.
The principal urinary metabolite of PCE is TCA, although trichloroethanol
(TCE) has been reported by the nonspecific Fujiwara reaction to be a secondary
metabolite. Other minor metabolites reported include oxalic acid, dichloro-
acetic acid, and ethylene glycol.
5.1.4.1 Controlled Studies—The concentration in alveolar air in the most
immediate post-exposure period (up to 2 hours) is a reflection of the PCE
concentration to which the individual was most recently exposed (Stewart et
al., 1970; Stewart 1961b). The breath decay curves shown in Figure 5-2 were
obtained from five males experimentally exposed to an average PCE concentra-
tion of 101 ppm (685 mg/m3) for 7 hd/day on 5 consecutive days. The curves
show that a high percentage of absorbed PCE was excreted during the 2-week
period following exposure. A single 7-hour exposure of 15 volunteers to an
average PCE concentration of 101 ppm (685 mg/m3) resulted in a similar alveolar
003PE1/E 5-13 11/22/83
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desaturation curve (Stewart et al., 1970). The range of concentrations from
which the average was obtained was 62 to 137 ppm (420 to 929 mg/m3). The
increase of initial alveolar air concentration after repeated exposure (Figures
5-2 and 5-3) was suggested by the authors to be a result of the accumulation
of PCE in the body tissues. However, as shown by Schumann et al. (1980), an
equilibrium is rapidly reached where the body burden of PCE would not increase
to a value greater than that of the previous day (Figure 5-2). The "hump" in
the curve (Figure 5-3) is unexplained but was not considered by Stewart and
co-workers to be artifactual. Breath samples, collected in glass pipettes and
Saran bags, were analyzed by infrared spectroscopy.
After an exposure of six volunteers to approximately 100 and 200 ppm PCE
(678 and 1356 mg/m3), analyses of the breath decay curves indicated that (1)
exposures of similar duration yielded decay curves with similar elimination rate
constants, (2) the average concentration in the expired air was proportional
to the vapor concentration for exposures of similar duration, and (3) the
length of time that PCE could be measured in expired air was proportional to
both the vapor concentration and the duration of exposure, i.e., it is pro-
portional to the acquired body burden, which would plateau upon repeated
exposure (Stewart et al., 1961a).
The concentration of PCE in the blood of those individuals exposed to 194
ppm (1316 mg/m3) for 83 and 187 minutes approached a plateau near the end of
the third hour of exposure. After exposure, PCE was rapidly cleared from the
blood and was undetectable after 30 minutes.
Monster et al. (1979), in agreement with the general findings of Stewart
and co-workers (Stewart, 1969; Hake et al., 1976; Stewart et al., 1970; Stewart
et al., 1961a,b; Stewart et al., 1977; Stewart et al., 1974), found that 80 to
100 percent of the PCE absorbed was excreted unchanged; metabolism to urinary
TCA accounted for less than 2 percent. Physical exercise resulted in increased
uptake and an increase of PCE in blood levels; similar observations were made
by Stewart et al. (1974).
In the study by Monster et al. (1979), discussed previously in Section
5.1.1.1, the concentration of PCE in exhaled air decreased when exposed indivi-
duals exercised. Recovery in exhaled air was 78 percent (exercise) as compared
to 92 percent when the individuals were at rest. Exercise had no effect on
the half-life of PCE elimination or on the rate constant of elimination.
Minute volume and lung clearance were three-fold higher than the values obtained
003PE1/E 5-14 11/22/83
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Mean and range of breath concentrations of tetrachloroethylene
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003PE1/E
5-16
11/22/83
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when the individuals were at rest. While exercise decreased the amount of PCE
in exhaled air, uptake increased about 40 percent (Figure 5-1).
The concentrations of PCE in blood and in exhaled air during the post-
exposure period are shown in Figure 5-4. Contrary to the finding of Stewart
et al. (1961a) that blood concentrations of PCE were undetectable 30 minutes
after exposure, Monster et al. (1979) found that the decrease of PCE in the
blood paralleled the decay in expired air. The slopes of the curves in Figure
5-3 suggest that the half-lives of PCE in exhaled air and blood for three body
compartments are for (1) tissues with high blood flow, (2) lean tissue, and (3)
adipose tissue. The concentration of TCA (from metabolism) in the blood increased
for 20 hours post-exposure before declining. Monster (1979) suggested that the
increase in TCA concentration is smaller and that the maximum is reached earlier
in contrast to exposures to trichloroethylene or methyl chloroform. The biolo-
gical half-life of TCA in blood is the same for all solvents (about 80 to 100
hours). Upon repeated exposure to PCE, TCA would increase (Monster et al., 1979).
However, TCA would be expected to eventually plateau and not increase beyond
steady-state.
The time course of the PCE concentrations in blood and in expired air
indicates that a long period (greater than 275 hours) is necessary to complete-
ly eliminate PCE from the body (Fernandez et al., 1976; Monster et al., 1979).
Fernandez et al. (1976) found that 2 weeks were necessary to completely eliminate
PCE from the body after exposure to 100 ppm (678 mg/m^) for 8 hours. These
findings agree with those of Monster et al. (1979). In these chamber studies,
24 volunteers were exposed for 1 to 8 hours to vapor containing 100, 150, and
200 ppm PCE (678, 1017, and 1356 mg/m3). When the exposure time increased,
the PCE concentration in alveolar air would plateau at steady-state.
Monster (1979) determined that the values for partition coefficients and
lung clearance measurements between blood and vapor after exposure of four to
six individuals to 70 to 140 ppm PCE (475 and 950 mg/m3) for 4 hours indicated
that (1) alveolar air concentration of PCE in the first few hours after exposure
will be proportional to exposure concentration and to the concentration in
blood and other rapidly exchangeable tissues, and (2) during that later phase
of elimination, the alveolar air concentration will be proportional to the
concentration in adipose tissue. As previously stated, the partition coeffici-
ent (37°C) for PCE between venous blood and alveolar air of 16 and the partition
coefficient between fat and blood of 90 suggests that adipose tissue is a
003PE1/E 5-17 11/22/83
-------
§8
< rJ
I ffl
X Z
UJ
1000
100
10
1
0.1
0.01
= A
I I I I I I I I I I I I I
O 72 ppm PCE AT REST
A 144 ppm PCE AT REST
D 142 ppm PCE AT REST AND WORKLOAD
EXHALED AIR
-I
50 100
POST EXPOSURE, houn
150
Figure 5-4. Tetrachloroethylene in blood and exhaled air following exposure
to tetrachloroethylene for 4 hours. Each point represents the
geometric mean ± standard deviation of six individuals.
Source: Monster et al. , 1979.
003PE1/E
5-18
11/22/83
-------
primary storage site. Monster estimated that 25 hours would be necessary for
PCE to saturate adipose tissue to 50 percent of its equilibrium concentration
with plasma.
Metabolic considerations investigated in this study are discussed in
Sections 5.1.3 and 5.1.4.
Evidence that blood levels of PCE may be useful in determining individual
uptake was obtained in a single exposure to 70 and 140 ppm (475 and 950 mg/m3)
for 4 hours (Monster and Houtkooper, 1979). Concentrations of PCE in blood,
urine, and exhaled air were determined at 2 and 20 hours after exposure.
While a high degree of correlation was obtained between exhaled air concentra-
tion and blood concentration, linear and multiple linear regression analysis
showed an inter-person coefficient of variation of 20 to 25 percent for blood
measurements at 2 and 20 hours and in exhaled air at 2 hours.
Verberk and Scheffers (1980) suggest that measurements of PCE in expired
air are useful in monitoring exposure in the general population as well as in
populations of occupationally exposed people. PCE was measured in the expired
air of residents living near 12 dry-cleaning shops. Residents were classified
into two groups, corresponding to the distance they lived from the shops. In
all cases, the residents living nearest the shop, the house next to the shop,
had higher levels of PCE in expired air than residents who lived one house
farther away from the shop.
In the study by Stewart and Dodd (1964), elimination of PCE via the lungs
following percutaneous absorption was shown to be minor. Each of five indi-
viduals immersed one thumb in a beaker of PCE located in a ventilated hood. At
intervals of 10 minutes, the concentration of PCE in exhaled air was measured.
Before and periodically during each skin exposure, samples of breathing zone
air were analyzed to preclude solvent vapor contamination. The mean peak
breath concentration after a 40-minute immersion was 0.31 ppm (2.1 mg/m3); 2
hours after exposure, the mean breath concentration was 0.23 ppm (1.6 mg/m3).
Five hours after exposure, PCE was still detectable (0.16 to 0.26 ppm; 1.1 to
1.8 mg/m3).
The elimination of PCE through the skin was•found by Bolanowska and
Golacka (1972) to be approximately 0.02 percent per hour of the dose inhaled.
As previously mentioned, in both controlled and occupational exposures of
humans to PCE, the principal urinary excretion product is TCA. Trichloroethanol
has been reported by a nonspecific assay as a metabolite. TCE was indirectly
measured by chromate oxidation of urine to TCA.
003PE1/E 5-19 11/22/83
-------
Hake and Stewart (1977) found only traces of TCA in 24-hour urine specimens
from individuals exposed to 150 ppm PCE (1017 mg/m3) and below. No TCE was
detected.
Ogata et al. (1971) found TCA in the urine of four individuals exposed to
87 ppm PCE (590 mg/m3) for 3 hours at a level of 1.8 percent of the total PCE
dose. TCE could not be detected, but the urine contained 1 percent of an
unidentified chlorine-containing compound. Urine was collected for 67 hours
into the post-exposure period.
In the previously discussed studies of Monster and co-workers (Monster et
al., 1979; Monster and Houtkooper, 1979), urinary TCA was found to represent
less than 1 percent of the absorbed dose of PCE. In blood, TCA continued to
increase until 20 hours post-exposure. From about 60 hours after exposure,
the concentration decreased exponentially. A base level of 0.6 mg of TCA per
day was found in the urine of subjects prior to exposure. Results of blood
and urine concentrations are shown in Figures 5-5 and 5-6. The ratio of TCA .
3 urine
to TCA. , . was three-fold higher in the period 0 to 22 hours after the start
blood
of the exposure. The relatively high concentration in urine possibly was
caused by an unknown compound measured by the non-specific Fujiwara reaction;
TCA in blood was measured by gas chromatography. The unknown compound was not
PCE or TCE.
Exposure combined with exercise resulted in 20 percent higher levels of
excreted TCA, while uptake of PCE increased 40 percent. The TCA concentration
at 20 hours after exposure was 1.6 times the concentration at the end of
exposure. Inferences drawn from these results regarding metabolism of PCE are
discussed in Section 5.1.3. Monster and Houtkooper (1979) concluded that TCA
is not a reliable indicator of exposure to PCE.
5.1.4.2 Occupational Studies — In a study involving six dry-cleaning-plant
workers exposed to PCE, an increase in urinary TCA was observed over the
50-hour sampling period (National Institute of Occupational Safety and Health,
1974). A control group not exposed to significant quantities of PCE also
evidenced a similar increase. The average length of exposure for these indivi-
duals was 17 months. The worker evidencing the highest level of TCA in the
urine had been exposed to an 8-hour time-weighted average of 168 to 171 ppm
PCE (1139 to 1160 mg/m3).
TCA and TCE (specific detection method not discussed) were found in the
urine of 40 workers exposed to PCE concentrations ranging from 58 to 134 ppm
003PE1/E 5-20 11/22/83
-------
1 -
I 0.6 H
I
0.4 —
§ 0.2 H
ffi
0.1 —
O 72 ppm PCE AT REST
4144 ppm PCE AT REST
0142 ppm PCE AT REST AND WORKLOAD
I I I I I I I I I I I I I I I |
50 100 150
TIME AFTER EXPOSURE, houn
Figure 5-5. Trichloroacetic acid in blood following exposure to tetrachloro-
ethylene for 4 hours. Each point represents the geometric mean
± the standard deviation of six subjects.
Source: Monster et al., 1979).
003PE1/E
5-21
11/22/83
-------
0
1*
< 8
H 6
UJ
Z 4
cc
3 2
ft
1 1 1 1 1
i i
i i
72ppmPCE WppmPCE 142ppmPCE
AT REST AT REST AT REST AND
T WORKLOAD
— —
^^
"" rri
-1 [*,*,
*
¥1
T
^ —
Tji,
0 22 46 70 0 22 46 70 0 22 46 70
TIME AFTER START EXPOSURE, hours
Figure 5-6. Urinary excretion of trichloroacetic acid following exposure to
tetrachloroethylene for 4 hours. Each point represents the geo-
metric mean ± the standard deviation of six subjects.
Source: Monster et al., 1979.
003PE1/E
5-22
11/22/83
-------
(393 to 909 mg/m3) (Medek and Kovarik, 1973). The maximum levels observed
were 41 mg TCA and 116 mg TCE per liter of urine. Seventy-two percent of
these workers reported subjective complaints. No relationship was proposed
between specific complaints and specific exposures or TCA and TCE levels.
Exposed groups were not compared with blind controls.
Muenzer and Heder (1972) reported that urinary TCA was found in 124 of
200 dry-cleaning-plant employees. Seventy-one individuals had more than 10
mg/1. Liver function tests for exposed and unexposed (control) groups were
comparable. The general room air in the work places contained between 200 to
300 ppm PCE (1357 to 2035 mg/m3). An association between workroom air con-
centrations and TCA levels was not made.
Ikeda et al. (1972) reported evidence that TCA and TCE concentrations in
the urine increased in proportion to environmental concentrations of PCE up to
50 ppm (339 mg/m3). In this study, urine samples were collected from 34 male
industrial workers who had been exposed to PCE vapors for 8 h/day for 6 days
per week. Concentrations of PCE in the work place ranged from 10 to 400 ppm
(68 to 2713 mg/m3). The plateau observed in the urinary excretion curve for
TCA suggested to the investigators that the capacity of humans to metabolize
PCE is limited. The maximum level of TCA observed was approximately 50 mg/1
of urine. For TCE, the maximum concentration reported was approximately 25
mg/1. TCE was measured indirectly by oxidation of urine with chromic oxide.
In another study, Ikeda and Ohtsuji (1972) reported a wide variation in
the TCA and TCE levels in urine from occupational ly exposed workers. One
group of four workers had been exposed to a concentration range of 20 to 70
ppm (136 to 475 mg/m3), while 66 workers in another group had been exposed to
between 200 and 400 ppm (1357 and 2713 mg/m3). The urine from the smaller
group contained between 4 and 35 mg TCA and 4 to 20 mg of TCE per liter of
urine. In the larger group, TCA levels were 32 to 97 mg/1 and TCE levels
ranged from 21 to 100 mg/1.
In their most recent study, Ohtsuji et al. (1983) obtained additional
evidence that confirms reports by others that PCE is metabolized to a limited
extent (approximately 2,percent) in humans. Personal monitoring of exposure,
using carbon felt dosimeters, was carried out in two groups of workers (36
males and 25 females). Comparison of urinary total trichlorocompounds (expres-
sed as TCA) with occupational levels of PCE suggested that metabolic capacity
becomes saturated at about 100 ppm (678 mg/m3) in air. Metabolite levels
increased in a quasi-linear fashion up to 100 ppm PCE (678 mg/m3)
003PE1/E 5-23 11/22/83
-------
High levels of TCA (greater than 60 mg/1 of urine) also were reported by
Weiss (1969) and by Haag (1958) in studies of individuals exposed occupational-
iy-
5.1.4.3 Population Studies--Pre1imi nary results reported by Ziglio (1981)
suggest that plasma TCA can be a reliable indicator of the extent of chronic
exposure to PCE. Ziglio chose 24 individuals in Milan, a city in which both
trichloroethylene and PCE were present in tap water at concentrations as high
as 150 to 200 ug/1 (Giovanardi, 1979). Subjects were divided into two groups:
exposed (those who drank water with halocarbon levels > 50 ug/1) and nonexposed
(those who drank water with levels < 1 ug/1). When exposure increased, plasma
TCA levels increased. However, the TCA levels were suggested by Ziglio to be
more indicative of the extent of exposure to trichloroethylene, because this
halocarbon is metabolized to a much greater extent than is PCE.
5.1.5 Estimates of Biological Half-life
Monster et al. (1979) determined from the concentration curves of PCE in
blood and exhaled air after exposure (Figure 5-4) that PCE was eliminated from
the body at three different rate constants with corresponding half-lives of 12
to 16, 30 to 40, and 55 hours, respectively; this result suggested that there
are three major body compartments for PCE. The half-life, derived from the
data of Stewart et al. (1970), was calculated to be 65 hours. The long half-
life does not pertain to the major portion of the exposure but to a small
percentage of the body burden (Hake et al., 1976). TCA in blood was reported
by Monster et al. (1979) to have a predominant half-life (60 hours after
exposure) of 75 to 80 hours.
Ikeda (1977) and Ikeda and Imamura (1973) reported that the mean bio-
logical half-life for PCE urinary metabolites is 144 hours. A possible sex
difference, indicated from exposures of nine males and four females, has not
been confirmed. The estimated biological half-life of PCE stored in adipose
tissue is 71.5 hours (Guberan and Fernandez, 1974).
5.1.6 Interaction of PCE with Other Compounds
Stewart et al. (1976) conducted a study designed to determine the effects
of alcohol and diazepam (Valium ) on 12 individuals exposed to 25 and 100 ppm
PCE (170 and 678 mg/m3) for 5.5 hours. Administration of alcohol to individuals
during exposure to 25 ppm (170 mg/m3) significantly increased blood levels of
003PE1/E 5-24 11/22/83
-------
PCE (p < 0.01), but there was no effect during exposure at 100 ppm (678 mg/m^).
Diazepam and alcohol each raised breath levels of PCE during exposure at 25
ppm (170 mg/m3) but not at 100 ppm (678 mg/m^). Results are shown in Table 5-2.
Neither diazepam nor alcohol changed the effects of PCE as measured by behavioral
and neurological tests.
5.1.7 Summary
Inhalation is the predominant route of exposure for humans, although
contamination of potable water supplies indicates that oral ingestion repre-
sents an additional route of exposure. Percutaneous absorption is of minor
consideration. In all species tested, PCE is rapidly absorbed and distributed
to major body compartments. Because of its high octanol/water and fat/blood
partition coefficients, PCE can be expected to selectively partition into
lipid-rich tissues until an equilibrium is reached between uptake and elimina-
tion, after which additional accumulation would not occur. While most of the
PCE is rapidly excreted, up to 2 weeks or more may be required to completely
eliminate PCE from the human body (Fernandez et al. , 1976). The estimated
biological half-life of PCE in adipose tissue is about 70 hours. PCE would
also be expected to partition into lungs, liver, kidney, spleen, and lean
muscle until equilibrium is reached. The rate of tissue loading in humans,
with a given inspired air concentration, is increased with .physical activity
and with exposure duration.
PCE is postulated to be metabolized via oxidation by the cytochrome P-450
mixed-function oxidase system. An epoxide intermediate, formed during metabo-
lism, is believed to represent a species capable of binding covalently to
tissue macromolecules. Jji vitro experiments (Costa and Ivanetich, 1980) have
shown that a cytochrome P-450 activation system from rats binds and metabolizes
PCE; TCA was identified as the principal reaction product. The _i_n vivo studies
by Pegg et al. (1979) in rats and by Schumann et al. (1980) in mice have
demonstrated irreversible binding of PCE and/or its metabolites to hepatic
macromolecules; binding to DNA, however, was not observed. Differences in
metabolism of PCE in animals and humans, who have shown very little metabolism
of PCE, seem to exist. In view of these metabolic differences between humans
and rodents, little binding of PCE metabolites to human hepatic macromolecules
would be expected.
Pegg et al. (1979) and Schumann et al. (1980) have suggested that the
metabolism of PCE is a saturable, dose-dependent process and that the extent
003PE1/E 5-25 11/22/83
-------
TABLE 5-2. ALCOHOL AND DIA2EPAM EFFECTS UPON TETRACHLOROETHYLENE (PCE) BLOOD AND BREATH CONCENTRATIONS (5.5-HOUR EXPOSURES)
cr
i
re
CT:
PCE in blood, ppm
2 hours into exposure
PCE in
chamber,
ppm
25
100
PCE
alone
1.65
(35)
8.25
(63)
PCE and
alcohol
2.92C
(15)
7.95
(29)
PCE andb
diazepam
1.76
(23)
8.47
(41)
PCE
alone
11.03
(35)
33.2
(68)
PCE in breath, ppm
2 hours into exposure 30
PCE and
alcohol3
12.35d
(15)
32.3
(28)
PCE andb
diazepam
11.72
(23)
35.5
(44)
PCE
alone
6.40
(35)
17.62
(64)
minutes post-exposure
PCE and
alcohol
7.40C
(14)
13.83C
(29)
PCE and
diazepam
6.96d
(22)
17.35
(42)
Alcohol blood levels of 30 to 100 rug percent.
Diazepam blood levels of 7 to 30 mg percent.
cSignificantly different from PCE alone at p <0.1.
Significantly different from PCE alone at p <0.05.
Numbers in parentheses indicate the number of determinations.
Source: Stewart et al., 1977.
-------
of reactive intermediate formation plateaus with increasing dose. This situa-
tion would explain why exposure to mice at 500 mg PCE/kg body weight would
result in a carcinogenic response (see Chapter 9) in the National Cancer Institute
bioassay but not an increase at a level of 1000 mg/kg.
Although TCA appears to be a predominant urinary metabolite in humans and
rodents, metabolism in humans (in the form of urinary TCA) accounts for 2
percent or less of the absorbed dose. The estimated mean biological half-life
for all PCE urinary metabolites in humans is 144 hours. Minor metabolites
reported include TCE, oxalic acid, and ethylene glycol. However, Pegg et al.
(1979) observed oxalic acid to be the principal urinary metabolite in rats.
TCA was not detected upon incubation of a rat hepatic microsomal activation
system with PCE. Costa and Ivanetich (1980) observed TCA to be the principal
reaction product.
Elimination of PCE in expired air differs somewhat between humans, mice,
and rats. Monster and coworkers (1979) have shown that humans may exhale as
much as 98 percent of PCE in unchanged form. Exercise increases uptake
(decreases pulmonary elimination) of PCE in the body. Repeated exposures to
levels of 72 ppm (488 mg/m^) or greater during sedentary conditions were
observed to result in more PCE being eliminated unchanged in expired air; this
situation is consistent with saturation of metabolism. At found or expected
ambient air levels of exposure, the fate of PCE would not be altered regardless
of exposure concentration in the linear pharmacokinetic range.
Elimination as the unchanged compound in expired air also is the pre-
dominant process in rats (Pegg et al., 1979), upon inhalation or oral exposure to
PCE. However, in mice, urinary metabolites represent the chief elimination
process at low exposures but not at high exposures (Schumann et al., 1980).
Upon high oral dose exposure of mice, pulmonary elimination of unchanged PCE
predominates (Schumann et al., 1980).
The dose level is a factor of particular importance when one considers
the potential effects of PCE in humans. There appears to be no evidence that
a different metabolite or a distinctively different pharmacokinetic pattern
exists at high, as compared to low, exposure levels. Monster et al. (1979)
has shown that a two-fold increase in exposure level resulted in a doubling of
uptake; a shift in the nature of the urinary metabolites was not observed.
003PE1/E 5-27 11/22/83
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man. Scand. J. Work Environ. Health 4:73-85, 1978.
Savolainen, H., P. Pfaffli, M. Tengen, and H. Vainio. Biochemical and behavioral
effects of inhalation exposure to tetrachloroethylene and dichloromethane.
J. Neuropath. Exp. Neurology 36(6):941-949, 1977.
Schumann, A. M., J. F. Quast, and P. G. Watanabe. The pharmacokinetics and
macromolecular interaction of perchloroethylene in mice and rats as related
to oncogenicity. Toxicol. Appl. Pharmacol. 55:207-219, 1980.
Stewart, R. D. Acute tetrachloroethylene intoxication. J. Am. Med. Assoc.
208(8):1490-1492, 1969.
Stewart, R. D., and H. C. Dodd. Absorption of carbon tetrachloride, tri-
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Stewart, R. D., H. H. Gay, D. S. Erley, C. L. Hake, and A. W. Schaffer. Human
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Stewart, R. D., C. L. Hake, A. Wu, J. Kalbfleisch, P. E. Newton, S. K. Marlow,
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6. TOXIC EFFECTS
6.1 HUMANS
The known effects of tetrachloroethylene (PCE) on humans have been estab-
lished primarily from individuals accidentally or occupationally exposed to
high (in some cases, unknown) concentrations of PCE. Exposure to high concen-
trations of PCE causes a variety of toxicological effects in humans. Effects
upon the central nervous system (CMS) are generally the most noticeable follow-
ing acute or excessive occupational exposures. Effects upon the liver and
kidney usually are observed after an elapsed period of exposure to high concen-
trations.
6.1.1 Effects on the Liver
6.1.1.1 Acute Exposure Situations—Transient, mild hepatitis was diagnosed by
Stewart (1969) in a worker occupationally exposed to high, anesthetic concen-
trations of PCE in a tank car for less than 30 minutes. Infrared analysis of
the patient's exhaled breath 1.5 hours after exposure showed 105 ppm PCE (712
mg/m3). Urinary urobilinogen levels were elevated on the ninth day of the
post-exposure period. The serum glutamic-oxaloacetic transaminase (SCOT)
level showed a slight increase on the third and fourth days. Stewart concluded
this patient had experienced marked depression of the CMS followed by tran-
sient, minimal liver injury. The increase in urinary urobilinogen was suggest-
ed as one indicator of hepatic injury caused by PCE. Elevations in the levels
of urobilinogen and other indicators of liver damage have been reported in
other case studies involving acute exposures (Stewart et al., 1961a; Saland,
1967).
Stewart et al. (1961a) reported a case in which a male construction
worker became semicomatose after being overexposed to fumes from a petroleum-
based solvent, during a 3.5-hour period. The solvent was reported to contain
50 percent Stoddard Solvent and 50 percent PCE. A neurological examination
conducted 1 hour after collapse indicated no abnormality of function; the
liver was not palpable. During a 6-week period subsequent to the incident,
neither clinical jaundice nor neurological symptoms were observed. Simulated
exposure conditions suggested that the estimated average concentration of PCE
in the work environment, during exposure, was 393 ppm (2666 mg/m3). The
average exposure for the first 3 hours was estimated to be 275 ppm (1864 mg/m3).
003PE5/A 6-1 11/22/83
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The exposure for the remaining 30 minutes with the air hose turned off was
1100 ppm (7458 mg/m3). Clinical tests revealed evidence of impaired liver
function, beginning 9 days following exposure; urinary urobilinogen and total
serum bilirubin were elevated. On the fourteenth post-exposure day, urinary
urobilinogen was in the normal range, but a slight elevation of alkaline phos-
phatase was observed. Serum glutamine-pyruvic transaminase (SGPT) and SGOT
values were in the normal range throughout the post-exposure period, with the
exception that on the eighteenth day, SGPT was slightly elevated. On the
sixteenth post-exposure day, the PCE concentration in expired air was sharply
elevated. However, the overall decay curve slope was not altered appreciably.
The investigators suggested that such an acute exposure may represent a contin-
uing insult to the liver in view of the observations that PCE had an exceedingly
long exponential decay in expired air, indicating slow release from body
tissues. On the twenty-first day, no detectable PCE was found on the exposed
worker's breath.
Elevated SGOT values and one enlarged liver were reported by Saland
(1967). Nine firefighters were exposed to an unknown concentration of PCE
upon responding to a complaint about fumes. All the firefighters were exposed
without masks to unknown levels above the odor threshold for 3 minutes. No
irritation of the eyes, mucous membranes, or respiratory tract occurred. For
some inexplicable reason, all the firefighters were admitted to a hospital for
study 12 days after the incident. All clinical tests indicated no abnormali-
ties; however, eight of nine individuals had elevated SGOT levels. In seven
individuals, SGOT values had returned to normal by 22 days after testing.
Hepatomegaly and splenomegaly were found in one individual 12 days post-exposure.
After 63 days post-exposure, the liver was not palpable.
An enlarged liver and obstructive jaundice were diagnosed by Bagnell and
Ellenberger (1977) in a 6-week-old, breast-fed infant. While this situation
is not uncommon in infants, the infant had been indirectly exposed to PCE.
The child's father worked as a leather and suede cleaner in a dry-cleaning
establishment where PCE vapors were present. During regular lunchtime visits
to the exposure site, the mother had been exposed to the same vapors. These
visits lasted between 30 and 60 minutes. The concentration of PCE in the work
place was unknown, although it was believed to be excessively high because of
reported episodes of dizziness. In the infant, bilirubin, SGOT, and serum
alkaline phosphatase were elevated; other blood and urinary parameters of
003PE5/A 6-2 11/22/83
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liver function were normal. Normal liver function was found in both parents.
Analysis of the mother's blood 2 hours after one of her lunchtime visits
indicated a PCE level of 0.3 mg per 100 ml. One hour after a visit, her
breast milk contained 1.0 mg PCE per 100 ml. After 24 hours, the concentration
of PCE in the breast milk decreased to 0.3 mg per 100 ml. Chlorinated hydro-
carbons were not found in the mother's urine. One week after breast feeding
was discontinued, serum bilirubin and serum alkaline phosphatase levels in the
infant returned to a normal range. The findings suggest that the neonatal
liver may be sensitive to toxicological effects of PCE, although other causal
factors in this instance cannot be ruled out.
Levine et al. (1981) reported a diffuse and marked fatty metamorphosis of
the liver of a 53-year-old male dry cleaner who succumbed to an overexposure
of PCE over an unspecified period of time. During autopsy, the PCE concen-
tration in the liver was 240 mg/kg. PCE was also detected in the blood (4.5
mg/1), brain (69 mg/kg), kidney (71 mg/kg), and liver (30 mg/kg). The authors
noted that the liver was extremely fatty and that an ethanol concentration of
0.20 percent (w/v) was found in the blood.
Hake and Stewart (1977) reported mild liver injury, as indicated by
elevated serum enzymes, in a 60-year-old male dry cleaning operator who was
overcome by PCE vapors. The investigators estimated that the individual had
been lying in a pool of solvent for about 12 hours. Although kidney and skin
damage also were evident, the individual was released after 21 days of hospi-
tal ization and appeared fully recovered.
6.1.1.2 Chronic Exposure Situations--Hepatotoxic effects associated with
persons working with unknown concentrations of PCE over extended periods have
been reported by a number of investigators (Coler and Rossmiller, 1953; Franke
and Eggeling, 1969; Hughes, 1954; Trense and Zimmerman, 1969; Meckler and
Phelps, 1966; Larsen et al., 1977; Moeschlin, 1965; Dumortier et al., 1964).
Liver function parameters that may be observed to be altered as a result of
excessive PCE exposure include sulfobromophthalein retention time, thymol
turbidity, serum bilirubin, serum protein patterns, cephalin-cholesterol
flocculation, serum alkaline phosphatase, SGOT, and serum lactic acid de-
hydrogenase (LDH). However, these parameters may be observed as a result of
many different causes that are completely dissociated with PCE.
Effects observed after exposure to PCE at high or unknown levels included
cirrhosis of the liver (Coler and Rossmiller, 1953), toxic hepatitis (Hughes,
003PE5/A 6-3 11/22/83
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1954; Meckler and Phelps, 1966), liver cell necrosis (Trense and Zimmerman,
1969; Meckler and Phelps, 1966), and enlarged liver (Meckler and Phelps,
1966). In some cases, liver dysfunction parameters returned to normal follow-
ing cessation of exposure (Hughes, 1954). In one case, the liver was enlarged
6 months after cessation of exposure (Meckler and Phelps, 1966).
Liver dysfunction was evidenced in a 24-year-old male who was admitted to
the hospital with abdominal pains and blood-tinged vomitus (Larsen et al.,
1977). Three days before hospitalization, this individual was exposed to PCE
vapors emanating from recently dry-cleaned clothes. In this study, SCOT values
increased four to five times above normal 1 day after initial symptoms. Two
days later, SCOT returned to a normal range. The subject was found to have
renal insufficiency and mild proteinuria. Despite diureses, serum creatinine
and blood area increased constantly. Serum bilirubin was normal throughout
hospitalization. Variations in the levels of SCOT and other liver function
parameters during the post-exposure period indicate that repeated testing
during this interval is required for complete diagnosis.
Larsen et al. (1977) also described an incident in which a female dry-
cleaning worker was hospitalized in a comatose state with a grand mal seizure.
Three days prior to hospitalization, she had worn clothing that had been
recently dry-cleaned. Upon her admission, SGOT and serum LDH levels were re-
ported to be strongly elevated; levels gradually decreased during hospitalization.
Blood and protein were found in the urine. Elevated serum creatinine levels
decreased after peritoneal dialysis. A cause-effect relationship is unproven.
Chmielewski et al. (1976) found that the activities of alanine and aspara-
gine aminotransferase were significantly elevated (t „ __ = 2.032) in a group
of 16 of 25 workers compared to non-exposed controls. This group of 16 workers
had been exposed to PCE vapors in the range of 59 to 442 ppm (400 to 3000 mg/m3)
for periods ranging from 2 months to 27 years. Aminotransferase activity in a
group of nine workers who were exposed to levels of PCE at or below 29 ppm (200
mg/m3) was normal. These enzyme imbalances were indicative of liver cell injury
by PCE. Low urinary excretion of 17-ketosteroids were observed in seven of nine
persons whose exposures were less than 29 ppm and 4 of 16 persons whose exposures
ranged from 59 to 442 ppm PCE. Abnormal electroencephalogram (EEC) tracings were
reported in 10 persons. Six people were diagnosed as having a pseudo-neurotic
syndrome, and in four other cases the researchers were led to diagnose or suspect
encephalopathy.
003PE5/A 6-4 11/22/83
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6.1.2 Effects on Kidneys
Diminished excretion of urine (5 to 10 ml of urine per hour), uremia, and
elevated serum creatinine was observed in a woman exposed to PCE vapors ema-
nating from clothes that had been dry-cleaned (Larsen et al., 1977). As pre-
viously described, she had been admitted to the hospital in a comatose state
with a grand mal seizure. During hospitalization, urine excretion and serum
creatinine returned to normal. Renal biopsy suggested toxic nephropathy.
Liver dysfunction also was evidenced by increased SGOT and bilirubin levels
(Section 6.1.1). Protein and blood were detected in the urine. A cause-effect
relationship was unproven.
In another situation in which an individual (male) had worn clothing
permeated with PCE vapors, elevated serum creatinine and blood uremia were
observed (Larsen et al. , 1977). This individual was admitted to the hospital
with abdominal pains and blood-tinged vomitus. Mild proteinuria and leuko-
cytes and erythrocytes were found in the urine. Despite diuresis, serum
creatinine and blood urea increased. Serum creatinine decreased with peritoneal
dialysis, over a 19-day treatment period. Renal biopsy evidenced necrosis in
the renal tubules.
Hake and Stewart (1977) reported renal damage, indicated by a proteinuria
that lasted 20 days and hematuria lasting for 8 days, in a 60-year-old male
dry-cleaning worker who was overcome by PCE vapors. Liver damage and skin
damage also were reported. The individual was reported fully recovered upon
release from the hospital (21 days post-exposure).
6.1.3 Effects on Other Organs/Tissues
6.1.3.1 Effects on the Pulmonary System—A 21-year-old male was found uncon-
scious in a library, presumably from a massive but unknown concentration of
PCE. After his admission to the hospital, the principal clinical feature was
consistent with acute pulmonary edema (Patel et al. , 1977). Bubbling rales
were heard over the entire lung field. Recovery was complete 4 days after
hospital admission. Liver and kidney function tests in this patient were
normal.
Levine et al. (1981) reported pulmonary damage in a 53-year-old male dry-
cleaner who succumbed upon overexposure to PCE vapors. Autopsy findings were
congestion superimposed on mildly fibrotic and diffusely emphysematous lungs
with apical bullous emphysema. The PCE concentration in lung tissue was
003PE5/A 6-5 11/22/83
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30 mg/kg. A co-worker, also overcome by vapors, was reported to have recovered
without complications after hospitalization. The PCE concentration in the
work place air at the time of the incident was not reported.
Alkaline phosphatase in leukocytes is a defense mechanism against bacterial
infection, and it plays an active part in phagocytosis. In an investigation
of the effects of PCE on alkaline phosphatase activity in human neutrophilic
leukocytes, Friborska (1969) found that activity was within the normal range.
In this study of occupational exposure, seven workers were exposed to PCE and
four had been exposed to both PCE and trichloroethylene. For controls, 20
unexposed individuals were used. Trichloroethylene exposure, as opposed to
PCE exposure, raised the activity of the alkaline phosphatase above the control
level. For those individuals exposed to both compounds, no synergistic or
additive effect was observed.
A slight depression in the total white blood cell count of three of nine
firefighters exposed for 3 minutes to unknown concentrations of PCE was reported
by Saland (1967). The white cell count was made 12 days after the firefighters
were exposed to PCE.
6.1.3.2 Effects on the Skin--The effects of PCE upon the skin range from a
mild to moderate burning sensation when direct contact occurs for 5 to 10
minutes, to a marked erythema after prolonged exposure, and finally, blistering
if PCE is trapped under clothing or in shoes (Hake and Stewart, 1977).
Stewart and Dodd (1964) reported that individuals experienced a mild
burning sensation on their thumbs after immersion of the thumbs in a solution
of PCE for 5 to 10 minutes. After the thumbs were withdrawn, burning per-
sisted without a decrease in intensity for 10 minutes before gradually sub-
siding after 1 hour. A marked erythema was present in all cases and subsided
between 1 and 2 hours post-exposure.
Morgan (1969) reported erythema and blistering over 30 percent of the
body of a worker who was anesthetized for 30 minutes after PCE overexposure in
a coin-operated laundry.
Ling and Lindsay (1971) reported severe burns when an individual, upon
losing consciousness, fell into a pool of PCE on the floor. The burns gradu-
ally healed within 3 weeks following exposure.
6.1.4 Behavioral and Neurological Effects
In nearly all the occupational situations involving short-term high-level
exposures to PCE, an initial characteristic response is depression of the CNS.
003PE5/A 6-6 11/22/83
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Subchronic exposures to high levels produce characteristics of a neurasthenic
syndrome; the most frequently reported subjective complaints are dizziness,
headache, nausea, fatigue, and irritation of the eyes, nose, and throat.
These symptoms are typical of many common ailments. Sensitivity may vary
greatly among individuals.
6.1.4.1 Effects of Short-term Exposures—Stewart (1969) reported normal
neurological findings, except for the Romberg test, in an individual who was
found unconscious at the bottom of a tank car. Analysis of the man's breath
1.5 hours after exposure showed approximately 105 ppm PCE (712 mg/m3). The
Romberg test is designed to detect swaying motions when the subject stands
with eyes closed. Upon return to work, the individual reported being very
fatigued after 4 hours of light work. The individual's fatigue with slight
exertion diminished over the 2 days following his return to work. Abnormal
results of the Romberg test are suggested to be the earliest indications of
signs of intoxication caused by PCE.
Rowe et al. (1952) reported that six individuals exposed to an average
PCE concentration of 106 ppm (719 mg/m3) (range = 83 to 130 ppm; 663 to 881
mg/m3) did not evidence CMS effects. At an average PCE concentration of 216
ppm (1465 mg/m3), all four individuals exposed for 45 minutes to 2 hours
experienced slight eye irritation, which developed 20 to 30 minutes into the
exposure period. Minimal, transient eye irritation led the authors to suggest
that the vapor concentration causing this effect in unacclimatized individuals
lies between 100 and 200 ppm (678 and 1356 mg/m3). Dizziness and sleepiness
also were noted. Recovery from all symptoms was complete within 1 hour after
exposure. An exposure to an average concentration of 280 ppm PCE (1899 mg/m3)
for up to 2 hours resulted in complaints of 1ightheadedness, burning sensation
in the eyes, congestion of frontal sinuses, and tightness about the mouth.
Transient nausea was reported by one individual. The subjects felt that motor
coordination was impaired and mental effort was required for coordination.
Motor coordination was accomplished only with mental effort when two indivi-
duals were exposed to an average PCE concentration of 600 ppm (4070 mg/m3).
Recovery was complete within 1 hour after exposure. An average exposure
concentration of 1060 ppm (7190 mg/m3) for 1 minute was intolerable to three
of four individuals. None experienced functional disturbances. Recovery was
rapid.
No behavioral or neurological effects were reported by Carpenter (1937)
when four individuals were exposed to 500 ppm (3391 mg/m3) PCE for 70 minutes.
003PE5/A 6-7 11/22/83
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Short-term exposures to higher concentrations resulted in subject's reports of
mental fogginess, lassitude, inebriation, loss of inhibition, and vertigo. At
an exposure level of 1500 ppm (10,174 mg/m3), shortness of breath, nausea,
mental sluggishness, and difficulties in maintaining balance were reported
during the post-exposure period. Tinnitus, ringing of the ears, was reported
upon exposure to 2000 ppm PCE (13,565 mg/m3) for 7.5 minutes.
Weichardt and Lindner (1975) recorded subjective responses of headaches,
giddiness, numbness, alcohol intolerance, and intolerance of fats and fried
foods as a result of exposures to between 11 to 45 ppm PCE (75 to 305 mg/m3)
for approximately 3 hours. Very little information about the work environment
was available. In addition, the study failed to include suitable controls.
In a case study involving a nursing mother exposed daily for 30 to 60
minutes to unknown but excessive PCE concentrations, Bagnell and Ellenberger
(1977) reported subjective complaints of dizziness.
6.1.4.2 Effects of Long-term Exposures—In a comprehensive 3-month chamber
study of six males and six females, designed to elicit interactions between
ethanol and diazepam (Valium ) and PCE, Stewart and coworkers (Stewart et al.,
1977; Hake and Stewart, 1977) found that exposure to 25 and 100 ppm PCE (170
and 678 mg/m3) alone or in combination with diazepam or alcohol had no effect
on the EEC tracings. A battery of neurological and behavioral tests, consisting
of the following, were administered at the peak blood levels of exposure:
Michigan eye-hand coordination, rotary pursuit, Flanagan coordination, and
Saccade eye, velocity, and dual attention tasks. EEG's were recorded for
spectral density analysis during the exposures. The only decrement in the
behavioral tests reported by the authors was a "non-consistent significant
detrimental effect of PCE alone (100 ppm; 678 mg/m3) in the performance of the
Flanagan coordination test." This test requires subjects to follow a spiral
pathway with a pencil, touching the sides of the pathway as few times as
possible.
Each subject was exposed 5.5 hr/day. The exposure duration was 11 weeks.
Monday and Tuesday were generally control days. Thursday was an intermediate
exposure day. Wednesday and Friday were 100-ppm-PCE exposure days. No other
unusual behavioral or neurological findings were noted for exposures to PCE
alone. However, subjective complaints were noted for the nine subjects who
completed the study. One subject accounted for one-third the incidence of
headache complaints and two-thirds the incidence of nausea complaints reported.
003PE5/A 6-8 11/22/83
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The same subject complained of eye, nose, and throat irritation on all but 8
of the 55 test days. Analysis of the times of occurrence of these subjective
symptoms revealed that there was no relationship to PCE vapor exposure. In
fact, the incidence decreased somewhat when the 100-ppm-PCE (678 mg/m3) expo-
sure concentrations were compared to non-PCE exposures. The absence of EEC
abnormalities, such as were found in an earlier study by Stewart and co-workers
(1974), suggest that EEC observations may not be reliable indicators of early
signs of PCE narcosis. In their earlier study, impairment of the Flanagan co-
ordination test was also occasionally noted during exposures to 150 ppm for
7.5 hours. Chmielewski et al. (1976) observed abnormal EEC tracings in 10 of
16 individuals exposed occupationally, from 2 months to 27 years, to 59 to 442
ppm PCE (400 to 3000 mg/m3). However, 6 of 16 cases were further diagnosed as
pseudo-neurotic syndrome, and in 4 of 16 cases the researchers "were led to a
diagnosis or suspicion at least of encephalopathy." No additional details
were given.
Stewart et al. (1970) conducted an experiment with 16 healthy technical
employees ranging in age from 24 to 64 years. Five of the male subjects,
ranging in age from 36 to 64 years, were repeatedly exposed to 100 ppm (678
mg/m3) PCE for 7 hours on 5 consecutive days. The remainder of the subjects
were exposed for one 7-hour period. When queried every hour during the exposure
about subjective response, 25 percent of the subjects reported that they had
developed a mild frontal headache. Sixty percent complained of mild eye,
nose, or throat irritation developing within the first 2 hours of exposure and
usually subsiding before 7 hours had elapsed. Forty percent of the subjects
reported that they felt "slightly sleepy" when inactive in the chamber.
The authors wrote, "Unfortunately, it was not possible to confine these
same subjects in the exposure chamber for a comparable control period, so the
clinical significance of the slightly sleepy sensation cannot be assessed."
Twenty-five percent of the subjects "noted some difficulty in speaking,
analogous to that noted during the early phases of intoxication." The single
untoward objective response was an abnormal Romberg test occurring in three
of the subjects within the first 3 hours of exposure to 100 ppm (678 mg/m3).
With greater mental effort, these three individuals were able to perform a
normal test when given a second chance. Those subjects repeatedly exposed had
fewer subjective complaints. One subject, who had low-grade chronic sinusitis,
developed a mild frontal headache during the course of each exposure. Two out
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of five subjects consistently reported mild eye and throat irritation and light-
headedness. Sensations of sleepiness and speech difficulty were not reported.
All other neurological test results were normal. The group exposed repeatedly
responded more comparably to subjects in two subsequent, more sophisticated
3-week studies performed by Stewart et al. (1974 and 1977) with repeated
exposure to 100 ppm (678 mg/m?) PCE alone and in combination with either
diazepam or ethanol.
Similar subjective complaints, as well as neurological effects, of PCE
also were reported in other studies (Method, 1946; Lob, 1957; Eberhardt and
Freundt, 1966; Gold, 1969).
6.2 ANIMALS
Reported toxic effects associated with PCE exposure in laboratory animals
include effects on the CNS, cardiovascular system, skin, liver, kidney, and
the immune system.
A number of previous reviews (NIOSH, 1976; Fuller, 1976; U.S. Environmental
Protection Agency, 1980; Walter et al., 1976; Parker et al. , 1978; FishbeiX
1976; U.S. Environmental Protection Agency, 1977) support the assessment of
the toxic effects of PCE in animals as presented below. Summaries of these
toxic effects and of toxic dose data appear in Tables 6-1 and 6-2.
6.2.1 Effects on the Nervous System
Effects of acute exposure to PCE are very much dominated by CNS depression.
Abnormal weakness, handling intolerance, intoxication, restlessness, irregular
respiration, muscle incoordination, unconsciousness, and ultimately, death are
among the symptoms, considered to be manifestations of effects on the CNS,
which have been observed in animals exposed to excessive levels.
Symptoms of acute CNS depression have been seen in experimental animals
(National Institute of Occupational Safety and Health, 1976; Fuller, 1976;
U.S. Environmental Protection Agency, 1980; Walter et al. , 1976; Parker et
al., 1978; Fishbein, 1976; U.S. Environmental Protection Agency, 1977) and in
dogs treated with therapeutic (anthelmintic) doses of PCE (Miller, 1966;
Snow, 1973; Christensen and Lynch, 1973) PCE.
Rowe et al. (1952) reported that behavioral changes were not observed in
rats, guinea pigs, rabbits, or monkeys exposed repeatedly for 7 hr/day at
vapor concentrations of PCE up to 401 ppm (2720 mg/m3).
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TABLE 6-1. SUMMARY OF THE EFFECTS OF TETRACHLOROETHYLENE ON ANIMALS
Animal
species
Dose
concentration
Route of
administration
Exposure variables
Effects
Reference
Rabbit
(female)
Rabbit
Mouse
Guinea
pig
N/A
13 mmole/kg
200 ppm
Mouse
Guinea
pig
Guinea
pig
200 ppm
100 ppm
200 ppra
400 ppm
skin
oral
inhalation
inhalation
inhalation
inhalation
inhalation
single application
single instillation
(eye)
single dose
4 hours
single exposure
4 hours/day
6 days/week
1-8 weeks
7 hours/day
5 days/week
132 exposures
7 hours/day
5 days/week
7 hours/day
5 days/week
169 exposures
primary eye and
skin irritant
marked increase in
serum enzymes, i.e. ,
alkaline phosphatase,
SCOT, and SGPT within
24 hours
moderate fatty infiltration
of the liver 1 day
after exposure but not
3 days after
fatty degeneration of
the liver
increased liver weights
in females
increased liver weights
with some fatty degenera-
tion in both males and
females - slight increase
in lipid content and
several small fat vacuoles
in liver
more pronounced liver
changes than at 200 ppm,
slight cirrhosis was
observed - increased liver
weight, increase in neutral
fat and esterified choles-
terol in the liver, moderate
central fatty degeneration,
cirrhosis
Ouprat et al., 1976
Fujii et al., 1975
Kyi in et al., 1963
Kyi in et al. , 1965
Rowe et al. , 1952
Rowe et al., 1952
Rowe et al. , 1952
-------
TABLE 6-1. (continued)
Animal
species
Dose
concentration
Route of
adainistrati on
Guinea
pig
(Ti
I
Rabbit
Rat
Monkey
Rabbit
Rat
Rat
Rat
Rabbit
Rabbit
Exposure variables
Effects
Reference
2500 ppm
inhalation
100-400 ppn
2500 ppm
2500 ppra
1600 ppm
3000-6000
15 ppn
2212 ppm
Q5 ara/1)
inhalation
inhalation
inhalation
inhalation
Inhalation
18 7-hour
exposures
7 hours/day
5 days/week
6 months
28 7-hour
exposures
inhalation
Inhalation
1-13 7-hour
exposures
18 7- hour
exposures
single exposure
up to 8 hours
3-4 hours/day
7-11 months
45 days
4 hrs/day
5 days/week
loss of equilibrium,
coordination, and strength,
increase in weights of liver
and kidney, fatty degenera-
tion of the liver, cloudy
swelling of tubular epithe-
lium of the kidney
no abnormal growth,
organ function, or
histopathologic findings
central nervous system
depression without
unconsciousness
loss of consciousness
and death
drowsiness, stupor, increased
salivation, extreme restless-
ness, disturbance of equili-
brium and coordination, biting
and scratching reflex
increase in liver weight, in-
crease in total lipid content
of liver accompanied by a few
diffusely distributed fat
globules
depressed agglutlnin
formation
liver damage
indicated by elevated
SGPT, SCOT. SGLOH:
marked reduction of
Schmidt index
Rowe et al., 1952
Rowe et al., 1952
Rowe et al., 1952
Rowe et al., 1952
Rowe et al., 1952
Rowe et al., 1952
Hazza, 1972
Mazza, 1972
-------
TABLE 6-1. (continued)
Animal
species
Rat
Rat
Rat
Rat
Rat
Rabbit
Dose Route of
concentration administration Exposure variables
70 ppra inhalation 8 hours/day
5 days/week
150 exposures
(7 months)
230 ppm inhalation 8 hours/day
5 days/week
150 exposures
(7 months)
470 ppm inhalation 8 hours/day
5 days/week
150 exposures
(7 months)
2750-9000 inhalation single exposure
ppn
19000 ppm inhalation 30-60 minutes
15 ppm inhalation 3-4 hours/day
7-11 months
Effects Reference
no pathological findings Carpenter, 1937
similar, but less severe, Carpenter, 1937
pathological findings as
with 470 ppm - congestion
and light granular swelling
of kidneys
congested livers with cloudy Carpenter, 1937
swelling, no evidence of
fatty degeneration or
necrosis: evidence of kid-
ney injury - increased
secretion, cloudy swelling.
and desquamation of kidneys:
congestion of spleen
no deaths Carpenter, 1937
congested livers with granular Carpenter, 1937
swelling, some deaths
moderately increased Navrotskii et al., 1971
urinary urobi 1 inogen,
Rabbit
2211 ppm
(15 mg/1)
inhalation
45 days
pathomorphological
changes in the
parenchyma of liver
and kidneys
significant reduction
of glomerular filtration
rate and the renal
plasma flow; decrease
of highest excretory
tubular capacity
(kidney damage)
Brancaccio et al., 1971
-------
TABLE 6-1. (continued)
CTi
Animal Dose
species concentration
Mouse 2.5 ml/kg
(Swiss
male,
10 animals)
House 5.0 ml/kg
(10
Animals)
Rabbit 2211 pprn
(15 mg/1)
House N/A
House N/A
Dog
Dog
Dog
Rat 300 ppro
Route of
administration Exposure variables
intraperitoneal
intrapeHtoneal
(urine samples were collected 24 hours
inhalation 45 days
intraperitoneal
intraperitoneal
intraperitoneal
intraperitoneal
intraperitoneal
inhalation 7 hours/day
days 6-15 of
gestation
Effects
100 mg percent or more
protein found in one of
six mice - proximal con-
voluted tubules were
swollen in all animals
and necrotic in one
two of four mice had
100 mg percent or
more protein in urine
post-injection)
increased plasma and urine
levels of adrenal cortical
and adrenal medullar hor-
mones; increased excretion
of principal catecholamine
metabolite (not statistically
significant)
liver dysfunction
L050
elevated SGPT
caused phenol sulfo-
nephthalein retention
indicating kidney dysfunction
LD50
decreased maternal
weight gains,
increased fetal
Reference
Plaa and Larson, 1965
Mazza and Brancaccio,
1971
Klaassen and Plaa, 1966
Klaassen and Plaa, 1966
Klaassen and Plaa, 1967
Klaassen and Plaa, 1967
Klaassen and Plaa, 1967
Schwetz et al. ,
1975
reabsorptions
-------
TABLE 6-1. (continued)
Animal
species
Oose
concentration
Route of
administration
Exposure variables
Effects
Reference
House
300 ppm
inhalation
Rat
44.2 ppm
Mouse 15-74 ppm
Rat 15 ppm
fiat
Dog
(male
beagles)
73 and
147 ppm
0.5-1.0%
v/v
5000 &
10000 ppm
inhalation
inhalation
inhalation
inhalation
inhalation
7 hours/day
days 6-15 of
gestation
entire gestation
period
5 hours/day
3 months
4 hours/day
5 months
4 hours/day
4 weeks
7 minutes house air
followed by 10
minutes of tetrachloro-
ethylene 8 M9/k9
epinephrine given I.V.
(1) a control dose
after 2 minutes of
breathing air (2) chal-
lenge dose after 5 min-
utes of breathing test
compound
maternal liver weights
increased relative to
body weight; increased
incidences of fetal
subcutaneous edema,
delayed ossification of
skull bones, and split
sternebrae
decreased levels of ONA
and total nucleic acids
in the liver, brain,
ovaries, and placenta
decreased electroconductance
of muscle and "amplitude"
of muscular contraction
EEC changes and proto-
plasmal swelling of
cerebral cortical cells,
some vacuolated cells and
signs of karyolysis
EEC and electromyogram
changes; decreased
acetylcholinesterase
activity
cardiac sensitization
(development of serious
arrhythmia or cardiac
arrest) was not induced
at the concentrations
tested (other similar
compounds gave positive
results at same concen-
tration,
Schwetz et al.,
1975
Aninina, 1972
Dmitrieva, 1968
Omitrieva, 1966
Dmitrieva, 1966
Reinhardt et al. , 1973
-------
TABLE 6-1. (continued)
CTi
M
CTi
Animal
species
Cat
Cat
House
Mouse
Rabbit
Cat
Dog
Dose
concentration
3000 ppm
14,600 ppm
40 rag/1
5,900 ppm
4-5 ml/kg
5 ml/kg
4 ing/ kg
9000 ppm
Route of
administration Exposure variables
inhalation 4 hours
inhalation 1-2 hours
inhalation
oral
oral (in oil)
oral (in oil)
inhalation
Effects
no anesthesia
anesthesia
minimal fatal concentration
death in 2-9 hours from
central nervous system
depression
death in 17-24 hours
death within hours
narcosis, marked
Reference
Lehmann, 1911
Lehmann and Schmidt-
Kehl, 1936
Lamson et al . ,
Lamson et al . ,
Lamson et al . ,
Lamson et al . ,
Lamson et al . ,
1929
1929
1929
1929
1929
salivation, "narrow
margin of safety"
Dog
4-25 ml/kg
oral (in oil)
death in 5-48 hours
Lamson et al. , 1929
-------
TABLE 6-2. TOXIC DOSE DATA
(Ti
Description
of exposure*
LOsb
LDsb
E050
LD&0
ED50
LOsb
E050
E050
'-Dso
ED50
LOso
LD50
LDSO
LCLo
L050
Species
mouse (male)
mouse
mouse
mouse
mouse
dog
dog
dog
mouse
mouse
mouse
mouse
mouse
mouse
rat
Route of
administration
oral
interperitoneal
interperitoneal
interperitoneal
interperitoneal
interperitoneal
interperitoneal
interperitoneal
subcutaneous
subcutaneous
oral
(undiluted)
oral
(in oil)
oral
inhalation
Dose
concentration
6100 mg/kg
2.9 ml/kg
28 mM/kg
4700 mg/kg
2.9 ml/kg
28-32 mM/kg
34 mM/kg
24 mM/kg
2.1 ml/kg
21 mM/kg
3400 rag/ kg
0.74 ml/kg
7.2 mM/kg
1.4 mf/kg
390 mM/kg
27 mM/kg
0.109 ml
0.134 ml
8850 mg/kg
23000 mg/m3
Toxic effect
endpoint
death
death
liver
dysfunction
death
liver toxicity
death
liver damage
kidney
dysfunction
death
liver toxicity
death
death
death
death
death
Time
36 hr
24 hr
24 hr
24 hr
24 hr
24 hr
10 days
unknown
unknown
unknown
2 hr
Reference
Wenzel and Gibson, 1951
Klaassen and Plaa, 1966
Gehring. 1968
Klaassen and Plaa, 1967
Klaassen and Plaa, 1967
Klaassen and Plaa, 1967
Plaa et al. , 1958
Plaa et al. , 1958
Dybing and Dybing, 1946
Dybing and Oybing, 1946
Handbook of Toxicology, 1959
Withey and Hall, 1975
-------
TABLE 6-2. (continued)
I
M
CO
Description
of exposure
LCLo
LCLo
LDLo
LDLo
LOLo
LDLo
Species
rat
rat
dog
dog
cat
rabbit
Route of
administration
inhalation
inhalation
oral
intravenous
oral
oral
Dose
concentration
4000 ppm
4000 ppm
4000 mg/kg
85 rag/ kg
4000 mg/kg
5000 mg/kg
Toxic effect
endpoint
death
death
death
death
death
death
Time
4 hr
4 hr
unknown
unknown
unknown
unknown
Reference
Handbook of Toxicology, 1959
Arch. Hyg. Bakteriol.
116:131, 1936
Carpenter et al. , 1949
Clayton, 1962
Lamson et al. , 1929
LCL - the lowest concentration of a substance, other than an LCSu, in air which has been reported to have caused death in
humans or animals.
LDL - the lowest dose of a substance, other than an LDSo, that is introduced by any route other than inhalation over any
given period of time and reported to have caused death in humans or animals introduced in one or more divided
portions.
-------
A single 4-hour exposure to 2270 ppm PCE (15400 mg/m3) caused rats to
suffer an 80-percent loss of both avoidance and escape responses (Goldberg et
al., 1964). Savolainen et al. (1977) demonstrated behavioral impairment in
rats exposed to 200 ppm (1357 mg/m3) PCE vapor for 6 hr/day for 4 days.
Marked increases in the frequency of ambulation in the open field were most
significant 1 hour after exposure when these responses were compared to control
responses. High tissue concentrations of PCE were detected in fat and brain
tissue after a relatively short exposure. A significant decrease in the ribo-
nucleic acid (RNA) content of the brain was measured as well as an increase in
nonspecific cholinesterase activity.
Dmitrieva and coworkers (Dmitrieva, 1966; Dmitrieva and Kuleshov, 1971;
Dmitrieva et al., 1968; Dmitrieva, 1973; Dmitrieva, 1968) have reported altera-
tions of EEC patterns in rats exposed to as low as 15 ppm PCE (102 mg/m3) 3 to
4 hr/day for 7 to 11 months. Such effects have not been observed by Western
investigators at much higher levels nor were details provided in the reports
of these Russian investigators to adequately assess the quality of the studies.
6.2.2 Effects on the Liver and Kidney
PCE is generally regarded as being both hepatotoxic and nephrotoxic when
exposure is excessive and prolonged.
Carpenter (1937) exposed three groups of 24 albino rats each to PCE vapor
concentrations averaging 70, 230, or 470 ppm (475, 1560, or 3188 mg/m3) for 8
hr/day, 5 days/week, for up to 7 months. The maximum exposure for any animal
during the 7-month period was 150 days (1200 hours). A group of 18 unexposed
animals served as controls.
The rats exposed to 470 ppm (3188 mg/m3) for 150 days followed by a
46-day rest period developed cloudy and congested livers with swelling; there
was no evidence of fatty degeneration or necrosis. These rats also had in-
creased renal secretion with cloudy swelling and desquamation of kidneys, as
well as congested spleens with increased pigment. The pathologic changes were
similar but less severe in the rats exposed to 230 ppm PCE (1560 mg/m3). In
some instances, there was congestion and light granular swelling of the kidneys
after 21 exposure days. After 150 days of exposure, followed by a 20-day
rest, congestion was found in the kidney and spleen. The livers showed reduced
glycogen storage. Microscopic evidence of damage to liver, kidney, or spleen
in rats exposed at 70 ppm (475 mg/m3) for 150 days was not observed. In addi-
tion, microscopic examination of heart, brain, eye, or nerve tissue did not
003PE5/A 6-19 11/22/83
-------
reveal any damaging effects in any of the chronically exposed rats. Functional
parameters, including icteric index, Van den Bergh test for bilirubin, and
blood and urine analysis, were normal after the exposures. -Fertility of
female rats, as measured by a fertility index (actual number of litters/possible
number of litters), was increased slightly after repeated exposures to 230 or
470 ppm PCE (1560 or 3188 mg/m3). No deaths or signs of anorexia or diminished
activity were observed during the chronic exposures.
Carpenter also tried to determine the highest concentration of PCE vapor
that would not anesthetize rats exposed for 8 hours. Exposure to 31,000 ppm
(210,273 mg/m3) was lethal within a few minutes. Rats exposed to 19,000 ppm
(128,871 mg/m3) died after 30 to 60 minutes. Animals that were exposed to
19,000 ppm (128,877 mg/m3) and removed from the inhalation chamber just prior
to unconsciousness developed congestion and granular swelling of the liver.
Similar liver effects were seen after exposure at 9000 ppm (61,047 mg/m3).
There was also marked granular swelling of the kidneys. A single exposure at
9000, 4500, or 2750 ppm (61,047, 30,523, or 15,261 mg/m3) did not cause death
to any of the rats in this study; however, post-mortem examinations of the
rats exposed to those concentrations revealed only a slight increase in the
prominence of liver and kidney markings.
Rowe et al. (1952) exposed rabbits, monkeys, rats, and guinea pigs to PCE
vapor for 7 hours, 5 days/week, for up to 6 months. Exposure concentrations
ranged from 100 to 2500 ppm (678 to 16,957 mg/m3). Three of the four species
tested -- rabbits, monkeys, and rats -- showed no effects of repeated exposures
to concentrations, up to 400 ppm (2713 mg/m3). There were no adverse effects
on growth, liver weight, or lipid content, or gross or microscopic anatomy
observed in any animal. In contrast, guinea pigs showed marked susceptibility
to PCE in this study. The liver weights of female guinea pigs increased
significantly after 132 7-hour exposures at 100 ppm (678 mg/m3). At 200 ppm
(1356 mg/m3), there was a slight depression of growth in female guinea pigs
and increased liver weights in both males and females. Slight to moderate
fatty degeneration of the liver also was observed. These effects were more
pronounced in guinea pigs that received 169 7-hour exposures at 400 ppm (2713
mg/m3). At this concentration, there also were increased amounts of neutral
fat and esterified cholesterol in livers. Gross and microscopic examination
of the tissues revealed slight to moderate fatty degeneration in the liver
with slight cirrhosis. Rowe et al. stated that at 395 ppm (2680 mg/m3),
003PE5/A 6-20 11/22/83
-------
increased kidney weights also were observed in guinea pigs but not in other
species. Guinea pigs have been shown to be extremely sensitive to toxicity
testing.
Klaassen and Plaa (1967) showed that intraperitoneal injections of PCE to
mongrel male and female dogs can produce damage to the kidney and liver. They
estimated the ED50 (effective dose in 50 percent of the animals tested) for
liver and kidney damage as well as the 24-hour LD50 value (lethal dose in 50
percent of the animals treated). The ED50 values were measured by sulfobromo-
phthalein (BSP), SGPT, glucose, protein, and phenolsulfonephthalein (PSP),
indicators of liver or kidney dysfunction. The LD50 was 2.1 ml/kg (21 mmoles/
kg) and the ED50 for elevation of SGPT was 0.74 ml/kg (7.2 mmoles/kg). The
ED50 for diminution of PSP excretion was 1.4 ml/kg (14 mmoles/kg). After
administration, effects on the liver and kidneys were determined by microscopic
examination. At near-lethal doses, PCE produced moderate neutrophilic infil-
trations in the sinusoids and portal areas; necrosis was not observed. Near
the EDr0, vacuolization of centralobular hepatocytes in about half the animals
was observed. After a single interperitoneal injection at 0.75 x 1-0™, SGPT
peaked at 2 days and rapidly declined throughout the 9-day measurement period.
Kidney dysfunction was deemed significant when PSP excretion was less than 39
percent (determined 24 hours after interperitoneal injection). At near PSP
excretion--EDcn doses, only mild dilation of the collecting ducts was seen in
some of the kidneys.
In a similar study using Swiss-Webster mice, Klaassen and Plaa (1966)
determined that the 24-hour LD™ was 2.9 ml/kg (28 mmoles/kg). The ED™ for
BSP retention was 2.9 ml/kg (32 mmoles/kg). The ED5Q for elevation of SGPT
was 2.9 ml/kg (28 mmoles/kg). When ethanol (5 gm/kg) was administered by
gavage for 3 days prior to injection of 1.0 ml/kg PCE, neither PSP excretion
nor BSP retention was significantly altered. At BSP-ED™, mice exhibited
predominantly inflammatory changes with trace to marked quantities of lipid
accumulation.
Effects upon the livers of rats exposed continuously for 3 months to 0.7
3
and 2.7 ppm (4.5 and 19 mg/m ) were reported by Bonashevskaya (1977). In the
low-exposure group, an increase in the activity of succinate dehydrogenase
(SDH) was reported. At the high level, SDH was slightly decreased in the
central lobular sections while activity in the peripheral zones was either
unchanged or increased. The activities of glucose-6-phosphate dehydrogenase
(G6PDH) were similar in pattern to that for SDH. RNA content was reported to
003PE5/A 6-21 11/22/83
-------
be decreased. LDH activity was either unchanged or increased. The enzyme
variations indicated the subject's adaptation to exposure. No significant
histomorphological changes in the liver were reported at either exposure
level.
A subcutaneous injection (0.4 ml daily, for 8 weeks) of PCE to groups of
rats that were fed a diet containing high protein had less of an adverse
effect upon the liver than PCE alone (Dumitrache et al., 1975). PCE was
observed to induce liver hypertrophy compared to the reaction of the controls
(p <0.001). Hypertrophy was most pronounced in the low-protein group. In
treating rats fed a diet containing low protein, cholesterol (p <0.001) and
total liver lipids (p < 0.001) were elevated compared to the reaction of the
controls. Twelve rats were used in each of the four groups.
Kyi in et al. (1963) noted moderate fatty degeneration of the liver with a
single 4-hour exposure to 200 ppm PCE (1356 mg/m3) in female albino mice that
were sacrificed 1 day after exposure. Degeneration was not observed in mice
sacrificed 3 days after exposure. The mice were exposed to PCE concentrations
of 200, 400, 800, or 1600 ppm (1356, 2713, 5426, or 10,852 mg/m ) for 4 hours.
Tissues were studied microscopically to assess the extent of necrosis and fat
infiltration of the liver. Moderate to massive infiltration was observed in
3
mice killed 1 or 3 days after exposure at 400 ppm (2713 mg/m ) or more, but no
cell necrosis was observed even after 4 hours exposure up to 1600 ppm PCE
3
(10,852 mg/m ). Kylin et al. (1963) exposed four groups of 20 albino mice
3
to 200 ppm PCE (1356 mg/m ). Each group was exposed for 4 hr/day, 6 days per
week, for 1, 2, 4, or 8 weeks. Microscopic examinations were performed on
livers and kidneys of the exposed mice and controls. Fatty degeneration was
particularly marked and tended to be more severe with longer exposure to PCE.
Chemical determination of the liver fat content was performed in addition to
the histologic examination. Correlation between the histologically evaluated
degree of fatty degeneration and the concentration of extracted fat was +0.74.
Liver fat content of the exposed animals was between 4 and 5 mg/g of body
weight, as compared to 2 to 2.5 mg/g for the control animals. The actual fat
content of the livers did not increase with duration of exposure as did the
extent of the fatty infiltration. No liver cell necrosis was observed. No
effect on the kidneys was reported.
Mazza (1972) exposed 15 male rabbits for 4 hr/day, 5 days a week, for 45
3
days, to 2790 ppm PCE (18,924 mg/m ). Mazza looked at the effect of PCE on
003PE5/A 6-22 11/22/83
-------
serum enzyme levels in an attempt to determine the specific location of initial
liver injury as well as the severity of the damage to the liver. The Schmidt
Index, which is the sum of SCOT and the SGPT divided by the serum glutamate
dehydrogenase (GDH), was used as an indication of hepatic disorders. Enzymatic
determinations were made before exposure and 15, 30, and 45 days after exposure
to PCE. All three of the enzymes showed an increase in activity, but the GDH
increased the most; GDH reduced the Schmidt Index from 6.70 to 1.79. Mazza
concluded that this reduction indicates the prevalence of mitochondrial injury
over cytoplasmic injury in the liver.
Mazza and Brancaccio (1971) exposed 10 rabbits for 4 hr/day, 5 days per
3
week, for 45 days, to 2790 ppm PCE (18,924 mg/m ). These investigators found
a moderate, but not statistically significant, increase in levels of adrenal
cortical and medullar hormones—plasma and urinary corticosteroids and cate-
cholamines--including increased excretion of 3-methoxy-l-hydroxymandelic- acid,
the principal catecholamine metabolite.
In another study, Brancaccio et al. (1971) exposed 12 male rabbits for 4
3
hr/day, 6 days per week, for 45 days, to 2280 ppm PCE (15,465 mg/m ) to look
at effects on kidney function. They noted a reduction in glomerular filtra-
tion and renal plasma flow and a decrease in the maximum tubular excretion
when measured upon cessation of the exposure regimen. Brancaccio et al.
concluded that PCE causes kidney damage, primarily in the renal tubule. These
findings agreed with earlier histological findings of Pennarola and Brancaccio
(1968) in which kidney injury, following exposure to PCE, appeared to be
primarily in the renal tubule.
Plaa and Larson (1965) dosed mice with PCE by interperitoneal injection.
Ten mice received 2.5 mg/kg and 10 others received 5.0 mg/kg. Urine samples
were collected from surviving mice 24 hours after the injection of PCE.
Protein was found in the urine of one of the six surviving mice injected with
the lower dose and in two of four survivors of the higher dose at levels of
100 or more mg percent. None of the survivors had greater than 150 mg percent
glucose in the urine. The kidneys of the mice given the lower dose were
examined microscopically. The proximal convoluted tubules were swollen in all
animals and necrotic in one.
Fujii (1975) observed an increase in serum enzyme activities (i.e.,
alkaline phosphatase, SCOT, and SGPT) within 24 hours after a single dose of
13 mmole PCE/kg given orally to rabbits. These changes in serum enzyme activi-
ties, indicative of liver damage, were mild and transient.
003PE5/A 6-23 11/22/83
-------
6.2.3 Effects on the Heart
The possible cardiovascular effects of PCE have not been systematically
investigated. Reinhardt et al. (1973) noted that PCE does not appear to
sensitize the myocardium to epinephrine. In this study of dogs, a response
that was considered to be indicative of cardiac sensitization was the develop-
ment of a seriously life-threatening arrhythmia or cardiac arrest following a
challenge dose of epinephrine. PCE inhalation exposure for 10 minutes at
3
concentrations of 5000 or 10,000 ppm (33,915 or 67,830 mg/m ) did not result
in a positive response in any of the 17 dogs.
The investigators noted the possibility that PCE has the potential for
cardiac sensitization, as do many organic solvents. However, the PCE concentra-
tion that failed to elicit a positive response (10,000 ppm) would be a level
immediately hazardous to life, based on CNS effects.
Christensen and Lynch (1933) observed depression of the heart and respira-
tion in five dogs, each given a single oral dose ranging from 4 to 5.3 ml/kg
PCE (approximately the LD5Q for dogs). Autopsy showed fatty infiltration of
both heart and liver tissue. The small intestine was extremely shriveled and
showed marked inflammation.
6.2.4 Effects on the Skin and Eye
Duprat et al. (1976) have shown PCE to be a primary eye and skin irritant
in rabbits. Instillation of the chemical into the eye produced conjunctivitis
with epithelial abrasion. However, healing of the ocular mucosa was complete
within 2 weeks. PCE had a severe irritant effect when a single application
was made to the skin of the rabbit.
6.3 ADVERSE EFFECTS OF SECONDARY POLLUTANTS
The level of phosgene derived from photodecomposition of PCE is not
likely to result in serious long-term effects. In occupational settings or
under certain conditions in which phosgene is directly formed at high tem-
peratures from halocarbons (cigarette smoke, welding), sufficient warning
from the generation of extremely irritating hydrogen chloride vapors would
prevent exposure to harmful concentrations of phosgene.
6.4 SUMMARY OF ADVERSE HEALTH EFFECTS AND ASSOCIATED LOWEST OBSERVABLE
EFFECT CONCENTRATIONS
6.4.1 Inhalation Exposure
A number of case reports describe accidental or occupational exposure to
PCE. However, the duration and extent of exposure either were unknown or
involved excessively high concentrations. The few controlled human studies
003PE5/A 6-24 11/22/83
-------
available generally provide information on effects resulting from short-term
3
exposures at levels near 100 ppm (678 mg/m ). Effects associated with chronic
exposures, on the other hand, are available from animal experiments.
6.4.1.1 Effects Associated With Intermittent or Prolonged Exposure—The data
available from both human and animal exposures to excessive concentrations of
PCE indicate that the CNS, liver, and kidneys may be adversely affected. In a
controlled study involving humans, Stewart and coworkers (1977) exposed 12
3
individuals to 25 and 100 ppm PCE (170 and 678 mg/m ), 5.5 hr/day, for 3
3
months (55 exposures). At 100 ppm PCE (678 mg/m ) there was a non-consistent
detrimental effect upon coordination, as measured by the Flanagan test, but no
adverse effects on the remainder of a battery of neurological and behavioral
tests and no alterations in the EEC patterns. There were no observed effects
at 25 ppm (170 mg/m3).
Both subchronic and chronic inhalation exposures of animals support the
observations in humans. Carpenter (1937) observed dose-related congestion and
swelling in the livers and kidneys of rats exposed for 8 hr/day, 5 days per
3
week, for 7 months, to 230 and 470 ppm (1559 and 3187 mg/m ). Rowe et al.
(1952) observed degenerative changes in the livers of guinea pigs, a species
that appears more sensitive to PCE, following an 8-month exposure to 200 and
3
400 ppm (1356 and 2712 mg/m ), for 7 hr/day, 5 days per week.
Carpenter (1937) reported a no-observable-effects level (NOEL) at 70 ppm
3
PCE (475 mg/m ) in. rats exposed 8 hr/day, 5 days per week, for 7 months. His-
tologic examination of the liver, kidneys, spleen, heart, brain, eye, and
nerve tissue, as well as blood and urine analyses, were performed. The NOEL
is defined as that exposure level at which there are no statistically signi-
ficant increases in frequency or severity of effects between the exposed
population and the appropriate control.
6.4.1.2 Effects Associated With Short-term Exposure—Controlled studies in
humans have provided information on a progression of effects upon the CNS,
ranging from lightheadedness to narcosis, and subsequently to death. A summary
of the estimated dose-response relationships for acute effects of single,
short-term exposures of humans is presented below:
>4000 ppm PCE: possibly life-threatening
2000 ppm PCE: light narcosis, possible liver
damage
1000 ppm PCE: lightheadedness
003PE5/A 6-25 11/22/83
-------
200 ppm PCE: possible reversible fatty changes
in the liver
100 ppm PCE: possible first signs of CNS depression,
giddiness, slight eye irritation
75 ppm PCE: very slight eye irritation
5 ppm PCE: odor threshold
Stewart et al. (1970) reported 1ightheadedness in 25 percent of the 15
volunteers exposed for 7 hours to an average PCE concentration of 101 ppm
3
(685 mg/m ). Two subsequent studies by Stewart et al. (1974, 1977), using
more updated technology and testing methodology, did not confirm this result
even when exposures to 100 ppm PCE were combined with diazepam or alcohol.
Rowe et al. (1952) reported 1ightheadedness in four individuals exposed for 2
3
hours to 216 ppm PCE (1465 mg/m ), but not in six individuals exposed 2 hours
3
to 106 ppm (719 mg/m ). The observations in these studies follow the expected
pattern for a non-specific anesthetic effect. These results show that a
short-term exposure to PCE in the range of about 100 to 200 ppm (678 to 1356
3
mg/m ) will result in some individuals experiencing the first gross signs of
CNS depression and behavioral alterations. Test results also indicate that
short-term exposures may damage the liver. Kyi in et al. (1963) reported
transient moderate fatty infiltration in livers of mice exposed for 4 hours to
3
200 ppm PCE (1356 mg/m ). However, liver damage in humans is generally asso-
3
ciated with short-term exposures greatly in excess of 100 ppm (678 mg/m ).
6.4.2 Oral Exposure
As summarized in Table 6-2, the acute oral toxicity of PCE has been
determined in rats, mice, cats, rabbits, and dogs. There are no subchronic
oral exposure studies and only one chronic study—the NCI bioassay. Neither
of these studies identifies either a NOEL or lowest-observed-adverse-effect
level (LOAEL).
6.4.3 Dermal Exposure
Although PCE can be absorbed through unbroken skin, absorption was esti-
mated to be minor (Stewart and Dodd, 1964). Toxic quantities would probably
not be absorbed through this route.
003PE5/A 6-26 11/22/83
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6.5 REFERENCES FOR CHAPTER 6
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Technical Report - 7143. February 1976.
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to their narcotic and lethal properties in mice. Toxicol. Appl. Pharmacol.
13:287-298, 1968.
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Goldberg, M. E.; H. E. Johnson; U. C. Pozzani; H. F. Smyth, Jr. Effect of
repeated inhalation of vapors of industrial solvents on animal behavior.
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Am. Ind. Hyg. Assoc. J. 25:369-375, 1964.
Hake, C. L.; R. D. Stewart. Human exposure to tetrachloroethylene: inhala-
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Handbook of Toxicology, Volumes II-V, Philadelphia: W. B. Saunders Co., 1959.
Volume V., p. 76.
Hughes, J. P. Hazardous exposure to a so-called safe solvent. J. Am. Med.
Assoc. 156:234-237, 1954.
Klaassen, C. D.; G. L. Plaa. Relative effects of various chlorinated
hydrocarbons on liver and kidney function in mice. Toxicol. Appl.
Pharmacol. 9:139-151, 1966.
Klaassen, C. D.; G. L. Plaa. Relative effects of various chlorinated
hydrocarbons on liver and kidney function in dogs. Toxicol. Appl. Pharmacol
10:119-131, 1967.
Kylin, B.; I. Sumegi; S. Yllner. Hepatotoxicity of inhaled trichloroethylene
and tetrachloroethylene - long-term exposure. Acta Pharmacol. Toxicol.
(Kbh). 22:379-385, 1965.
Kylin, B. ; H. Reichard; I. Sumegi; S. Yllner. Hepatotoxicity of inhaled
trichloroethylene, tetrachloroethylene, and chloroform—single exposure.
Acta Pharmacol. Toxicol. 20:16-26, 1963.
Lamson, P. D.; B. H. Robbins; C. B. Ward. The pharmacology and toxicology
of tetrachloroethylene. Am. J. Hyg. 9:430-444, 1929.
Larson, N. A.; B. Nielsen; A. Ravin-Nielsen. Perchloroethylene intoxica-
tion—a hazard in the use of coin laundries. Ugeskr. Laeg. 39(5):270-275,
1977. (English translation)
Lehmann, K. B. Experimental studies on the influence of technically and
hygienically important gases and vapors on the organism. Arch. Hyg.
74:1-60, 1911. (German)
Lehmann, K. B.; L. Schmidt-Kehl. The 13 most important chlorinated hydro-
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Levine, B.; M. F. Fierro; S. W. Goza; J. C. Valentour. A tetrachloro-
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1971.
Lob, M. Dangers of perchloroethylene. Arch. Gewerbepath. Gewerbehyg. 16:45-52,
1957. (English translation)
003PE5/A 6-29 11/22/83
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Mazza, V. Enzyme changes in experimental tetrachloroethylene intoxication.
Folia Med. 55(9-10):373-381, 1972. (English translation)
Mazza, V.; A. Brancaccio. Adrenal cortical and medullar hormones in
experimental tetrachloroethylene poisoning. Folia Med. 54:204-207,
1971. (English translation)
Meckler, L. C.; D. K. Phelps. Liver disease secondary to tetrachloroethylene
exposure. J. Am. Med. Assoc. 197(8):144-145, 1966.
Method, H. C. Toxicity of tetrachloroethylene. J. Am. Med. Assoc. 131:1468,
1946.
Miller, T. A. Anthelmintic activity of tetrachloroethylene against various
stages of Ancylostoma cam'urn in young dogs. Am. J. Vet. Res. 27(119):
1037-1040, 1966.
Moeschlin, S. Poisons, diagnosis and treatment. New York, N.Y.: Gruner and
Stratton, 1965.
Morgan, B. Dangers of perchloroethylene. Brit. Med. J. 2:513, 1969.
National Institute for Occupational Safety and Health. Criteria for a recom-
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Saland, G. Accidental exposure to perchloroethylene. N. Y. State J. Med.
67:2359-2361, 1967.
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7. TERATOGENICITY, EMBRYOTOXICITY, AND REPRODUCTIVE EFFECTS
Because of its widespread use, PCE has been studied for teratogenic
potential. Teratology studies have been performed in rats, mice, and rabbits,
using doses of PCE which, in some studies, produced slight signs of maternal
toxicity. Other studies in chicken embryos (Elovaara et al., 1979) have
indicated that PCE disrupts embryogenesis in a dose-related manner. However,
since administration of PCE directly into the air space of chicken embryo is
not comparable to administration of dose to animals with a placenta, it is not
possible to interpret this result in relationship to the potential of PCE to
cause adverse effects in animals or humans.
The following discussion of studies subscribes to the basic viewpoints
and definitions of the terms "teratogenic" and "fetotoxic" as summarized and
stated by the U.S. Environmental Protection Agency (1980):
Generally, the term "teratogenic" is defined as the tendency to
produce physical and/or functional defects in offspring i_n utero. The
term "fetotoxic" has traditionally been used to describe a wide variety
of embryonic and/or fetal divergences from the normal which cannot be
classified as gross terata (birth defects) -- or which are of unknown or
doubtful significance. Types of effects which fall under the very broad
category of fetotoxic effects are death, reductions in fetal weight,
enlarged renal pelvis edema, and increased incidence of supernumerary
ribs. It should be emphasized, however, that the phenomena of terata and
fetal toxicity as currently defined are not separable into precise cate-
gories. Rather, the spectrum of adverse embryonic/fetal effects is
continuous, and all deviations from the normal must be considered as
examples of developmental toxicity. Gross morphological terata represent
but one aspect of this spectrum, and while the significance of such
structural changes is more readily evaluated, such effects are not neces-
sarily more serious than certain effects which are ordinarily classified
as fetotoxic--fetal death being the most obvious example.
In view of the spectrum of effects at issue, the Agency suggests
that it might be useful to consider developmental toxicity in terms of
three basic subcategories. The first subcategory would be embryo or
fetal lethality. This is, of course, an irreversible effect and may
occur with or without the occurrence of gross terata. The second subcate-
gory would be teratogenesis and would encompass those changes (structural
and/or functional) which are induced prenatally, and which are irreversible.
Teratogenesis includes structural defects apparent in the fetus, functional
deficits which may become apparent only after birth, and any other long-
term effects (such as carcinogenicity) which are attributable to i_n utero
exposure. The third category would be embryo or fetal toxicity as com-
prised of those effects which are potentially reversible. This subcategory
would therefore include such effects as weight reductions, reduction in
the degree of skeletal ossification, and delays in organ maturation.
003PE2/E 7-1 11/22/83
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Two major problems with a definitional scheme of this nature must be
pointed out, however. The first is that the reversibility of any phenom-
enon is extremely difficult to prove. An organ such as the kidney, for
example, may be delayed in development and then appear to "catch up."
Unless a series of specific kidney function tests are performed on the
neonate, however, no conclusion may be drawn concerning permanent organ
function changes. This same uncertainty as to possible long-lasting
aftereffects from developmental deviations is true for all examples of
fetotoxicity. The second problem is that the reversible nature of an
embryonic/ fetal effect in one species might, under a given agent, react
in another species in a more serious and irreversible manner.
7.1 ANIMAL STUDIES
7.1.1 Mice
Schwetz et al. (1975) reported a study of Swiss-Webster mice exposed to
3
PCE via inhalation at concentrations of 300 ppm (2034 mg/m ) for 7 hours daily
on days 6 through 15 of presumed gestation. Day 0 of gestation was designated
the day a vaginal plug was observed. This concentration was cited as twice
the maximum allowable exclusion limit for human industrial exposure, with the
® •*
Threshold Limit Value (TLV) of 100 ppm (678 mg/m ). Concurrent controls were
exposed to filtered air.
Following maternal Caesarean-sectioning on day 18 of gestation, all
fetuses were examined for external anomalies. One-half the fetuses were
examined for soft tissue malformations using a free-hand sectioning technique,
and the other one-half of the fetuses were cleared, stained, and examined for
skeletal malformations. One fetus in each litter was processed and examined,
using histopathological techniques. Seven litters were examined. The pups in
the exposed group were significantly smaller, as measured by decreases in
fetal body weight. Also, slight but not statistically significant increases
in the number of runts were observed, as well as increases in the numbers of
fetuses with subcutaneous edema, delayed ossification of the skull, delayed
ossification of the sternebrae, and splits in sternebrae. No other remarkable
malformations were reported in fetuses. Increases in the absolute and relative
mean maternal liver weights were reported. No evidence of teratogenicity of PCE
was found at the concentration tested.
7.1.2 Rats
Schwetz et al. (1975) also administered PCE to Sprague-Dawley rats by
3
inhalation (300 ppm; 2034 mg/m ) for 7 hours daily, on days 6 to 15 of gesta-
tion. Control rats were exposed to filtered air. Day 0 of gestation was
003PE2/E 7-2 11/22/83
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designated as the day when spermatozoa were observed in smears of vaginal
contents. Rats were sacrificed on day 21 of gestation. Caesarean sections
were performed, and fetuses were examined for external malformations. One-
half of the fetuses in each litter was examined for soft tissue malformations
and the remaining one-half of the fetuses were examined for skeletal malforma-
tions. One fetus from each litter was randomly selected for serial sectioning
and histological evaluation. Average maternal body weight gain was slightly
reduced in the rats exposed to PCE. A slight but statistically significant
increase in resorption was reported in 9 of 17 PCE-exposed litters evaluated.
Exposure of dams to PCE produced no effect on the average number of implanta-
tions per litter, fetal sex ratios, or fetal body measurements. No soft
tissue or skeletal anomalies were reported in the offspring of rats exposed to
PCE. No evidence of teratogenicity of PCE was found at the concentration
tested.
Beliles et al. (1980) exposed Sprague-Dawley rats to PCE at 300 ppm
3 •
(2034 mg/m ) for 7 hours daily, 5 days per week. Controls were administered
filtered air. One-half of the rats were exposed for 3 weeks prior to mating.
All rats were exposed during gestation on either days 0 through 18 or days 6
through 18, with day 0 designated as the day when spermatozoa were observed in
smears of vaginal contents. It should be noted that the information in the
report indicated that a Monday through Friday treatment period was used. It
is unclear, based on information in the report, if the rats were indeed exposed
on the gestational days cited. Three rats (days 6-18) died on the second day
of pregestational treatment. Signs of ataxia and loss of balance were observed
in all of the other rats of this group on the same day. The authors thought
this response was most likely due to the high levels in the inhalation chamber
during the last 2 hours of the day. Measurements taken 15 minutes before the
3
end of the exposure showed 568 ppm (4061 mg/m ) but could have been higher
previously. The maternal body weight gain in the PCE-treated rats was not
statistically different from that in controls during the pregestational period.
Inhibition of maternal body weight gain occurred during the first week, as
well as increases in mean absolute, but not relative, kidney and liver weights.
No embryotoxic effects were observed which were attributable to maternal
exposure to PCE, except for delays in skeletal ossification. This effect,
however, is thought to be a reversible effect and is not considered a malformation
as such. No teratogenic effects were observed.
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Nelson et al. (1980) evaluated the ability of PCE to elicit a behavioral
teratogenic response in Sprague-Dawley rats. In a pilot dose-range finding
study, groups of 3 rats were exposed by inhalation to 1800 or 3600 ppm (12,204
3
or 24,408 mg/m ) PCE. Narcotization was observed in dams. Therefore, a
3
concentration of 900 ppm (6102 mg/m ) PCE was used in the pilot study. Signs
of maternal toxicity such as severe reductions in average food intake and
decreased average body weight gain were observed. Food consumption and body
weight gain were also reduced but were not statistically significant in rats
3
exposed to 900 ppm (6102 mg/m ) PCE on days 14 to 20 of gestation. Dams and
pups exposed to 100 ppm (678 mg/3) PCE on days 14 to 20 of gestation did not
show any adverse effects as compared to the controls.
In the behavioral testing study, 102 pregnant rats were exposed to PCE by
inhalation as follows:
(1) 900 ppm, days 7-13 of gestation (N=19)
(2) 900 ppm, days 14-20 of gestation (N=21)
(3) sham exposed, days 7-13 of gestation (N=13)
(4) sham exposed, days 14-20 of gestation (N=19)
(5) 100 ppm, days 14-20 of gestation (N=15)
(6) sham exposed, days 14-20 of gestation (N=15)
Seven behavioral tests were performed on days 4 through 46 postparturition,
using one male and one female per litter. One group of rats, consisting of a
male/female pair from each litter, was tested for ascent on a wire mesh
screen and rotorod balancing. One male and one female rat per litter were
tested for open-field activity, activity wheel, and avoidance conditioning. A
third pair was tested for operant conditioning. Catecholamines (norepinephrine
and dopamine), acetylcholine, and protein, measured in the brain tissue, were
evaluated in 10 pups (no more than 2 per litter per treatment group) at birth,
or at 21 days postparturition. Histopathological evaluation of brains for
neuropathology was performed in an unreported number of pups.
3
Rats from dams exposed to 900 ppm (6102 mg/m ) PCE on days 7 to 13 of
gestation, performed less well on discrete testing of ascent and rotorod
tests, but only on certain days of testing. Offspring exposed to 900 ppm
3
(6102 mg/m ) on days 14 to 20 of gestation performed less well on one test day
in the ascent test, but later performed better than controls in the rotorod
test, and were relatively more active than controls in the open-field tests.
Acetylcholine levels were reduced in 21-day-old rats of dams exposed to 900 ppm
003PE2/E 7-4 11/22/83
-------
3
(6102 mg/m ) PCE. Dopamine levels were reduced in rats of dams exposed on
days 7 to 13 of gestation.
3
In the group of rats exposed to 100 ppm (678 mg/m ) PCE during days 14-20
of gestation, no significant differences were observed between the offspring
of these animals and their controls on any of the behavioral tests. The
authors summarized that "there were generally few behavioral or neurochemical
o
differences observed between offspring of animals exposed to 900 ppm (6102 mg/m )
PCE during either days 7-13 or 14-20 of gestation. When significant differ-
ences did appear, they occurred more often when the group exposed during days
14-20 of gestation was compared with its control group."
It should be noted that the field of behavioral teratology is in its
early stages of development (Buelke-Sam and Kimmel, 1979), and such behavioral
alterations cannot at this time be interpreted in terms of human effect. It
also must be noted that these observed changes may be related to maternal
toxicity and do not represent a direct toxic effect.
Tepe et al. (unpublished, 1982) exposed Long-Evans hooded female rats to
3
1000 ± 125 ppm (6780 ± 847 mg/m ) PCE vapors to ascertain if exposure before
mating and during pregnancy was more detrimental to the embryo than exposure
during pregnancy alone. Four treatment groups (30 rats each) were utilized in
a two-by-two factorial design: exposure to PCE for two weeks (6 hours daily,
5 days per week) before mating, through day 20 of pregnancy; PCE before mating
and filtered air during pregnancy; filtered air before and PCE during pregnancy;
and filtered air throughout. One-half the dams in each treatment group were
sacrificed on day 21 by ether anesthesia. Elevations in relative maternal
liver weights and reduction in fetal body weight were observed without altera-
tion in maternal weight-gains in groups exposed during pregnancy. An excess
of skeletal variations were seen in the group exposed before mating and during
pregnancy; and excessive soft tissue variations (e.g. kidney dysplasia) occurred
in the group exposed during pregnancy alone. These effects are consistent
with embryotoxicity. Elevation in ethoxycoumaren dealkylase activity (an
indicator of P45Q activity) was observed in maternal livers but not fetal
livers with pregnancy exposure. Ethoxyresorufin dealkylase activity (an
indicator of ?..„ activity) was not elevated in maternal livers with PCE
exposure and was not detectable in fetal livers. No other effects were observed
in pregnant rats exposed for 6 h per day on days 0-20 of gestation.
Hanson et al. (unpublished, 1982) conducted a study on postnatal evaluation
3
of offspring of the female rats exposed to 1000 ± 125 ppm (6780 ± 847 mg/m ) PCE
before and/or during pregnancy. The four treatment groups are those described
003PE2/E 7-5 11/22/83
-------
above (Tepe et al., 1982). The purpose of the postnatal evaluation was to deter-
mine if the reduction in body weight or the excess skeletal and soft tissue
variants observed in term fetuses from the Tepe et al. (1982) study persisted,
and if PCE possessed transplacental carcinogenic activity or neurobehavioral
toxicity. Weight gain and survival of offspring up to 18 months of age, and
frequency of any gross lesions observed at 6- and 18-month autopsies, were not
influenced by prenatal exposure to PCE. Neurobehavioral tests of general acti-
vity in open field tests at 10 and 20 days of age, and in running wheels from 40
to 100 days of age, did not indicate treatment-related effects. Likewise, results
from the visual discrimination test on offspring from 130 to 170 days of age
were negative. Prenatal exposure to PCE did not exert a detrimental effect on
any of the parameters of postnatal maturation examined.
7.1.3 Rabbits
Beliles et al. (1980) also investigated the teratogenic potential of PCE
in rabbits. The rabbits were divided into six groups as follows:
Days of Exposure
Group Cone (ppm) Pregestational Gestational
1
2
3
4
5
6
0 (control)
0 (control)
500
500
500
500
none
5 days/week;
none
5 days/week;
none
5 days/week;
3 weeks
3 weeks
3 weeks
0-21
0-21
0-21
0-21
0-21
0-21
After the pregestational exposure period was complete, rabbits from each group
were mated. A positive identification of spermatozoa in the vaginal canal was
taken as evidence of mating and designated as day 0 of gestation. The two
control groups (1 and 2) were exposed to filtered air and the exposure groups
3
(3, 4, 5, and 6) to 500 ppm (3390 mg/m ) PCE, 7 h per day. It is unclear if
the animals were exposed on each day, i.e., on days 0 through 21 of gestation
or only on a Monday through Friday basis. All rabbits could not have been mated
on the same calendar day; thus, all rabbits would not have been at the same stage
of gestation during the exposure period.
Mean body weights of rabbits during pregestational and gestational exposure
indicated no significant difference between controls and treated groups.
Reduced food consumption during the approximate period of days 10 through 22
of gestation was observed in groups 3 and 5 and may have been related to PCE
003PE2/E 7-6 11/22/83
-------
exposure. During the investigation, 21 female rabbits died (9 in the control
groups and 12 in the treatment groups). Placental abnormalities were found at
all exposure levels. In four litters examined in group 5, 19 anomalies were
found (p < 0.05 compared to group 1 and 2 controls.) Rank Sum analysis did
not confirm this significance and the authors judged it to reflect a change
confined to a few litters. Fifteen of the 19 abnormal placentas in group 5
occurred in two litters. Histopathologic evaluation failed to reveal any
significant change in the placentas. Of the skeletal changes noted in the
fetuses examined in each group, neither the frequency nor the character of
changes in the treated groups indicated an adverse effect on fetal growth and
development, nor any teratogenic effects.
7.2 SUMMARY
The mammalian animal tests performed to date do not indicate any signifi-
cant teratogenic potential of PCE. On this basis, there is no evidence that
suggest that the conceptus is uniquely susceptible to the effects of PCE. The
anatomical effects observed reflect delayed development and can be considered
reversible. The minor behavioral changes observed probably reflect maternal
nutritional deprivation rather than a direct effect of PCE. It is important
to note, however, that the reversible nature of an embryonic/fetal effect in
one species might, in another species, be manifested in a more serious and
irreversible manner. Thus, the teratogenic potential of PCE for humans must
be considered unknown.
003PE2/E 7-7 11/22/83
-------
7.2 REFERENCES
Beliles, R. P.; Brusick, D. J.; Mecler, F. J. (1980). Teratogenic-mutagenic
risk of workplace contaminants: trichloroethylene, perch!oroethylene, and
carbon disulfide. U.S. Department of Health Education and Welfare, Contract
No. 210-77-0047.
Buelke-Sam, J.; Kimmel, C. A. (1979). Development and standardization of
screening methods for behavioral teratology. Teratology 20:17-30.
Elovaara, E. ; Hemminki, K.; Vainio, E. (1979). Effects of methylene chloride,
trichloroethane, trichloroethylene, tetrachloroethylene and toluene on
the development of chick embryos. Toxicol. 12:111-119.
Lanham, S. Studies on placental transfer: trichloroethylene. Ind. Med.
39:46-49, 1970.
Manson, J. M.; Tepe, S. J.; Lowrey, B.; L. Hastings (1982). Postnatal evaluation
of offspring exposed prenatally to perchloroethylene. Unpublished.
Nelson, B. K.; Taylor, B. J.; Setzer, J. V.; Hornung, R. W. (1980). Behavioral
teratology of perchloroethylene in rats. J. Environ. Pathol. Toxicol.
3:233-250.
Schwetz, B. A.; Leong, B. K.; Gehring, P. J. (1975). The effect of maternally
inhaled trichloroethylene, perchloroethylene, methyl chloroform, and
methylene chloride on embryonal and fetal development in mice and rats.
Toxicol. Appl. Pharmacol. 32:84-96.
Tepe, S. J.; Dorfmveller, M. K.; York, R. G.; J. M. Manson (1982). Teratogenic
evaluation of perchloroethylene in rats. Unpublished.
U.S. Environmental Protection Agency. Proposed guidelines for registering
pesticides in the United States. 43 FR, #163, August 22, 1978. pp.
37382-37388.
U. S. Environmental Protection Agency. Proposed health effects test standards
for Toxic Substances Control Act test rules and proposed good laboratory
practice standards for health effects. 44 FR, #145, July 26, 1979, pp.
44089-44092.
U.S. Environmental Protection Agency. Determination not to initiate a rebut-
table presumption against registration (RPAR) of pesticide products con-
taining carbaryl availability of decision document. Fed. Regist. 45:
81,869-81,876, December 12, 1980.
003PE2/E 7-8 11/22/83
-------
8.1 MUTAGENICITY
The objective of this mutagenicity evaluation is to determine whether
tetrachloroethylene has the potential to cause mutations in germ and somatic
cells of humans. This qualitative assessment is based on data derived from
several short-term tests that measure different types of genetic alterations:
gene mutation, chromosomal aberrations, unscheduled DMA synthesis, and mitotic
recombination (Table 8-1). These tests were conducted using bacteria, Drosophila,
yeast, cultured mammalian cells, whole mammal systems, and cytogenetic analyses
of exposed humans. Consideration will also be given to studies concerning the
mutagenicity of known and expected metabolites.
8.1 GENE MUTATION TESTS
8.1.1 Bacteria
The ability of tetrachloroethylene to cause gene mutations in bacteria has
been studied by several investigators. Many of these investigators used the
Ames Salmonel1 a/microsome test or modifications of that test. Different purities
of tetrachloroethylene (stabilized* and low-stabilized materials) have been
evaluated. (Bacterial studies are summarized in Table 8-2.)
Tetrachloroethylene is a volatile chemical, and thus, the standard Ames/
Salmonella plate test, in which precautions are not taken to prevent escape of
evaporated material, is not entirely suitable for its testing. Williams and
Shimada (1983, sponsored by PPG Industries, Inc.), however, modified the standard
plate procedure by exposing the bacteria to the test agent in a sealed chamber.
*Stabilization is the intentional addition of material to increase the
stability of tetrachloroethylene. Typically, the added stabilizers are acid
and free radical scavengers (Dr. A. Philip Leber of PPG Industries, Inc.,
personal communication, September 1983).
8-1
-------
TABLE 8-1. SUMMARY OF MUTAGENICITY TESTING OF TETRACHLOROETHYLENE
A. Gene Mutation Tests Results*
Ames/Salmonella assay +**,-
Escherlchla coll K12/343/113 ( )
Multi-purpose test
Saccharomyces cerevlslae D7 reverse
mutation test (lvl-1 locus)
Drosophlla sex-linked recessive
lethal test
Host-mediated tests In mice: bacteria wk
yeast
B. Chromosomal Aberration Tests
Rat bone marrow assay
Mouse bone marrow assay
Peripheral lymphocytes from
exposed humans
Drosophlla sex chromosome loss
assay
Rat dominant lethal assay
_t
References
Williams and Shlmada 1983,
Margard 1978, SRI Inter-
national 1983, NTP 1983,
Bartsch et al. 1979, Cerna
and Kypenova 1977 (abstract)
Henschler 1977, Grelm et al.
1975
Call en et al. 1980,
Bronzettl et al. 1983
Bellies et al. 1980
Bellies et al. 1980,
Cerna and Kypenova 1977
(abstract)
Bronzettl et al. 1983
Rampy et al. 1978,
Bellies et al. 1980
Cerna and Kypenova
1977 (abstract)
Ikeda et al. 1980
Bellies et al. 1980
Bellies et al. 1980
C. Other Tests Indicative of DNA Damaging Activity
Unscheduled DNA synthesis In WI-38 -f
Hepatocyte primary culture/DMA repair *,-tt
test
Mltotlc recombination tests In +,-
Saccharotnyces cerevlslae D7
Bellies et al. 1980
Williams and Shlmada
1983, Williams 1983
Callen et al. 1980,
Bronzettl et al. 1983
D. DNA Binding Studies
Whole mice
E. Germ Cell Tests
Altered sperm morphology
mouse rat
Schumann et al. 1980
Bellies et al. 1980
* + designates positive; - negative; wk weak response. Dose-response
relationships were not established for the reported * results or wk results.
**AHhough Increases several fold over background were observed, the positive
results are considered weak because large amounts of material were needed to
elicit the responses. Positive results were only obtained using airtight
chambers (except for the study by Cerna and Kypenova 1977).
Questionable evidence for weak or borderline activity In specific data sets.
ftPos1t1ve results were found with vapor phase exposure and negative results
were obtained using conventional phase exposure.
8-2
-------
TABLE 8-2. RESULTS OF BACTERIAL TESTS OF DIFFERENT PURITIES AND SOURCES OF TETRACHLOROETHYLENE
co
Test system/strain
Ames/Salmonella
TA98, TA100, TA1535
Ames/Salmonella
TA98, TA100, TA1535
TA1537, TA1538
Purity/source
Perch! or- 200-
low-stabil ized
99.93% purity
PPG Industries,
Inc.
Perchlor 230-
stabil ized
99.80% purity
PPG Industries,
Inc.
Concentrations Metabolic
tested activation
1% v/v for TA98, Aroclor-
TA1538, and induced rat
TA1537; 0.1 S9 mix
1.0, 2.5, 5.0,
7.5, and 10%
for TA100 and
TA1535
1% v/v for TA98, Aroclor-
TA1538, and induced rat
TA1537; 0.1 S9 mix
1.0, 2.5, 5.0,
7.5, and 10%
Protocol
Gas-phase
exposure in
airtight
chamber's
Gas-phase
exposure in
airtight
chambers
Reported
result Reference
Positive
at 2.5% (>97%
toxicity) in
base-pair
substitution-
sensitive strain
(two to tenfold
increases
+/- M.A.*)
dose-response
not established
Positive
at 2.5% (>97%
toxicity)
in base pair-
substitution
Williams
and
Shimada
1983
Wi 1 1 i ams
and
Shimada
1983
for TA100 and
TA1535
sensitive
strains (three to
ten-fold increase
+/- M.A.)
dose-response
not established
Ames/Salmonella
TA100, TA1535
High purity
Perchlor
low-stabil ized,
99.98+% purity,
PPG Industries,
Inc.
0.1, 1.0, and
2.5% v/v
Aroclor-
induced rat
S9 mix
Gas-phase
exposure
in airtight
chambers
Negative
Wi11i ams
and
Shimada
1983
*+/- M.A. designates response similar in the presence and absence of
metabolic activation
(continued on the following page)
-------
TABLE 8-2. (continued)
Test system/strain
Ames/Salmonella
TA98, TA100, TA1535
TA1537, TA1538
Ames/Salmonella
TA98, TA100, TA1535
TA1537, TA1538
Purity/source
Nonstabilized
high purity
Detrex
Industries, Inc.
Stabilized
99.84% purity
Detrex
Industries, Inc.
Concentrations
tested
0.01, 0.05,
and 0.1 ml/
plate
0.01, 0.05,
and 0.1 ml/
plate
Metabolic
activation
Aroclor -
induced r-at
S9 mix
Aroclor -
induced rat
S9 mix
Protocol
Standar-d
plate test
in airtight
chambers
Standar d
plate test
in airtight
chambers
Reported
result
Negative
Refer ence
Margard
1978
Positive Margard
(twofold 1978
increases
in frameshift-
CO
sensitive
strains and
TA100) at 0.1
ml/pi ate (160
mg/plate). >90%
toxicity. S9 mix
increased r-esponse
(10-to 17-fold
increases)
Ames/Salmonella
TA98, TA100, TA1535
TA1537
Ames/Salmonella
TA98, TA100, TA1535
TA1537
99+% purity 0.025, 0.05,
Aldrich 0.1, 0.5, 1.0
and 1.5 added
to petri plate
at bottom of
desiccator-
Technical grade 3, 10, 33, 100,
Fisher 333 ug/plate
Ar odor-induced
female and male
rat liver S9 mix
and Aroclor-- in-
duced female and
male mouse liver
S9 mix
Aroclor- induced
rat liver S9 mix
Aroclor-induced
hamster- liver
S9 mix
Gas-phase Negative
exposure in
airtight
chambers
Preincu- Negative
bation assay
10 min. 37°C
SRI
Inter-
national
1983
NTP 1983
(continued on the following page)
-------
TABLE 8-2. (continued)
00
Test system/strain
Ames/Salmonella
TA100
Ames/Salmonella
tester- strains
not reported
E. coli K12/343/113 (>)
Host-mediated assay
ICR mice/Salmonella
TA1950, TA1951,
TA1952
Host-mediated assay
male and female CD
mice/Salmonella
TA98
Purity/source
99. 7% purity
Merck-Darmstadt
Not
reported
analytical grade
Merck-Darmstadt
Not
reported
91.43% purity
North Strong
Division
Chemicals
Concentrations
tested
0 to 663
mg/plate
(4xlO-3n)
Not
reported
0.9 mM
Not
reported
Inhalation at
100 ppm and
500 ppm for
5 consecutive
days
Metabolic
activation
Phenobar bital-
induced mouse
1 iver microsomes
with and without
cofactors
Phenobar bital-
induced mouse
1 iver- micro-
somes
Female mice
Male and
female mice
Protocol
Standard
plate test
Spot test
Liquid
suspension
2 hours at
37°C
Not
reported
Bacteria in-
jected intra-
per itoneally
after last
exposure and
removed 3
hours
Reported
result
Negative
Positive
Negative
Positive
(mortality
not
reported)
Positive
(twofold
increases
in revertants
from 100 ppm
in males;
and fourfold
increases
in revertants
from 500 ppm
females).
Reference
Bartsch
et al.
1979
Cerna and
Kypenova
(1977,
abstract)
Greim et
al. 1975
Cerna and
Kypenova
(1977,
abstract)
Beliles
et al.
1980
-------
In their procedure, a known volume of test chemical was added to a glass petri
plate containing a magnetic bar to ensure continuous stirring for even dissipation
of vapors. The chamber was initially incubated at room temperature for 20
minutes, and then exposure was continued at 37°C for 18 hours, after which the
bacterial test plates were removed from the chamber, covered with lids, and
incubated another 30-54 hours at 37°C.
In the Williams and Shimada (1983) study, three types of material (provided
by PPG Industries, Inc.) were tested in the presence and absence of S9 mix
(derived from livers of Aroclor-induced rats): Perchlor 200 (low-stabilized,
99.93% purity), high purity Perchlor (low-stabilized, 99.98+% purity), and
Perchlor 230 (stabilized, 99.80% purity).* Perchlor 200 and Perchlor 230 were
evaluated in tester strains TA98, TA1537, and TA1538 at 1% (v/v) and in tester-
strains TA100 and TA1535 at 0.1, 1.0, 2.5, 5.0, 7.5, and 10% (v/v). High-
purity Perchlor was evaluated in TA100 and TA1535 at 0.1, 1.0, and 2.5% (v/v).
These concentrations represent the predicted concentration of the compound in
the gas phase based upon calculations that consider the chamber volume and absolute
temperature of the chamber, atmospheric pressure, and density of the test compound.
(For example, 0.06, 0.62, 1.55, 3.1, 4.65, and 6.2 ml of tetrachloroethylene
was added to the chamber- for a 0.1, 1.0, 2.5, 5.0, 7.5, and 5% v/v vapor- tar-get
concentration, respectively.) Positive responses were obtained for Perchlor
200 and 230 at 2.5% v/v (or 1.55 ml per desiccator) but not for high-purity
Perchlor at the concentrations tested. A slightly higher response was observed
with stabilized material. Positive results were repeated in a second experiment
and were similar in the presence or absence of S9 mix. For example, in the
absence of S9 mix Perchlor 200 (at 2.5%) increased the number of revertant
colonies in the base-pair substitution-sensitive strains TA1535 (six to tenfold
increases) and TA100 (two to threefold increases); Perchlor 230 (at 2.5% v/v) also
caused increases in revertant numbers in TA1535 (approximately tenfold increases)
*Perchlor 200 contained 0.012 (% by weight) of hydroquinone monomethyl
ether (HQMME) which provides minimal stabilization; Perchlor 230 contained 0.011%
HQMME, 0.07% cyclohexene oxide, and 0.05% B-ethoxy propionitrile; high-purity
Perchlor contained 0.01% HQMME (written communication from Dr. A. Philip Leber
of PPG Industries, Inc., 1983).
8-6
-------
and TA100 (approximately threefold increases). (See Figure 8-1 for an illustra-
tion of these responses). Negative results were found in frameshlft-sensitive
strains (TA98 and TA1538) for both Perch!or 200 and 230 in the presence or
absence of S9 mix. At the next highest concentration (5% v/v) the revertant
counts decreased to zero as toxicity became an important factor in the tests.
The authors also indicated the test materials were toxic (> 97% killing) at
the concentration (i.e., 2.5% v/v) that induced the mutagenic responses as
determined by a simultaneous cytotoxicity test. Although the test materials
may be quite toxic at 2.5% v/v, the investigator's method of quantifying
cytotoxicity may not be accurate because the cell density that was used on
each plate for the determination of toxicity is several orders of magnitude
lower than that used for determination of mutagenicity (i.e., 107-109 cells).
Nevertheless, it should be noted that because of toxicity, the mutagenic
responses observed for Perchlor 200 and 230 are within a narrow range of
concentrations. Because of this narrow range of effective concentrations and
the concentration increments tested, a clear dose-response was not demonstrated.
In the Williams and Shimada (1983) study, high-purity material did not
produce a detectable response under- experimental conditions, as did the lower-
purity materials (i.e., Perchlor 200 and 230). Therefore, the increases in
revertants observed may be due to a contaminant(s). However, the high-purity
Perchlor appeared to be more toxic (zero revertants observed at 2.5%) than the
other test materials, and it is possible that a weak mutagenic response may
have been masked by its toxicity. A weak response was observed for high-purity
material at 1% in TA1535 (with S9 mix), but was not repeatable. Also, it
should be pointed out that concurrent negative and positive controls
were not used in this study. This weakens the negative conclusions for the
high-purity material and makes interpretation of the magnitude of the responses
in the presence of S9 mix for the lower-purity materials difficult.
Margard (1978) examined tetrachloroethylene (provided by Detrex Chemical
Industries, Inc.) in the Ames/Salmonella test to determine whether- mutagenic
activity of technical grade samples is the result of added stabilizers.
Precautions to prevent escape of material were taken (Margard, personal
communication 1981), but they were not specified in writing. Tester strains
TA1535, TA1537, TA1538, TA98, and TA100 were used. Tests were conducted in
8-7
-------
TA153S
300
mo
100
0
ut
a.
i
4
ec
UJ
>
UJ
CC
C
uj
TA1535
300
no
100
2.6
1041
CD
I
CXI
GOO
500
400
30O
200
100
TA 100
300
2OO
100
0.1 2.8 5.0 7.6 10.0
TETRACHLOROETHYLENE. Stabilized (%V/V)
TA100
0.1 1.0 2.6 5.0 7.5 10.0
TETRACHLOROETHYLENE. Low-Stabilized (%V/V)
Figure 8-1. Oose-response curves for Perchlor 200 (low-stabilized tetrachloroethylene) and Perchlor 230
(stablllz tetrachloroethylene) using Salmonella typhlmurlum tester strains TA100 and TA1535 in the
presence of (O--O) and 1n the absence of (•—•) 59 mix (150 ul of Aroclor-lnduced rat liver).
Each data point represents the geometric mean of triplicate plates from one experiment. (Williams
and Shimada 1983)
-------
the presence and absence of S9 mix (prepared from livers of Aroclor-1254 induced
rats). Nonstabilized and stabilized materials were added directly to the
petri plates at 0.01, 0.05, and 0.1 ml per plate rather than vapor exposure.
The nonstabilized test material was described as purified tetrachloroethylene
that contained no detectable epoxides or other stabilizing components, and the
stabilized material was identified as an industrial degreasing grade of
tetrachloroethylene that contained 0.07% (by weight) epichlorohydrin, 0.007%
N-methylmorpholine, 0.07% beta-hydroxypronitrile, and 0.01% hydroquinone monoethyl
ether (written communication from L. Schlossberg, Detrex Chemical Industries
Inc., 1981). Nonstabilized test material was not detected as positive in the
presence or absence of S9 mix. But stabilized material at 0.1 ml per plate
(equivalent to 160 mg*) caused a weak response in the absence of S9 mix;
twofold increases were observed for TA1538, TA98, and TA100. In the presence
of S9 mix, greater increases in the number- of revertants for- tester- strains
TA1538 (17-fold increase), TA98 (10-fold increase), and TA100 (1.7-fold increase)
at 0.1 ml per plate, were found. A clear dose-response was not observed for
any of the tester strains. The toxicity of the test material was reported as
greater than 90% killing at concentrations which caused increases in the number
of revertants. The method used to determine toxicity may be inaccurate, as
discussed previously for the Shimada and Williams study (1983). Negative
results were reported for stabilized material in TA1535 and TA1537. Although
the stabilized test material contained epichlorohydrin, which has been shown
to be strongly mutagenic in Salmonella, it is primarily active in base-pair
substitution-sensitive strains (i.e., TA1535 and TA100; McCann et al. 1975,
Anderson et al. 1978), and thus it does not seem likely that this agent can
solely account for the activity observed for stabilized material in the
frameshift-sensitive strains TA1538 and TA98. Nevertheless, it appears
that the mutagenic activity is due to the presence of a contaminant(s).
Other- investigators have conducted studies in which precautions were
taken to prevent escape of test material during testing. SRI Inter-national
(1983 EPA-sponsored test) reported tetrachloroethylene (99+% purity,
Calculation based on the density of tetrachloroethylene as 1.586 g/ml.
8-9
-------
Aldrlch) to be negative when tested as a vapor In sealed desiccators using
TA1535, TA1537, TA98, and TA100 with or without S9 mix prepared from livers of
female and male Aroclor-1254 induced rats and mice. Cells were exposed at
37°C for 8 hours in desiccators, and then incubated at 37°C for an additional
42 hours. The concentrations tested were 0.05, 0.1, 0.5, 1.0, 1.5 ml added to
a petri plate at the bottom of the desiccator chamber in one experiment and
0.025, 0.05, 0.1, 0.5, and 1.0 ml added to a plate at the bottom of the desiccator
in another experiment. A toxic response (reduction or absence of bacterial
lawn) was found at 1.5 ml/per desiccator.
The National Toxicology Program (NTP 1983) sponsored Salmonella testing
on tetrachloroethylene (technical grade, source Fischer) and obtained negative
results. Four standard tester strains were used: TA98, TA100, TA1535, and TA1537.
A preincubation protocol was followed in which the cells, S9 activation system,
and chemical are preincubated in test tubes for 20 minutes at 37°C before
addition of the top agar and plating in petri plates. Two different S9 systems
were used; S9 mix derived from livers of Aroclor-1254 induced rat liver and
Aroclor-1254 induced hamster liver. Tests were also conducted in the absence
of S9 mix. Six concentrations (0, 3, 10, 33, 100, 333 ug/plate) were evaluated.
(In TA100, up to 10,000 ug/plate was evaluated.) Although incubation was
carried out in capped tubes, it is possible that evaporation and some escape
may have occurred. However, it should be noted that toxic levels were tested
as indicated by absence or reduction of bacterial lawn.
Bartsch et al. (1979) investigated the mutagenicity of tetrachloroethylene
(99.7% purity, Merck-Darmstadt) using the standard Ames/Salmonella plate
test in which precautions to prevent escape of test material are not taken.
Negative results were obtained using tester strain TA100 in the presence of
mouse liver microsomes with and without cofactors. The authors indicated that
toxic concentrations (above 82.9 mg/plate) were tested, but did not indicate
their- criteria for determining toxicity. Interpretation of these negative
conclusions is limited by the amount of data presented and because only one
tester strain was evaluated.
In an abstract, Cerna and Kypenova (1977) reported that in a spot test
protocol of the Ames/Salmonella assay, tetrachloroethylene (purity and source
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not reported) induced both base-pair substitution and frameshlft mutations in
Salmonella typhimurium. The results were obtained in the absence of exogenous
activation. These authors also reported that in a host-mediated assay using
female ICR mice, tetrachloroethylene induced significant increases in the
number of revertants in tester strains TA1950, TA1951, and TA1952 at dosage
levels reported as representing the 1059 and one-half of the 1059. These
results were reported as not being dose-dependent. Because this report was in
abstract form and did not provide details of the protocol nor present the
data, the acceptability of the test results is indeterminate. Also, the
possibility of mutagenic contaminants must be considered.
Beliles et al. (1980) used a host-mediated assay to evaluate the effects
of tetrachloroethylene (91.43% pure from North Strong Division Chemicals) in
the presence of whole mammal metabolism. In this study, Salmonella tester
strain TA98 was used as the indicator organism, and male and female mice (strain
CD-I) were the hosts. The animals were exposed by inhalation 7 hours per day
for 5 days to either 100 ppm or 500 ppm. Bacteria were injected intraperitoneally
into the mice after the last exposure. The bacteria were removed from the mice
3 hours later. At the 100 ppm dose level, increases in the number of revertants
were observed for males (approximately twofold increase) but not for females.
At 500 ppm, a positive response was reported for females (approximately fourfold
increase) but not for- males. Because of the lack of a dose-response and the
weak responses, these findings should be viewed with caution. Also, the material
used was of low-purity, and the responses may have been due to contaminants
and/or added stabilizers. In addition, parallel in vitro plate tests using
Salmonella were not conducted. The parallel plate tests are important the
determination of the requirement of whole-mammal activation.
Studies on tetrachloroethylene have also been conducted in Escherichia coli.
Henschler (1977) reported that tetrachloroethylene was not mutagenic when tested
in E. coli K12 with metabolic activation (liver microsomal fraction prepared
from phenobarbital-induced mice). This report, however, is difficult to evaluate
because actual revertant count data (experimental and control number of revertants)
and the details of the protocol are not provided. It appears that the conclusions
presented in this report are actually based on data derived from the study by
Grelm et al. (1975).
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Greim et al. (1975) reported that negative results were obtained when
tetrachloroethylene (purity reported as analytical grade; Merck-Darmstadt)
was assayed in the multi-purpose test system of Escherichia coli K12/343/113 (>).
Tests were conducteo in the absence and presence of metabolic activation
(phenobarbital-induced mouse liver microsomal fraction plus TIADPH cofactors)
at a concentration of 0.9 mM and a treatment time of 2 hours in liquid suspension
at 37°C. These treatment conditions resulted in 99 + 1% survival. The genetic
markers evaluated for mutation induction were the missense marker arg+ and
the frameshift marker nad+ for reverse mutation, and gal+ and MTR for forward
mutation. Deficiencies in the study design and reporting of the results
reduce the weight of the negative conclusion. These deficiencies are as
follows: 1) only one concentration was evaluated, 2) adequate exposure may
not have been achieved, as indicated by the high survival, 3) there was no
reporting of revertant count data (experimental and control), and 4) there
was no reporting of the number of replicate plates used or the number of the
tests conducted.
The bacterial tests discussed above do not clearly demonstrate that
tetrachloroethylene Itself is mutagenic. The positive responses found may
be due to contaminants and/or added stabilizers. The induced increases in
revertants did not require exogenous metabolic activation and were observed in
both frameshift and base-pair substitution-sensitive tester strains. The
positive findings were not considered strong in that large amounts of material
(estimated vapors at 2.5% v/v and at 160 ing/plate) were needed for the detection
of mutagenicity. Also, there was a very narrow range of effective concentrations
because of toxidty; thus dose-response relationships were not established.
When tested, highly purlfed samples were not detected as mutagenic under the
conditions 1n which technical samples caused increases 1n the number of revertants.
Although some technical tetrachloroethylene samples were positive, there were
other samples that were not detected as positive. The available results provide
suggestive evidence that certain technical samples of tetrachloroethylene are
weakly mutagenic 1n Salmonella and that the positive responses may be due to
impurites and/or added stabilizer's.
8.1.2 Drosophila
Bellies et al. (1980) used the sex-linked recessive lethal assay in
DrosophUa melanogaster to test tetrachloroethylene (91.43% purity, North Strong
Division Chemicals) and reported negative results. Adult male flies were
exposed for 7 hours by Inhalation at 100 ppm and 500 ppm. Treated males were
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mated to nontreated females at various times (2-3-3-2 day mating scheme) to
test specific germ cell stages. No significant increases (P < 0.05) over the
background values were observed. However, only a small sample size was examined.
A total of 3804 chromosomes for the 100 ppm dose and 3956 chromosomes for the
500 ppm dose were evaluated. This sample size was only large enough to
exclude the induction of an approximately fourfold increase in mutation frequency
(Kastenbaum and Bowman 1970). Ideally, at least 7000 chromosomes at each dose
level should be screened to preclude a doubling in mutation frequency, which is
generally considered to be the increase of biological significance. Survival
was not reported in this study, and thus it is uncertain whether a sufficient dose
was given. These deficiencies prevent a judgment regarding the mutagenic
activity of tetrachloroethylene in Drosophila.
The ability of tetrachloroethylene to cause gene mutations in an eucaryotic
organism has not been adequately examined. The only available study was a sex-
linked recessive lethal test in Drosophila in which tetrachloroethylene was not
properly evaluated.
8.2 CHROMOSOMAL ABERRATION TESTS
8.2.1 Whole-Mammal Bone Marrow Cells
Rampy et al. (1978) examined bone marrow cells for chromosome aberrations
from male and female Sprague-Dawley rats after tetrachloroethylene exposure.
Animals were exposed to 300 ppm (2.03 mg/1) or 600 ppm (4.07 mg/1)
tetrachloroethylene* by inhalation 6 hours/day, 5 days/week, for one year.
Three animals per dose were examined. The authors reported "zero" chromosomal
aberrations pet cell for both females and males. The data for females, however,
are inadequate for a clear interpretation because of the very low number of
metaphases scored (less than 25 cells per animal). In male rats, 150 cells
were scored (50 cells per animal). The negative controls were also reported
as "zero" aberrations. This observation is very unusual because, in general,
most laboratories have reported about 1 to 2% total aberrations for background
values. In this study, it is not known whether the highest exposure level was
near the maximum tolerated dose (MTD) for females because no weight loss and
no mortality was observed. In males there was no weight loss, but significant
increases in mortality above control values were observed at the highest dose
*Formulation (liquid volume percent): trichloroethylene, 3 ppm; hexachloro-
ethane, < 12 ppm; carbon tetrachloride, 2 ppm; 4-methyl morpholine, 44 ppm; non-
volatile residue, 2 ppm; and tetrachloroethylene balance.
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tested, and therefore an MTD may have been approached. It is not apparent that
the Investigators determined the toxicity of the test material to arrive at an
MTD for this study, because the dosage levels used were based on the threshold
limit value of 100 ppm for tetrachloroethylene.
A rat bone marrow assay was also performed by Bellies et al. (1980) and
reported as negative. Ten males and ten females [CRL:COBS CD(SD)BR] were
exposed to an acute dose of 100 ppm and 500 ppm tetrachloroethylene (91.43%
purity, North Strong Division Chemicals) by Inhalation for 7 hours. Bone
marrow cells were harvested 6, 24, and 48 hours later. No increase in aberrations
was found for females, but for males, weak clastogenic effects (breaks, fragments,
deletions, and aneuploid cells) were observed. At 500 ppm, 3.3% cells with
aberrations versus 0.7% in the control were found at the 24-hour kill. A
subchronic study (five exposures, 7 hours per day) at 100 ppm and 500 ppm was
also conducted. Animals were killed 6 hours after the last exposure. No
increase in abnormalities was found in males, and only a slight increase at
100 ppm (1% cells with aberrations) was found in females. This response was
not dose-related. The female subchronic control group had a very low background
(0.3% cells with aberrations). Although isolated incidences of increases in
chromosomal aberrations were observed, the lack of dose-responses precludes an
unequivocal positive conclusion. On the other hand, a negative conclusion
cannot be drawn because the authors did not discuss the criteria used to select
the dosage levels, and thus, there is the possibility that a toxic dose may
not have been evaluated.
Cerna and Kypenova (1977) reported in an abstract that mice (ICR)
given an acute intraperitoneal dose one half of the 1050 of tetrachloroethylene
or dosed intraperitoneally for five applications in 24 hour intervals (dose of
one injection equalled 1/6 LDso) did not show cytogenetic effects in the bone
marrow cells. Details of the protocol and the cytogenetic data were not available
for an evaluation. Hence, the negative conclusion of the authors cannot be
considered definitive.
8.2.2 Human Peripheral Lymphocytes
Ikeda et al. (1980) studied chromosomal aberrations, sister chromatid exchanges
(SCEs), and variation 1n the mitotic index of peripheral lymphocytes cultured from
ten workers (seven males and three females) occupationally exposed to technical grade
tetrachloroethylene (impurities not reported). The workers were divided into
high (Group 1) and low (Group 2) exposure groups. Group 1 consisted of six
workers (five males and one female aged 20 to 66 years) from a degreasing workshop.
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These workers had a geometric mean exposure of 92 ppm (range 30 to 220 ppm). The
five males of Group 1 had work histories of 10 to 18 years, whereas the one female
had worked in the degreasing shop for only 1 year. Group 2 included four
workers (two males and two females aged 17 to 31 years) from a support department
with a shorter work history (3 months to 3 years) and with an exposure range
of 10 to 40 ppm. The control group consisted of six males and five females.
The authors did not indicate if this was a matched control, and did not indicate
the medical histories of the subjects (e.g., recent illnesses, radiation
exposures). There were no statistically significant differences (P > 0.05)
in the incidences of chromosomal aberrations (structural and numerical) and
SCEs between the exposed workers and control group. The mitotic index was
similar for both exposed and control groups.
8.2.3 Drosophila
In addition to performing the sex-linked recessive lethal assay discussed
earlier, Bellies et al. (1980) also conducted a sex chromosome loss assay on
low-purity (91.4%, North Strong Division Chemicals) tetrachloroethylene for
nondisjunction in Drosophila melanogaster. Males were exposed to 100 and 500 ppm
of tetrachloroethylene for- 7 hours. The phenotypic classes used allow for
detection of losses of the entire X or- Y chromosome and the short or- long arm of
the Y chromosome. Although marginal increases, which were not dose-related,
were observed after tetrachloroethylene treatment, they are not considered
sufficient to judge the data as positive or- negative.
The cytogenetic tests discussed above using mice, rats, Drosophila, and
exposed humans have been reported as negative. Although these studies are
not considered to be a thorough evaluation of the ability of tetrachloroethylene
to cause chromosomal aberrations, the data collectively indicates that
tetrachloroethylene is not strongly clastogenic. However, there have been no
adequate studies on the ability of tetrachloroethylene to cause chromosome
nondisjunction (aneuploidy).
8.3 OTHER TESTS INDICATIVE OF DNA DAMAGE
8.3.1 DNA Repair
Unscheduled DNA synthesis (UDS) is measured by repair of DNA lesions,
which is indicative of DNA.damage. Bellies et al. (1980) assessed the ability
of tetrachloroethylene (91.43% purity, North Strong Division Chemicals) to
cause UDS in human fibroblast (WI-38) cells. Because WI-38 cells have little
if any enzyme activation capability, the tests were conducted with an exogenous
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source of metabolic activation (i.e., S9 mix). The test material was examined
at 0.01, 0.05, 0.1, and 0.5% (v/v) using conventional liquid phase exposure.
Scheduled DNA synthesis was blocked by treatment with hydroxyurea, and DOS was
measured by liquid scintillation counting of incorporated tritiated thymidine
(C3H]-TdR) into DNA. A very slight increase was seen at 0.01% (v/v)
tetrachloroethylene both in the presence (1.5-fold of control) and absence
(1.35-fold of control) of Aroclor rat liver S9 mix. No increases occurred at
0.1 and 0.5%. A toxic response was reported at 0.5% (i.e., a decrease in
total amount of DNA as well as in the Incorporation of [3H]-TdR). An increase
in total amount of DNA at 0.01% was found, suggesting that more cells are entering
the S phase. Cells accelerated into S phase would account for the increases
seen in [3H]-TdR incorporation at 0.01%. Therefore, it is uncertain whether
tetrachloroethylene induced a weak UDS response in this system. Also, other
problems were found in this study. The positive controls gave weak responses.
N-methyl-N'-n1tro-N-nitrosoguanidine elicited a weak response (1.8-fold
increase), but at a toxic concentration (5 ug/ml) as indicated by a decrease
1n total DNA (6.27 ug of DNA versus 23.76 ug of DNA in solvent control).
Benzo[a]pyrene gave a response of 1.16-fold above background. This is an
equivocal response, particularly because there was also an increase found in
the total amount of DNA.
Williams and Shimada (1983, sponsored by PPG Industries, Inc.) evaluated two
different samples of tetrachloroethylene, Perchlor 200 (low-stabilized, 99.93%
purity) and Perchlor 230 (stabilized, purity 99.80%), for their ability to
cause UDS 1n the hepatocyte primary culture (HPO/DNA repair test. The target
cells in this test system have a capability to metabolize xenobiotics.
Williams and Shimada measured UDS by autoradiographic determination of the
amount of [3H]-TdR (10 uCi/ml) incorporated into nuclear DNA. Hepatocytes
were isolated from adult male Fischer 344 rats. Cells were treated for 18 hours
or 3 hours. For the 3-hour exposure, cultures were incubated another 15
hours in the absence of tetrachloroethylene to allow for DNA repair synthesis.
The criterion that was used for a positive result was a net nuclear grain count
of 5 in triplicate coverslips.* Negative results were reported for both Perchlor
230 (at concentrations of 0.001, 0.01, 0.1, and 1.0% v/v) and Perchlor 200 (at
*Nuclear grain counts were reported as the mean +_ standard deviation.
Cytoplasmic grain counts 1n three nuclear size areas adjacent to the nucleus
were determined. The highest cytoplasmic grain cbunt was subtracted from the
nuclear count. This value is referred to as "net" grain count.
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concentrations of 0.0001, 0.001, 0.1, and 1.0%) when tests were conducted
using conventional liquid-phase exposure in which tetrachloroethylene was
added to the culture medium.
For both materials, positive responses were reported when testing was
performed using vapor-phase exposure in gas tight chambers. Testing was conducted
at 0.1, 1.0, and 2.5% v/v (desired air concentration) for 3 and 18 hours.
Perchlor 230 at 0.1% caused an increase in UDS when the cells were exposed for
3 hours (6.2 +_ 4.9 net nuclear grain count, 50% cells showing toxic effects*)
and for 18 hours (15.9 +_ 1.6 net nuclear- grain count, 25% cells showing toxic
effects). Perchlor 200 caused an increase in UDS at 0.1% v/v for the 3-hour
treatment (10.8 +_ 6.1% net nuclear grain count, 75% cells showing toxic effects)
but not for the 18-hour treatment. At 1.0% and 2.5% for both materials at
both treatment times, nearly 100% of cells showed toxic effects. It should be
pointed out that the background control values were taken from the conventional
exposure experiment. This makes intrepretation of the results difficult,
particularly because the conventional-phase media was different from the gas-
phase media. Because these data are based on one test in which there were no
concurrent positive and negative controls, the results are considered only
suggestive of a positive effect. To validate these findings it is necessary
to repeat experiments using appropriate concurrent controls and a concentration
range to demonstrate a dose-response. Although the possibility exists that an
impurity(ies) may be responsible for- the observed effects, the authors, did not
examine high-purity Perchlor (99.98+%) as they did in the Salmonella tests
discussed earlier-. If high-purity material had tested negative under the same
experimental conditions under which the lower-purity materials tested positive,
as in the Salmonella test, a stronger argument could be made that impurities were
causing the effects.
Williams (1983) conducted HPC/DNA repair tests using hepatocytes from male
B6C3F1 mlce and ma1e Osborne-Mendel rats on tetrachloroethylene (99+% purity,
Aldrich Chemical Company). Conventional liquid-phase exposure conditions
were used. Negative results were reported for both rat and mice hepatocytes
when tetrachloroethylene was added to the culture medium at 0.00001, 0.0001,
'Toxicity identified by the absence of S phase cells and general cellular-
morphology. This is not an accurate method of determining toxicity.
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0.001, 0.01, and 0.1% (v/v) for 18 hours. The material was reported as "toxic"
at 0.01% and higher. The highest cytoplasmic grain count was subtracted from
the nuclear count. This reduces the possibility of false positives, but the
chance of missing a weak UDS inducer would be increased, especially if the
cytoplasmic grain count is high. This consideration also applies to the Williams
and Shimada conventional-phase exposure study discussed previously. The cytoplasmic
grain counts were not reported in any of the HPC/DNA repair studies.
8.3.2 Mitotic Recombination
Call en et al. (1980) evaluated the ability of tetrachloroethylene (purity
not reported, stabilized 0.01% thymol, Eastman Kodak) to cause mitotic gene
conversion (nonreciprocal recombination) at the trp-5 locus, mitotic crossing-
over (reciprocal recombination) at the ade-2 locus, and gene mutation (reversion)
at the ilv-1 locus in log phase cultures of Saccharomyces cerevisiae D7.
Cells were incubated for one hour in culture medium containing 0, 4.9, 6.6, and
8.2 mM tetrachloroethylene. No exogenous source of metabolic activation was
used in these studies. At 4.9 mM (84% survival), no significant increases
in the frequency of gene conversion (1.9 convertants/lO^ survivors versus 1.4
convettants/lO^ survivors in background control) and mitotic crossing-over
(5.3 mitotic recombinants/lO4 survivors versus 3.3 mitotic recombinants/104
survivors in background control) occurred. As shown in Figure 8-2, when the
concentration of tetrachloroethylene was increased to 6.6 mM (58% survival),
increases in mitotic gene conversion and mitotic crossing-over did occur (8.3
convertants/105 survivors and 52.6 mitotic recombinants/104 survivors,
respectively). Mitotic recombinational activity was not determined at 8.2 mM
because less than 0.1% survival was found. No significant increases in gene
mutations were observed at 4.9 mM (3.8 revertants/106 survivors versus 2.9
revertants/106 survivors in control). The reverse mutation frequency was
not determined at 6.6 mM. Therefore, the induced number of revertants is too
low to be indicative of a positive result, but the high values for the
recombinational events do indicate a positive effect at 4.9 mM test material.
The possibility that the effects were caused by a mutagenic impurity(ies) . .
should be considered. It should be pointed out that in the study by Callen et
al. the Ade* recombinants were estimated from a total of 30 plates, ten of
which contained minimum medium (plus adenine and isoleucine) used for estimating
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4 6
TETRACHLOROETHYLENE (mM)
10
Figure 8-2. Induction of mitotic recombination by tetrachloroethylene in
Saccharomyces cerevisiae D7. The frequency of mitotic crossing-
over (•—•} and gene conversion (A---A) was determined. Log-
phase yeast cells were treated with test chemical for one hour
without the addition of an exogenous metabolic activation system.
(Adapted from Call en et al. 1980)
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the number of trp-5 convertants. Because mitotic crossing-over- and gene conversion
are not necessarily distinct events, in that they probably depend on a common
inducible mechanism, the number of Ade+ recombinants may have been overestimated
by including those Ade+ that were already Trp+. There is experimental support
that these recombinational events are inducible by a common factor(s) (see Fabre
Fabre and Roman 1977, Fabre 1978). The positive findings of Call en and coworkers
for mitotic recombination should be confirmed by repeating the assay using
appropriate selection conditions. In addition, because the responses were
observed within a narrow "window," at least one additional concentration between
the low and high doses used should be tested to demonstrate a clear dose-response.
Bronzetti et al. (1983) also used the yeast S. cerevisiae D7 to evaluate
the effects of tetrachloroethylene (99.5% pure, stabilized with 0.01% thymol,
Carlo Erba Co.) at the trp-5, ade-2, and ilv-1 loci. In this study, however,
negative results were obtained after a 2-hour treatment at 5, 10, 20, 60, and
85 mM in the presence or- absence of S9 mix (Aroclor-1254 induced rat liver).
Differences in the experimental protocol from that used by Call en et al. may
explain these negative findings. Bronzetti et al. used cells in the
stationary phase of growth rather than in the log phase of growth. For some
chemicals it has been observed that stationary cells are more refractory than
log cells to mutagenic treatment (Mayer and Coin 1980, Shahin 1975).
Bronzetti et al. were able to test higher concentrations than Callen and
coworkers. In the absence of S9 mix, Callen et al. reported less than 0.1%
survival at 8.2 mM, while Bronzetti et al. did not observe complete killing
until 85 mM. Another difference between these two studies is the source of
tetrachloroethylene. Bronzetti et al. purchased tetrachloroethylene from
Carlo Erba Co. (Milan, Italy) and Callen et al. obtained the test agent from
Eastman Kodak Co. (Rochester, N.Y.); thus, it is possible that the test samples
may have contained different impurities.
Bronzetti et al. (1983) also obtained negative results in an intrasanguineous
host-mediated assay using _S. cerevisiae 07 as the indicator organism and CD-I
mice as the host. Stationary yeast cells (4 x 10^ cells) were injected into
the retro-orbital sinus of mice. After injection of the yeast, an acute oral
dose of 11 g tetrachloroethylene/kg body weight (b.w.) was given. A subacute
dose of 2 g/kg b.w. given 5 days a week for a total of 12 administrations was
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also used (the last test dose was 4 g/kg b.w.; therefore, the total dose was 26
g/kg b.w.). In the subacute study, yeast cells were injected after the last
dose of tetrachloroethylene (i.e., 4 g/kg b.w.). Four hours after injection of
yeast, in both the acute and subchronic study, the cells were recovered from
the liver, lungs, and kidneys of three animals. No concurrent positive control
chemical was used in these studies to ensure that the system was functioning
properly.
The positive responses discussed above for mitotic recombination in yeast
and DMA repair synthesis in mammalian cells provide suggestive evidence that
certain technical samples of tetrachloroethylene may be active in damaging DMA.
However, toxic concentrations of material were needed to elicit these responses.
The possibility that impurities and/or added stabilizers caused the increased
effects should be considered.
8.4 DMA BINDING STUDIES
Chemical adduct formation is a critical step in certain types of mutagenesis.
Schumann et al. (1980) reported that there was no detectable binding of ^4C-
labeled tetrachloroethylene (99% purity) to DNA when mice were treated by inhala-
tion at 600 ppm for 6 hours or when they were given an acute dose of 500 mg/kg
orally. The specific activity of the 14C-label (26,593 dpm/umol in the
inhalation study and 133,273 dpm/umol in the oral study) is too low, however,
to preclude the possibility of very low levels of DNA binding. For example, at
the specific activities used under the experimental conditions, at an assumed
binding level of 10~5 alkylations per nucleotide (Stott and Watanabe 1982)
there would be 5-10 dpm per- sample. This is at the limit of practical
detection. Therefore, the possibility of binding at slightly less than 10~5
alkylations per- nucleotide cannot ruled out. These negative findings, however,
are consistent with the negative and weak results reported in the mutagenicity
tests discussed above.
8.5 STUDIES INDICATIVE OF MUTAGENICITY IN GERM CELLS
An important aspect of a mutagenicity evaluation is to assess the potential
of the chemical to reach the germinal tissue of humans and cause mutations
that may contribute to the genetic disease bur-den. This assessment is almost
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always based on animal experimentation. The ability of tetrachloroethylene to
cause genetic damage in germinal tissue has not been well studied. The only
test results available were from a dominant lethal study in rats [CRL: COBS
CD(SD)BR] and a sperm morphology assay in both rats and mice (strain CD-I).
Tetrachloroethylene (91.43% purity, North Strong Division Chemicals) did not
cause an increase in dominant lethals in rats when unexposed females were
mated during a 7-week period to exposed adult males given an acute dose of 100
ppm and 500 ppm by inhalation for 7 hour's per day for 5 days (Beliles et al.
1980). The dominant lethal assay is generally thought to measure gross chromosome
damage (Bateman and Epstein 1971). Also, this test is not considered a sensitive
assay because of the high spontaneous level of lethal events (Russell and
Matter 1980), and thus, negative results do not necessarily indicate that the
chemical does not reach and damage the germ cell DMA.
Beliles et al. (1980) also examined tetrachloroethylene for altered sperm
morphology in treated rats and mice. After- dosing at 100 ppm and 500 ppm by
inhalation 7 hours per day for 5 consecutive days, groups of four animals
were killed at the end of 1 week, 4 weeks, and 10 weeks to examine effects on
various germ cell stages. Sperm was collected from the cauda epididymis, and
at least 500 cells were examined. Negative results were reported in rats;
the mice, however-, showed positive responses. At 500 ppm, 19.7% abnormal
sperm were observed (versus 6.0% in the negative controls) during the fourth
week after exposure (corresponding to the spermatocyte stage). By themselves,
these positive findings alone are not sufficient to conclude that tetrachloroethy-
lene alters germ cell DMA because this assay is only an indicator of chemical
effects on sperm and does not provide definitive evidence that a chemical
reached germinal tissue and damaged DNA. Therefore, because limited information
was provided, it is not clear whether tetrachloroethylene (or impurities)
reaches the germ tissue. However, if tetrachloroethylene reached germ cell
DNA, there may be no serious risk of mutation because of the largely negative
or marginal results found in mutagenicity tests discussed previously.
8.6 MUTAGENICITY OF METABOLITES
Trichloroacetic acid (TCA) is a known human metabolite of tetrachloroethylene.
The formation of TCA is thought to occur through the formation of an epoxide,
tetrachloroethylene oxide, and its subsequent rearrangement to trichloroacetyl
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chloride or trlchloroacetaldehyde, which then rapidly hydrolyzes to TCA. (See
chapter 5 on metabolism for a more detailed discussion.) These intermediates are
considered relevant in assessing the mutagenicity of tetrachloroethylene.
Tetrachloroethylene oxide, which is considered to be the biologically active
intermediate of the parent compound tetrachloroethylene, was assayed for' mutagenicity
1n the absence of exogenous metabolic activation in several bacterial tests
(Kline et al. 1982). Tetrachloroethylene oxide increased the number of revertants
in a dose-dependent manner in Salmonella tester strain TA1535 when assayed by
a preincubation liquid protocol (20 minutes at 37°C), but did not cause an
Increase in revertants using Escherichia coli WP2 uvr A. In Salmonella, a
14-fold increase occurred at 2.5 mM and a 20-fold increase occurred at 5mM.
Tetrachloroethylene epoxide was toxic at 25mM.
This epoxide was also evaluated by Kline et al. (1982) in the E. coli pol
A assay. Positive effects were observed at 0.04, 0.09, and 0.44 mM/ml as
measured by differential growth inhibition of a DMA polymerase-deficient strain
in comparison with its polymerase-proficient parent.
Waskell (1978) tested TCA at 0.45 mg/plate in Salmonella TA98 and TA100
and obtained negative results. There are other intermediates which have not
been identified in humans but are thought to occur- (e.g., trichloroethanol
chloral hydrate). Waskell (1978) obtained negative results for trichloroethanol
1n Salmonella TA98 and TA100 up to a dose of 7.5 mg/plate. Gu et al. (1981),
however-, reported suggestive evidence that trichloroethanol weakly induced
sister chromatld exchange (SCE) formation in primary cultures of human lympho-
cytes.
Chloral hydrate was reported to be marginally mutagenic (less than twofold
increase over- a dose range of 0.5 to 10 mg per plate) in Salmonella TA100 (Waskell
1978). Gu et al. (1981) also provided suggestive evidence than chloral hydrate
at 54.1 mg/1 caused a weak increase in SCEs in cultured human lymphocytes.
Chloral hydrate has also been shown to block spindle elongation in insect
spermatocytes (Ris 1949). Data on metabolites of tetrachloroethylene
suggest that if the parent compound was biotransformed, its metabolites may be
genotoxic; these data are limited, however, and additional studies are needed
on metabolism and on the mutagenicity of metabolites to reach a clear- conclusion.
8-23
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8.7 SUMMARY AND CONCLUSIONS
Tetrachloroethylene itself has not been clearly shown to be a mutagen.
Certain commercial and technical preparations have elicited positive responses
in the Ames bacterial test, a yeast recombinogenic assay, a host-mediated
assay using Salmonella, and DMA repair assays. In general, the responses were
weak, and eliciting them required rather high toxic concentrations of
tetrachloroethylene. No dose-response relationships were established in
these studies. The positive findings may be explained by the presence of
mutagenic contaminants and/or- added stabilizers. Highly purified tetrachloro-
ethylene has only been evaluated in the Ames/Salmonella test, where negative
results were obtained.
Several other- tests of commercial and technical samples have been reported
to be negative. In addition, The National Toxicology Program (NTP) has recently
sponsored mutagenicity testing (a modified Ames Salmonella test, a sex-linked
recessive lethal test in Drosophila, sister chromatid exchange formation and
chromosome aberrations in Chinese hamster ovary cells in vitro on a technical
sample of tetrachloroethylene; the preliminary results were negative. (These
studies are in the process of being peer-reviewed, and were not discussed in
this chapter, except for the Ames test.) The inconsistencies of available
results on different samples of tetrachloroethylene may be a function of the
toxicity of the test material, of exposure conditions used for testing this
volatile chemical, or of differences in sample contaminants and/or added stabil-
izers. Information on chemical composition of the tetrachloroethylene test
samples was scarce.
Although tetrachloroethylene itself has not been shown to be mutagenic,
it should be emphasized that the negative results are not wholly unequivocal.
Appropriate concurrent controls, adequate sample sizes, and exposure conditions
were sometimes not used, and in some cases the available data are not sufficient
to determine whether an adequate test was conducted. Also, there have been no
reliable studies investigating the ability of tetrachloroethylene to cause
chromosome nondisjunction, which would result in aneuploidy, a significant
genotoxic effect.
Because the epoxide of tetrachloroethylene was mutagenic in bacterial studies,
the concern should be raised that it may pose a mutagenic hazard. It should be
8-24
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noted, however, that the parent compound was assayed in the presence of several
types of metabolic activation systems (i.e., liver homogenates, intact hepatocytes,
and whole mammals) and the results were largely negative or weakly positive.
Therefore, it is uncertain whether these negative or weak findings were the
result of limitations of the activation systems, the epoxide not being produced
in sufficient quantities, or the epoxide possessing too short a half life to
cause a detectable mutagenic response.
In conclusion, inadequate information exists to war-rant a provisional
classification of tetrachloroethylene either as nonmutagem'c or- mutagenic. If
tetrachloroethylene is a mutagen, the evidence available thus far indicates
that it is only weakly so. (Because of insufficient information, this conclusion
is not made with regard to its potential for causing chromosome nondisjunction.)
Certain commercial and technical preparations of tetrachloroethylene may contain
mutagenic impurities and/or added stabilizers. Although there may be mutagenic
agents in certain preparations of tetrachloroethylene, usually large amounts of
material (at toxic levels) were required to elicit weak responses.
8-25
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8.8 REFERENCES
Bartsch, H., C. Malaveille, A. Barbin, and G. Planche. 1979. Mutagenic and
alkylating metabolites of halo-ethylenes, chlorobutadienes and dichlorobutenes
produced by rodent or human liver tissues. Arch. Toxicol. 41:249-277.
Bateman, A.J. and S.S. Epstein. 1971. Dominant lethal mutations in mammals. In:
Chemical mutagens: Principles and methods for their detection, Vol. 2
(A. Hollaender, ed.) Plenum Press, New York, pp. 541-568.
Beliles, R.P., D.J. Brusisk, and F.J. Mecler. 1980. Teratogenic mutagenic
risk of workplace contaminants: trichlorethylene, perchloroethylene, and
carbon disulfide. Contract No. 210-77-0047. Litton bionetics, Inc.,
Kensington, Maryland.
Bronzetti, G., C. Bauer, C. Corsi, R. Del Carratore, A. Galli, R. Nieri, and
M. Paolini. 1983. Genetic and biochemical studies on perchloroethylene
in vitro and in vivo. Mutat. Res. 116:323-331.
Callen, D.F., C.R. Wolf, R.M. Philpot. 1980. Cytochrome P450 mediated genetic
activity and cytotoxicity of seven halogenated aliphatic hydrocarbons in
Saccharomyces cerevisiae. Mutat. Res. 77:55-63.
Cerna, N., and H. Kypenova. 1977. Mutagenic activity of chloroethylenes
analyzed by scr-eening system test. Mutat. Res. 46:36 (Abst.).
Fabre, F. 1978. Induced intragenic recombination in yeast can occur- during
the GI mitotic phase. Nature 272:795-798.
Fabre, F., and H. Roman. 1977. Genetic evidence for- inducibility of recombination
competence in yeast. Proc. Natl. Acad. Sci. USA 74:1667-1671.
Greim, H., G. Bonse, Z. Radwan, D. Reichert, and D. Henschler. 1975. Mutagenicity
in vitro and potential carcinogenicity of chlorinated ethylenes as a function
oT metabolic oxirane formation. Biochem. Phar-macol. 24:2013-2017.
Gu, Z.W., B. Sele, P. Jalbert, M. Vincent, C. Marka, D. Charma, and J. Faure.
1981. Induction d'echanges entre les chromatides soeurs (SCE) par- le
trichloroethylene et ses metabolites. Toxicological European Research
3:63-67.
Henschler, D. 1977. Metabolism and mutagenicity of halogenated olefins: a
comparison of structure and activity. Environ. Health Per spec. 21:61-64.
Ikeda, M., A. Koizumi, T. Watanable, A. Endo, and K. Sato. 1980. Cytogenetic
and cytokinetic investigations on lymphocytes from workers occupationally
exposed to tetrachloroethylene. Toxicol. Letters. 5:251-256.
Kastenbaum, M.A. and K.O. Bowman. 1970. Tables for determining statistical
significance of mutation frequencies. Mutat. Res. 9:527-549.
8-26
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Kline, S.A., E.G. McCoy, H.S. Rosenkranz, and B.L. Van Duuren. 1982. Mutagenicity
of chloralkene epoxides in bacterial systems. Mutat. Res. 101:115-125.
Margard, W. 1978. In vitro bioassay of chlorinated hydrocarbon solvents.
Battelle Laboratories. Unpublished proprietary document for Detrex Chemical
Industries.
Mayer, V.W., and C.T. Coin. 1980. Induction of mitotic recombination by
certain hair-dye chemicals in Saccharomyces cerevisiae. Mutat. Res. 78:243-252.
McCann, J., E. Choi, E. Yamasaki, and B.N. Ames. 1975. Detection of carcinogens
as mutagens in the Salmonella/microsome test: Assay of 300 chemicals. Proc.
Natl. Acad. Sci. USA 72:5125-5139.
NTP. 1983. Unpublished data on tetrachloroethylene: on Salmonel1 a/mi crosome
preincubation test provided by Or. E. Zeiger.
Rampy, L.W., J.F. Quast, M.F. Balmer, B.K.J. Leong, and P.J. Gehring. 1978.
Results of a long-term inhalation toxicity study. Perchloroethylene in rats.
Toxicology Research Laboratory. Health and Environmental Research. The Dow
Chemical Company. Midland, Michigan. Unpublished.
Ris, H. 1949. The anaphase movement of chromosomes in the spermatocytes of the
grasshopper. Biol. Bull. 96:90-106.
Russell L.B., and B.E. Matter. 1980. Whole mammal mutagenicity tests. Evaluation
of five methods. Mutat. Res. 75:279-302.
Schumann, A.M., J.F. Quast, and P.G. Watanabe. 1980. The pharmacokinetics and
macromolecular interactions of perchloroethylene in mice and rats as related to
oncogenicity. Toxicol. Appl. Pharm. 55:207-219.
Shahin, M.M. 1975. Genetic activity of niridazole in yeast. Mutat. Res.
30:191-198.
SRI International. 1983. Salmonella test results on tetrachloroethylene.
Prepared for- U.S Environmental Protection Agency, Dr. Harry Milman, project
officer. Unpublished.
Stott, W.T. and P.G. Watanabe. 1982. Differentiation of genetic versus epigenetic
mechanisms of toxicity and its application to risk assessment. Drug Metabolism
Reviews 13:853-873.
Schlossberg, L. (Detrex Chemical Industries, Inc. Detroit, Michigan) January
5, 1981. Memorandum to Dr. V. Vaughan-Dellarco of the U.S. Environmental
Protection Agency, Reproductive Effects Assessment Group.
Waskell, L. 1978. A study of the mutagenicity of anesthetics and their
metabolites. Mutat. Res. 57:141-153.
8-27
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Williams, G.M. 1983. DMA repair tests of 11 chlorinated hydrocarbon analogs.
Prepared for ICAIR Life systems, Inc. TR-507-18 and U.S. Environmental
Protection Agency. Dr. Harry Mllman, project officer. Unpublished.
Williams, G.M., and T. Shimada. January 1983. Evaluation of several halogenated
ethane and ethylene compounds for genotoxicity. Final report for PPG
Industries, Inc., Pittsburgh, Pennsylvania. Unpublished.
8-28
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9.0 CARCINOGENICITY
The purpose of this section is to provide an evaluation of the likelihood
that tetrachloroethylene (perchloroethylene) is a human carcinogen and, on the
assumption that it is a human carcinogen, to provide a basis for estimating its
public health impact, including a potency evaluation, in relation to other
carcinogens. The evaluation of carcinogenic!ty depends heavily on animal
bioassays and epidemiologic evidence. However, information on mutagenicity and
metabolism, particularly in relation to interaction with DNA, as well as to
pharmacokinetic behavior, has an important bearing on both the qualitative and
quantitative assessment of carcinogenicity. The available information on these
subjects is reviewed in other sections of this document. This section presents
an evaluation of the animal bioassays, the human epidemiologic evidence, the
quantitative aspects of assessment, and finally, a summary and conclusions dealing
with all of the relevant aspects of the carcinogenicity of tetrachloroethylene.
9.1 ANIMAL STUDIES
Two long-term animal bioassays have been performed to assess the carcino-
genic potential of tetrachloroethylene. In one study involving exposure of rats
and mice to tetrachloroethylene by gavage, the National Cancer Institute (NCI)
(1977a) reported the induction of hepatocellular carcinomas in male and female
mice, but determined that the test with rats was inconclusive because of excessive
mortality. In the other study, in which Sprague-Dawley rats were exposed to
tetrachloroethylene by inhalation, the Dow Chemical Company (Rampy et al. 1978)
reported no evidence for- the carcinogenicity of the chemical. However, limita-
tions in this study make it difficult to assess the carcinogenic potential of
tetrachlor oethylene.
9.1.1 National Cancer Institute Bioassay (1977a)
The tetrachloroethylene sample used in this bioassay was purchased from
the Aldrich Chemical Company, Milwaukee, Wisconsin. Analysis by gas-liquid
chromatography and infrared spectroscopy yielded results indicating a purity
of > 99% with at least one minor impurity not identified in the report.
Identification of the impurities in the test sample was not made (personal
communications with the NCI and the Aldrich Chemical Company).
9-1
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The carcinogenicity of tetrachloroethylene was tested in Osborne-Mendel
rats and B6C3F1 mice. The initial age of the weanling animals was 25 days for
the mice and 35 days for the rats. Two treatment groups consisted of 50 males
and 50 females, and matched vehicle (corn oil) and untreated control groups
comprised 20 animals of each sex. Selected dosage levels were those determined
to be maximally tolerated in an 8-week subchronic study, i.e., a dosage that
was not fatal and/or did not reduce body weight gain more than approximately
20%, and one-half maximally tolerated in an 8-week subchronic toxicity test.
Time-weighted average doses (mg/kg/dose) in the chronic study were 941 and 471
for- male rats, 949 and 474 for female rats, 1,072 and 536 for male mice, and
772 and 386 for female mice. Tetrachloroethylene was administered to the animals
by gastric intubation in corn oil once each day, 5 days/week, for 78 weeks.
During the final 26 weeks of treatment, doses were administered to rats in a
cyclic pattern of 1 week without treatment followed by 4 weeks with treatment.
Body weights and food consumption were obtained weekly for the first 10 weeks
and monthly thereafter. Mice and rats were permitted to survive an additional
12 and 32 weeks after treatment, respectively, until sacrifice.
Each animal was submitted to extensive gross and microscopic examinations.
Specified organs, plus any other tissue containing visible lesions, were fixed
in 10% buffered formalin, embe'dded in paraplast, and sectioned for slides.
Hematoxylin and eosin staining (H and E) was used routinely, but other- stains
were employed when needed. Diagnoses of observed tumors and other lesions were
coded according to a modified Systematized Nomenclature of Pathology (SNOP)
originally developed by the College of American Pathologists in 1965.
Tetrachloroethylene was found to be carcinogenic in mice in this study.
Results summarized in Table 9-1 indicate that tetrachloroethylene induced highly
statistically significant (P < 0.001) increases in the incidence of hepatocellular
carcinomas in both sexes of mice in both treatment groups as compared to untreated
controls or vehicle-controls. The microscopic appearance of carcinomas was
variable, with some tumors composed of well-differentiated hepatocytes arranged
in rather uniform hepatic cords, and other lesions consisting of anaplastic cells,
often with inclusion bodies with vacuolated, pale cytoplasm. Mitotic figures
were often present. In male mice, the first hepatocellular carcinomas were
detected at 27 weeks in the low-dose group, 40 weeks in the high-dose group,
9-2
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TABLE 9-1. INCIDENCE OF HEPATOCELLULAR CARCINOMAS IN B6C3F1 MICE
FED TETRACHLOROETHYLENE
(National Cancer Institute 1977a)
Dose (nig/kg/day)3
Male
untreated
vehicle-control
536
1072
Femal e
untreated
vehicle-control
386
772
Hepatocellular carcinomas
2/17 (12%)
2/20 (10%)
32/49 (65%)
27/48 (56%)
2/20 (10%)
0/20 (0%)
19/48 (40%)
19/48 (40%)
P valuesb
P < 0.001
P < 0.001
P < 0.001
P < 0.001
«nme-weignted^ average doses.
^Probability level (P-values) for the Fisher Exact Test comparison of dose
groups with vehicle-control group.
and 90 and 91 weeks in vehicle-control and untreated control groups. In female
mice, the first hepatocellular carcinomas were observed at week 41 in the
low-dose group, week 50 in the high-dose group, and week 91 in the untreated
control group. Metastases of hepatocellular carcinomas occurred in the kidneys
of one untreated control male and in the lung of three low-dose males, one
low-dose female, and one high-dose female.
Toxic nephropathy in mice was apparent in 40/49 low-dose males, 45/48 high-
dose males, 46/48 low-dose females, and 48/48 high-dose females. Control animals
did not exhibit this lesion. Chronic murine pneumonia was also a frequently
observed finding. A low incidence of bloating or abdominal distension was
9-3
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noted 1n treated animals during the second year of the study. Body weight gain
was comparable between groups (Figure 9-1). Median survival times were greater
than 90 weeks in control males, 78 weeks in low-dose males, and 43 weeks in
high-dose males (Figure 9-2). Median survival times for females were greater
than 90 weeks in control females, 62 weeks in low-dose females, and 50 weeks in
high-dose females (Figure 9-2).
In rats, toxic nephropathy, not found in control animals, was detected in
43/49 low-dose males, 47/50 high-dose males, 29/50 low-dose females, and 39/50
high-dose females. Figure 9-3 indicates that treated rats gained less weight
than controls, though the difference was slight, with maximum reduction being
13% during the first year and 19% during the second year. Clinical signs
apparent in treated animals included a hunched appearance and urine stains on
the lower abdomen. Respiratory abnormalities, characterized by dyspnea, wheezing,
and/or reddish nasal discharge, were noted with increased incidence in all groups
during aging of the animals, and chronic murine pneumonia was diagnosed in >^ 62%
of the animals in each group. As indicated in Figure 9-4, median survival times
were greater than 88 weeks in the control groups, 68 weeks in low-dose females,
66 weeks in high-dose females, 67 weeks in low-dose males, and 44 weeks in
high-dose males. The survival was not adequate to support any conclusions
about the cardnogenicity of tetrachloroethylene in rats.
In an attempt to characterize impurities present in the tetrachloroethylene
product used in this study, documented chemical analyses of the test samples
performed at the carcinogenicity testing laboratory of the NCI bioassay program
were examined by the analytical department of the Diamond Shamrock Corporation
(interoffice memorandum from E. A. Rowe to G. K. Hatfield, Diamond Shamrock
Corporation, October 8, 1979, obtained with the documented chemical analyses
from G. K. Hatfield, January 28, 1981). As indicated in the interoffice memo-
randum, tetrachloroethylene samples used in the National Cancer Institute
bioassay (NCI 1977a) were not available for- analysis at Diamond Shamrock;
however, the analytical method, with the same type of instrument and column
used at the carcinogenicity testing laboratory, was reproduced. Evaluation of
the analytical method led to the conclusions presented in the previously mentioned
memorandum that the method could not distinguish epichlorohydrin from trichloro-
ethylene, and that the tetrachloroethylene product used in the NCI (1977a)
9-4
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-40
- 30
UNTREATED CONTROL
• VEHICLE CONTROL
• •••••••< LOW DOSE
— — HIGH DOSE
-20
- 10
45 60 75
TIME ON TEST (WEEKS)
90
105
120
3U —
_ 40-
cc
0
- 30-
I
O
UJ
§2°-
m
2
UJ
5 10-
0-
/x
-^t^"*™--«Sr^ — ^C^J^^'^^r
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• «•••••• LOW DOSE
FEMALE MICE HIGH DOSE
i 1 i 1 i 1 i 1 i I i 1 i 1 i
j la jG **5 6^ /o SO iuj "A
— su
-40
— 30
- 20
- 10
^
TIME ON TEST (WEEKS)
Figure 9-1. Growth curves for male and female mice in the tetrachloroethylene
chronic study. (National Cancer Institute 1977a).
9-5
-------
0.8-
_J
>
>
OC
V) 06 —
u.
O
>
h-
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5 0.4-
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t| *•• • • • •« • ^
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• •••••• LOW DOSE
— HIGH DOSE
1 1 ' 1 ' 1 ' 1 ' | ' 1 ' I
) 15 30 45 60 75 90 105 12
1.0
— 0.8
-
-06
- 0.4
-0.2
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TIME ON TEST (WEEKS)
0.8-
^
^
—
-------
750-T-
UNTREATED CONTROL
— VEHICLE CONTROL
• •••••••• LOW DOSE
— — — - HIGH DOSE
I
15
I
30
I
45
60
I
75
I
90
I
105
750
— 600
— 450
— 300
- 150
120
TIME ON TEST (WEEKS)
750
750
— 600
— 450
VEHICLE CONTROL
• ••«••• LOW DOSE
— — — HIGH DOSE
- 300
- 150
15
105
120
TIME ON TEST (WEEKS)
Figure 9-3. Growth curves for male and female rats in the tetrachloroethylene
chronic study. (National Cancer Institute 1977a).
9-7
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0.8-
cc
3
u.
O
03
<
m
O
QC
Q.
0.6-
0.2-
o.o
••• 1
• •••••
MALE RATS
UNTREATED CONTROL
VEHICLE CONTROL
LOW DOSE
— — HIGH DOSE
I 'I
15 30
45
1^
60
I
75
I
90
105
-1 0
L-0.8
-0.6
-0.4
— 0.2
•0.0
120
TIME ON TEST (WEEKS)
0.8 -J
_j
<
>
1
=i 0.4
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cc
0.
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0.0-
• ••••• * —1 j 1
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- •••» '^
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-
_
_ VEHICLE CONTROL
• •••••• LOW DOSE
— — — HIGH DOSE
FEMALE RAlb
1
•1 ' 1
^\ \
• ••• "^
"1 '• ^
L ••*_..
~~ "^" ~~ t_ **i*
h__^
M «^» M^ •
i n
_
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— 06
-
— 04
—
- 0.2
n n
15 30 45 60 75
TIME ON TEST (WEEKS)
90
105
120
Figure 9-4. Survival comparisons of male and female rats in the tetrachloroethylene
chronic study. (National Cancer Institute 1977a).
9-E
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bloassay could have contained one or both of these compounds as impurities.
The documented chemical analyses show contaminant levels of 0.055%, 0.041%,
and 0.010% in tetrachloroethylene samples analyzed at the beginning of the
bloassay, at 1 year into the bloassay, and at 2 years into the bloassay,
respectively. The conclusion stated in the memorandum 1s that the contaminant
was probably eplchlorohydrln, since: 1) epichlorohydrln was a commonly used
stabilizer for tetrachloroethylene at that time; 2) a reported analysis by a
competitor of Diamond Shamrock showed a maximum of 0.015% trlchloroethylene in
any tetrachloroethylene product manufactured 1n the United States, and 3) the
decreased amount of impurity found in the tetrachloroethylene samples as the
bloassay progressed suggests some decomposition 1n that, according to the
experience of Diamond Shamrock, eplchlorohydrln but not trlchloroethylene
levels would decrease during storage. Nonetheless, although the evidence
provided by Diamond Shamrock indicates that eplchlorohydrln was present 1n the
tetrachloroethylene product used in the NCI (1977a) bloassay, the quantity of
epichlorohydrln in the test material used in this study remains uncertain.
Furthermore, the different levels of unknown material shown 1n the documented
chemical analysis may be due to the use of the indicated different lots, possibly
containing unequal amounts of impurities.
An estimate of the likelihood that the eplchlorohydrln impurity in the
tetrachloroethylene material used in the NCI experiment could have been large
enough to account for the positive results can be obtained by considering the
results of Laskin et al. (1980), who exposed rats to epichlorohydrln via inhalation.
They found that 1/100 rats exposed to 30 ppm (0.115 mg/1 air, equivalent to
13.2 mg/kg/day) of epichlorohydrln for 6 hours/day, 5 days/week, for 730 days
developed squamous cell nasal carcinomas, compared with 0/50 in control animals.
The eplchlorohydrln impurity 1n the NCI tetrachloroethylene high-dose male
mice experiment was approximately 1,032 mg/kg/day x 0.041% = 0.42 mg/kg/day, from
the discussion above. This is only about 0.42/13.2 ~ 0.03 times the dose that
gave an incidence of only 1/100 in the Laskin et al. rat experiment. Therefore,
it 1s unlikely that epichlorohydrln impurities at the level estimated to be present
1n the NCI experiment could have contributed appreciably to the positive response.
In a second cardnogeniclty bloassay sponsored by the National Toxicology
Program (NTP 1983, draft), tetrachloroethylene was given orally to B6C3F1 mice
9-9
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and to four strains of rats (Sherman, Fischer 344, Long-Evans, and Wistar). The
tetrachloroethylene sample used in this ongoing carcinogenicity bioassay did
not contain detectable amounts of epoxide contaminants. In the study with
female B6C3F1 mice (NTP 1983, draft), groups of 100 females received 25, 50,
100, or 200 mg of more than 99% pure tetrachloroethylene per kg body weight in
corn oil by gavage for 103 weeks, 5 days/week. Vehicle-control and untreated
control groups of 100 mice each were used. Survival was not affected significantly.
Doses of 50 mg/kg or- more produced a dose-related cytomegaly of the kidneys.
Other signs of liver toxicity appeared during the study as increases of sorbitol
dehydrogenase activity, and increases in relative liver weight and lipid levels
were associated with increasing dose levels. Doses of 50 mg/kg or more produced
time- and dose-related increases in the incidence of hepatocellular adenomas
and carcinomas bearing no causal relationship to the observed renal damage.
The first adenoma appeared at week 46; the first carcinoma at week 58. The
NTP intends to conduct an audit of the raw data of this study before the final
technical report is prepared. The findings of the audit will determine the
validity of the study. If the audit reveals no major problems, the NTP expects
to present the report to a peer review group no sooner then early 1984 (letter
from Dr. Schetz, NTP, to Dr. Haberman, CAG, September 29, 1983).
9.1.2 Dow Chemical Company Inhalation Study (Rampy et a!. 1978)
Two groups of weanling Sprague-Dawley (Spartan substrain) rats, each
composed of 96 males and 96 females, were exposed to 600 ppm (4.07 mg/1 air)
or 300 ppm (2.03 mg/1 air) of tetrachloroethylene 6 hours/day, 5 days/week, for
52 weeks. An untreated group of 192 males and 192 females served as controls.
Controls were not put in inhalation chambers, but were in the treatment room
during exposure. At the end of the treatment period of 52 weeks, animals
were allowed to survive until sacrifice at 31 months. The composition of the
test material (Lot A12282D) by gas chromatographic analysis was as follows:
trichloroethylene, 3 ppm (liquid volume %); hexachloroethane, < 12 ppm; carbon
tetrachloride, 2 ppm; 4-methyl morpholine, 44 ppm; nonvolatile residue, 2 ppm;
tetrachloroethylene, balance. Exposure was done in 3.7 m3 inhalation chambers
with a dynamic airflow system. Analyses of tetrachloroethylene levels in the
inhalation chambers during the treatment period revealed analytical concentrations
(mean +_ standard deviation) of 310 +_ 32 ppm (244 analyses) and 592 +_ 62 ppm
9-10
-------
(1,245 analyses). Analytical concentrations by infrared analysis within jf 10%
of nominal levels were achieved on 89.3% (low-dose) and 97.1% (high-dose) of
the exposure days. Animals were evaluated for clinical toxic signs, body
weight changes, urinalysis at 24 months, and hematology at 12 and 24 months.
Survivors and decedents were given gross and histopathologic examinations.
Bone marrow samples were taken from three males and three females in each group
sacrificed at 1 year for cytogenetic evaluation.
Clinical signs of toxicity were not observed with the nominal concentrations
of tetrachloroethylene used in this study. Mean body weight gains were similar
among groups. Hematology and urine analyses showed no treatment-related effects
of tetrachloroethylene. The mortality patterns exhibited in the study are des-
cribed in Table 9-2. Mortality in high-dose males was slightly greater than that
of controls during months 5 to 24; the earlier- onset of chronic renal disease in
this treatment group was considered to be a contributing factor in increased
mortality.
No carcinogenic effects of tetrachloroethylene were observed from pathologic
examination of the animals. Statistical analysis of the data showed numerous
non-neoplastic abnormalities that occurred spontaneously and were within the
normal variation encountered in lifetime studies with this strain of rat. With
respect to tumor findings, statistical analysis of the data did not reveal a
definite increased tumor incidence in animals exposed to tetrachloroethylene.
Tumors or tumor-like changes in the kidney were found in 1/189 control, 2/94
low-dose, and 4/94 high-dose males during gross necropsy; however, light micro-
scopic examination of kidney lesions did not show a statistically significant
tumor- incidence compared to controls. Although many tumor- types were found in
treated and control animals, there was no statistically significant (P > 0.05)
increase in tumor incidence at any anatomical site.
The results of this study do not indicate a definite carcinogenic effect
of tetrachloroethylene in Sprague-Dawley rats; i.e., the tumor incidence between
control and treated rats was similar. However, this study has the following draw-
backs: 1) the period of exposure was only 12 months rather than the lifetime of
the animals, which would have been a more appropriate duration for- carcinogenicity
studies; and 2) the dose levels in this study do not appear to have been high
enough to provide maximum sensitivity.
9-11
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TABLE 9-2. CUMULATIVE SURVIVAL OF SPRAGUE-DAWLEY RATS EXPOSED TO
TETRACHLOROETHYLENE FOR 12 MONTHS
(adapted from Rampy et al. 1978)
Month of Study 0 ppm
Male Female
300 ppm
Male Female
600 ppm
Male Female
Initial
6
12
18
24
31
189
187
183
155
44
1
189
188
185
151
70
12
94
94
91
74
26
1
91
91
91
77
37
6
94
88a
84*
55a
13*
1
94
94
91
863
493
5
dP < 0.005 by the FisnerTxact Test.
9.1.3 Intraperitoneal Administration Study (Theiss et al. 1977)
Thelss et al. (1977) tested tetrachloroethylene for cardnogenicity in the
strain A mouse pulmonary tumor induction system. The test sample, a product of
the Aldrich Chemical Company, was reagent grade with a purity exceeding 95 to 99%.
Strain A/St male mice, 6 to 8 weeks old, were used in this assay. The maximum
tolerated dosage, defined as the dosage which five mice tolerated after- six
intraperitoneal injections over a 2-week period followed by a 4-week observation
period, was determined and used in the bioassay. In the main test, 20 mice per
treatment group received three intraperitoneal injections of 80, 200, or 400 mg/kg
of tetrachloroethylene weekly until total dosages of 1120, 4800, and 9600 mg/kg,
respectively, were achieved. Survivors were sacrificed at 24 weeks after the
first injection, and the number of surface adenomas was counted. Results were
compared with findings in vehicle (tricaprylin) and untreated controls by the
Student t test. Tetrachloroethylene did not statistically increase (P > 0.05)
the incidence of pulmonary tumors in this study (Table 9-3). This strain was
sensitive to the positive control chemical urethan, as shown in Table 9-3.
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TABLE 9-3. PULMONARY TUMOR RESPONSE TO TETRACHLOROETHYLENE
(adapted from Theiss et al. 1977)
No. survivors/ No. lung
Compound Dosage (mg/kg) No. animals tumors/mouse
Tricaprylln
Tetrachl oroethyl ene
—
80
200
400
46/50
15/20
17/20
18/20
0.39 +_ 0.063
0.27 + 0.07
0.41 + 0.10
0.50 T 0.12
Urethan 1,000 mg/kg 20/20 19.6^2.4
{1 injection)
aMean +_ S.E.
A negative result in this assay is not considered conclusive, since several
chemicals known to be carcinogenic in chronic rodent bioassays induce no response
in this assay.
The strain A mouse pulmonary tumor assay is relatively insensitive to mouse
carcinogens for which the effect is confined to the liver (Theiss et al. 1977).
For example, chloroform, 2-chloroethyl ether, and hexachlorocyclohexane induce
tumors of the liver (not other sites) in mice (NCI 1976, 1977c; Innes et al.
1969) but were not carcinogenic in the assay by Theiss et al. (1977). The
reasons for the negative lung response are not understood, but it may be due to a
smaller concentration of activating enzymes in the lung than in the liver.
9.1.4 Skin Painting Study (Van Duuren et al. 1979)
A carcinogenicity study of purified tetrachloroethylene in ICR/Ha Swiss
mice was described by Van Duuren et al. (1979). Maximum tolerated dosages were
determined in range-finding studies 6 to 8 weeks in duration, and were
selected as dosages that did not affect body weight gain or- produce clinical
signs of toxicity. This study included the following experiments: 1) 30
females were treated topically on the dorsal skin with a single application of
163 mg tetrachloroethylene followed 14 days later by the applications of 5.0 ug
phorbol myristate acetate (PMA) to the skin three times weekly until termination
of the study at 428 to 576 days; median survival time was 428 to 576 days. 2)
30 females were given thrice weekly topical applications of 54 mg tetrachloro-
ethyl ene for the duration of the test (440 to 594 days) with a median survival
time of 317 to 589 days. A vehicle (acetone) control group of 30 mice and an
untreated control group of 100 mice were included in these experiments.
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In the initiation-promotion experiment, 210 mice treated with PMA alone were
also on test. The mice were 6 to 8 weeks old at the beginning of the study, and
were housed six to a cage. Test sites on the skin were shaved as necessary and
were not covered; however, it was the authors' impression that tetrachloro-
ethylene was immediately absorbed and that evaporation from test sites was
minimal (personal communication, B. L. Van Duuren, New York University). The
animals were weighed monthly, and each animal was examined by necropsy. Tumors
and lesions as well as skin, liver, stomach, and kidneys were examined histo-
logically.
Tetrachloroethylene did not show initiating activity in the initiation-
promotion experiment; the number of mice with skin papillomas (squamous cell
carcinomas) was: 4 (0) initiated with tetrachloroethylene, 15 (3) treated
with PMA alone, and 0 (0) in the control groups. The study involving repeated
application to the skin produced lung and stomach tumors in 16 and 0 tetrachloro-
ethylene-treated mice, respectively; 11 and 2 vehicle-treated controls,
respectively; and 30 and 5 untreated controls, respectively.
The negative results for tetrachloroethylene in mouse skin as an initiator
and as a complete carcinogen can be reconciled with the positive mouse liver-
response in the NCI study if it is hypothesized that the skin does not have the
necessary enzymes to convert tetrachloroethylene to an active metabolite,
whereas the liver does have this capability.
The lack of sensitivity of skin application tests, as compared with tests
using other routes of exposure, is apparent from the results of Van Duuren et
al. (1979) with 1-chloroprene, cis-l,3-dich1oropropene, and 2-chloroproponal.
They found that none of the three compounds induced a response as initiator's in
initiation-promotion experiments or with repeated topical application on the
skin. However, they did observe a statistically significant increase in the
incidence of forestomach tumors in female HatlCR Swiss mice dosed by gavage
with 1-chloroprene (P < 0.0005) and 2-chloroproponal (P < 0.05) and in the
Incidence of local sarcomas In female Ha:ICR Swiss mice treated with cis-1,3-
dlchloropropene (P < 0.0005) by subcutaneous injection.
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9.2 EPIDEMIOLOGIC STUDIES
There are six epidemiologic studies that relate to tetrachloroethylene
exposure. Only one of these studies, however, has actually identified workers
exposed to tetrachloroethylene. Because tetrachloroethylene has been used in the
dry-cleaning industry, the present discussion also includes three proportionate
mortality studies of decedents who had worked in the dry-cleaning industry, as
well as two case-control studies in which the cases and controls were asked
about their employment histories, including employment in the dry-cleaning
industry.
9.2.1 Kaplan (1980)
Kaplan (1980) did a retrospective cohort mortality study of dry-cleaning
worker-s exposed to tetrachloroethylene for at least one year prior to 1960.
The study was performed under contract to the Biometry Section of the National
Institute for Occupational Safety and Health (NIOSH) Industry-Wide Studies Branch.
In a preface to the discussion of the study, Kaplan reported that levels
of tetrachloroethylene exposure were "much higher" for- cleaners (machine opera-
tors) than for- other employees of dry-cleaning establishments. A geometric
mean time-weighted average for tetrachloroethylene was 22 ppm for the machine
operators. For- all other jobs, the highest corresponding value was reported
to be 3.3 ppm. These data were provided through a NIOSH industrial hygiene
survey of dry-cleaning facilities.
The study cohort, selected from records maintained by several labor unions,
consisted of 1,597 individuals employed more than 1 year- prior to 1960 in dry-
cleaning establishments. The primary solvent was tetrachloroethylene. Efforts
were made by the author to exclude all persons with previous occupational exposure
to carbon tetrachloride or- trichloroethylene. By September- 30, 1977, the end of
the study period, 1,058 individuals were found to be alive, 285 were deceased,
and 254 were of unknown vital status. The extent of follow-up varied by sex;
8% of the males and 20.4$ of the females remained lost to follow-up. Race was
known only for deceased workers, and was obtained from death certificate data.
Because of the lack of information regarding race, observed deaths by cause
were compared to expected deaths by means of a standardized mortality ratio
(SMR) for whites, an SMR for blacks, and a composite point estimate SMR for both.
Based on the assumption that every member of the cohort was white, expected
9-15
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deaths for whites were derived by multiplying the person-years accumulated for
the entire cohort by the white death rates within 5-year age groups (separately
for males and females and then with the two combined). Expected deaths for
blacks were similarly generated by assuming that all cohort members were black.
Composite expected deaths were calculated by weighting the total accumulated
person-years for the cohort in each 5-year age group by the proportion of
person-years attributable to the deceased blacks and the proportion attributable
to deceased whites, and then multiplying each total separately by the corresponding
death rates for whites and blacks, and finally adding across age groups to get
the "composite" expected deaths.
Because death certificates could not be located for all of the deceased, it
was assumed that those deaths for which no death certificates could be found had
the same distribution by cause as those for which death certificates were
available. Thus the SMRs for each cause of death were corrected to reflect
the missing death certificates. Using the SMR for- deaths from malignant
neoplasms of the colon in whites as an example, this correction was made in the
follow.ing manner:
(11/247 x 38) + 11 x 11 x 100 = 182
TT ~
where: 247 is the number of deaths in the cohort with death certificates;
38 is the number- of deaths in the cohort without death certificates;
11 is the observed number of colon cancer deaths identified by death
certificates;
6.98 is the expected number- of white colon cancer deaths;
100 is a constant used in calculating SMRs (by convention, SMRs are
expressed as a factor of 100); and
182 is the corrected white colon cancer SMR.
No significance tests of any of the SMRs were done by the author. However, the
author called attention to the elevated SMR from malignant neoplasms of the colon
as possibly related to occupational exposure (6.98 white expected deaths, 6.77
black expected deaths, and composite SMR = 182). Using an observed number- of
colon cancer cases corrected for the loss of death certificates, the Carcinogen
Assessment Group (CAG) found that the SMR for either whites or- blacks would be
statistically signficant (P < 0.05).
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The author points out that because the expected numbers of deaths were cal-
culated using U.S. rates, which include a higher socioeconomic class than tKe
dry-cleaners in this study, and because higher socioeconomic class is associated
with a risk of colon cancer, the risk of colon cancer from exposure to tetra-
chloroethylene found in this study is probably underestimated. Furthermore,
the expected number of colon cancer cases in this study was calculated using
mortality rates for neoplasms of the intestine, except rectum. Although most
of the deaths expected using mortality rates for neoplasms of the intestine
would be deaths from malignant neoplasms of the colon, some deaths would be
from neoplasms of the small intestine. Since all of the observed deaths were
from malignant neoplasms of the colon, the comparison of the observed to expected
colon cancer deaths in this study would also tend to underestimate the colon
cancer risk from exposure to tetrachloroethylene.
It should also be noted that although it could be argued that the number
of observed colon cancer deaths in each of the four- union locals in this study
was small, an elevated colon cancer SMR did exist in each of the four locals.
Finally, it should be noted that the colon cancer- SMR appeared to demonstrate a
positive correlation with the length of the follow-up period. This last finding
must be viewed with a great deal of caution, however, because of the author's
difficulty in defining length of exposure and hence length of follow-up.
In addition to colon cancer, the SMRs for cancer of various other sites
were also elevated. These included rectum, pancreas, respiratory system,
urinary organs, and "other and unspecified sites (major)." None of these were
significant at the P < 0.05 level when tested using an observed number- that was
corrected for lost death certificates; however, the SMRs for- cancer of three of
these sites [respiratory system, urinary organs, and "other and unspecified
sites (major-)"] could be considered borderline significant (0.10 < P < 0.05).
Perhaps the major weakness of this study with regard to evaluating
tetrachloroethylene as a carcinogen is that the history of solvent exposure
prior to 1960 was unknown for nearly half of the union member- shops. Because
the majority of dry-cleaning establishments in the United States used petroleum
distillates as the primary cleaning agent prior to 1960, it is quite likely
that most of the shops in this study used petroleum distillates as the cleaning
solvent prior- to changing to tetrachloroethylene.
9-17
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Other Important confounding variables were also not controlled. For example,
smoking is a major confounding variable to be considered when evaluating a
potential risk for respiratory or bladder cancer, both of which were found in
excess in this study. Socioeconomic status, as has been discussed, is a con-
founding variable for colon cancer.
Another weakness of this study is that 16 percent of the study cohort
was lost to follow-up. Currently, NIOSH is attempting to improve the percentage
of follow-up as well as to add to the length of follow-up. In addition, NIOSH
has identified other individuals who were exposed to tetrachloroethylene for
at least one year prior to 1960, so that the size of the cohort has also been
increased.
In summary, this study appears suggestive that dry-cleaning workers exposed
to tetrachloroethylene are at an elevated risk of colon cancer mortality.
Potential exposure to petroleum distillates for- approximately half of the cohort,
however, limits any conclusions with regard to the carcinogenicity of tetra-
chloroethylene in humans.
9.2.2 Blair et al. (1979)
Blair et al. (1979) reported, as the preliminary results of a cohort study
of 10,000 laundry and dry-cleaning workers, a proportionate mortality analysis
of 330 of the workers who had died during the period 1957-1977. Deaths were
identified from the mortality records of two union locals in St. Louis, Missouri.
The distribution by cause of death among the 330 was compared to that expected
based on the proportionate mortality experience of the United States. Only
279 of the 330 had worked exclusively in dry-cleaning establishments; however,
the authors did not report the proportionate mortality analyses of these 279.
Furthermore, the dry-cleaning agent used by the dry-cleaners in this study is
unknown. Among the 330 deaths, deaths from cancer of the lung, cervix, and
skin contributed to the finding of a significant (P < 0.05) excess proportion
of deaths from cancer at all sites. For lung cancer, there were 17 deaths
observed versus 10 expected (P < 0.05); for cervical cancer, 10 observed deaths
versus 4.8 expected (P < 0.05); and for skin cancer, 3 observed deaths versus
0.7 expected (P < 0.05). On the other hand, a significant deficit of deaths
occurred in the cause category identified as "all circulating diseases," with
100 observed versus 125.9 expected (P < 0.005)—a finding which could account
for the excess proportion of cancers observed.
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Only limited conclusions can be drawn from this study as to the potential
carcinogenicity of tetrachloroethylene. This is true for several reasons. First,
the study did not report the distribution of deaths for dry-cleaners alone.
Second, the dry-cleaning agent these workers used is not known. Lastly, certain
possibly confounding variables were not considered in the analysis of the results.
These variables include smoking (with regard to the lung cancer excess), socio-
economic status (with regard to the cervical cancer excess), and sunlight
exposure (with regard to the skin cancer excess).
9.2.3 Katz and Jowett (1981)
Katz and Jowett (1981) analyzed the death certificate records of 671 white
female laundry and dry-cleaning workers who had died during the period 1963-
1977. The records of dry-cleaning workers were not studied separately from
those of laundry workers, since both groups of workers shared the same occupa-
tional code. Furthermore, it is not known whether those decedents who were
dry-cleaners used tetrachloroethylene as a dry-cleaning agent. During the
1940s, petroleum derivatives were the predominant dry-cleaning agents used in
the United States. A shift to the use of tetrachloroethylene began in the
late 1940s and gained momentum in the 1950s and 1960s. By 1977, approximately
75% of the dry-cleaning plants in the United States were using tetrachloro-
ethylene. However, in the period before 1960, petroleum distillates were
still the dominant solvents in use (Kaplan 1980).
In this study, cause-specific proportionate mortality for the 671 deceased
laundry and dry-cleaning workers was compared to that for the deaths of all
other working females in Wisconsin during the same period, and to deaths of
females in "lower-wage occupations" in Wisconsin during the same period.
Significantly elevated proportionate mortality ratios (PMRs) were found for
deaths from cancer of the genitals (unspecified) (P < 0.01) and for deaths from
cancer of the kidney (P < 0.05) when deaths among women of all occupations and
deaths among women of "lower-wage occupations" were used as the comparison
groups. In summary, this study, although suggesting an association of employment
in the dry-cleaning industry with excess risks of certain types of cancer, cannot
be said to demonstrate conclusively that such an association exists, because of
the possible confounding influence of other dry-cleaning agents, and the fact
that no distinction was made between laundry and dry-cleaning workers.
9-19
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9.2.4 L1n and Kessler (1981)
Lin and Kessler (1981) did a case-control study of 109 pancreatic cancer
cases diagnosed during the period 1972-1975 from 115 hospitals in five metro-
politan areas. The control group was composed of subjects who were free from
cancer but who were similar to the study group in age (+_ 3 years), sex, race,
and marital status, and who were selected at random from among contemporaneous
admissions to the same hospital. Cases and controls were asked about demographic
characteristics, residential history, toxic exposures, animal contacts,
smoking habits, diet, medical history, medications, and family history. Males
were asked about sexual practices and urogenital conditions. Females were
questioned on these topics and also on their- marital, obstetric, and gynecologic
histories. Among other statistically significant associations, an association
was found between pancreatic cancer- and employment either as a dry-cleaner or
in a job involving exposure to gasoline. It is not known, however, how many
individuals with pancreatic cancer were employed as dry-cleaners and how many
were employed in occupations involving exposure to gasoline. Furthermore, the
dry-cleaners may have used a variety of dry-cleaning agents other than tetra-
chloroethylene. Thus this study, although suggestive of an association between
employment as a dry-cleaner and an excess risk of pancreatic cancer, cannot be
said to demonstrate an association between pancreatic cancer and tetrachloro-
ethylene exposure.
9.2.5 Asa! (personal communication 1983)
In two epidemiologic studies conducted by Dr. Nabih Asal of the University
of Oklahoma, an excess of cancer among dry-cleaners was found (personal communi-
cation 1983). These studies are as yet unpublished. One of the studies was
a kidney cancer- case-control study in which Asal found an association between
kidney cancer and the dry-cleaning occupation. The second study was a propor-
tionate mortality study of dry-cleaning workers in Oklahoma. Asal reported
that results from the latter study suggest an increased mortality from kidney
and lung cancer among the workers. As in the Blair et al., Katz and Jowett,
and Lin and Kessler studies previously discussed, exposure in the studies by
Asal was not specific for- tetrachloroethylene. Moreover, the conclusions of
these studies cannot objectively be evaluated until published descriptions of
them are available.
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9.3 QUANTITATIVE ESTIMATION
This quantitative section deals with estimation of the unit risk for
tetrachloroethylene as a potential carcinogen in air and water, and with the
potency of tetrachloroethylene relative to other carcinogens that have been
evaluated by the CAG. The unit risk for an air or water pollutant is defined
as the lifetime cancer risk to humans from daily exposure to a concentration of
1 ug/m3 of the pollutant in .air by inhalation, or to a concentration of 1 ug/1
in water by ingestion.
The unit risk estimate for tetrachloroethylene represents an extrapolation
below the dose range of experimental data. There is currently no solid scientific
basis for any mathematical extrapolation model that relates exposure to cancer
risk at the extremely low concentrations, including the unit concentration
given above, that must be dealt with in evaluating environmental hazards. For
practical reasons the correspondingly low levels of risk cannot be measured
directly either by animal experiments or- by epidemiologic study. Low-dose
extrapolation must, therefore, be based on current understanding of the mechanisms
of carcinogenesis. At the present time the dominant view of the carcinogenic
process involves the concept that most cancer-causing agents also cause
irreversible damage to DMA. This position is based in part on the fact that a
very large proportion of agents that cause cancer are also mutagenic. There
is reason to expect that the quantal response that is characteristic of
mutagenesis is associated with a linear non-threshold dose-response relationship.
Indeed, there is substantial evidence from mutagenicity studies with both
ionizing radiation and a wide variety of chemicals that this type of dose-
response model is the appropriate one to use. This is particularly true at
the lower end of the dose-response curve; at high doses, there can be an upward
curvature, probably reflecting the effects of multistage processes on the
mutagenic response. The linear- non-threshold dose-response relationship is
also consistent with the relatively few epidemiologic studies of cancer responses
to specific agents that contain enough information to make the evaluation
possible (e.g., radiation induced leukemia, breast and thyroid cancer-, skin
cancer induced by arsenic in drinking water, liver cancer induced by aflatoxins
in the diet). Some supporting evidence also exists from animal experiments
(e.g., the initiation stage of the two-stage carcinogenesis model in rat liver
and mouse skin).
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Because its scientific basis, although limited, is the best of any of the
current mathematical extrapolation models, the non-threshold model, which is
linear at low doses, has been adopted as the primary basis for risk extrapolation
to low levels of the dose-response relationship. The risk estimates made with
such a model should be regarded as conservative, representing the most plausible
upper limit for the risk; i.e., the true risk is not likely to be higher than
the estimate, but it could be lower.
For several reasons, the unit risk estimate based on animal bioassays is
only an approximate indication of the absolute risk in populations exposed to
known carcinogen concentrations. First, there are important species differences
in uptake, metabolism, and organ distribution of carcinogens, as well as species
differences in target site susceptibility, immunological responses, hormone
function, dietary factors, and disease. Second, the concept of equivalent
doses for humans compared to animals on a mg/surface area basis is virtually
without experimental verification as regards carcinogenic response. Finally,
human populations are variable with respect to genetic constitution and diet,
living environment, activity patterns, and other cultural factors.
The unit risk estimate can give a rough indication of the relative potency
of a given agent as compared with other carcinogens. Such estimates are, of
course, more reliable when the comparisons are based on studies in which the
test species, strain, sex, and routes of exposure are similar.
The quantitative aspect of carcinogen risk assessment is addressed here
because of its possible value in the regulatory decision-making process, e.g.,
in setting regulatory priorities, evaluating the adequacy of technology-based
controls, etc. However, the imprecision of presently available technology for
estimating cancer risks to humans at low levels of exposure should be recognized.
At best, the linear extrapolation model used here provides a rough but plausible
estimate of the upper limit of risk—that is, with this model it is not likely
that the true risk would be much more than the estimated risk, but it could be
considerably lower. The risk estimates presented in subsequent sections should
not be regarded, therefore, as accurate representations of the true cancer
risks even when the exposures involved are accurately defined. The estimates
presented may, however, be factored into regulatory decisions to the extent
that the concept of upper-risk limits is found to be useful.
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9.3.1 Procedures for the Determination of Unit Risk
9.3.1.1 Low-Dose Extrapolation Model--The mathematical formulation chosen to
describe the linear nonthreshold dose-response relationship at low doses is the
linearized multistage model. This model employs enough arbitrary constants to
be able to fit almost any monotonlcally increasing dose-response data, and it
incorporates a procedure for estimating the largest possible linear- slope (in
the 95% confidence limit sense) at low extrapolated doses that is consistent
with the data at all dose levels of the experiment.
Let P(d) represent the lifetime risk (probability) of cancer at dose d.
The multistage model has the form
P(d) = 1 - exp [-q0 + qj^d + qjd2 + ...+ qkdk)]
where
q, _> 0, 1 = 0, 1, 2, ..., k
Equivalently,
Pt(d) = 1 - exp [qjd + q2d2 + ... + qkdk)]
where
P (d) = P(d) - P(0)
t 1 - P(0)
is the extra risk over background rate at dose d.
The point estimate of the coefficients qi, 1 = 0, 1, 2, ..., k, and
consequently, the extra risk function, Pt(d), at any given dose d, is
calculated by maximizing the likelihood function of the data.
The point estimate and the 95% upper confidence limit of the extra risk,
P^(d), are calculated by using the computer program, GLOBAL79, developed by
Crump and Watson (1979). At low doses, upper 95% confidence limits on the
extra risk and lower 95% confidence limits on the dose producing a given risk
are determined from a 95% upper confidence limit, q* on parameter qj. When-
ever qi > 0, at low doses the extra risk Pt(d) has approximately the form
Pt(d) = q* x d. Therefore, q* x d is a 95% upper confidence limit on the
extra risk and R/q* is a 95% lower confidence limit on the dose, producing an
extra risk of R. Let LQ be the maximum value of the log-likelihood function.
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The upper-limit q* is calculated by increasing q^ to a value q* such that when the
log-likelihood is remaximized subject to this fixed value q* for the linear coef-
ficient, the resulting maximum value of the log-likelihood LI satisfies the equation
2 (L0 - l\) = 2.70554
where 2.70554 is the cumulative 90% point of the chi -square distribution with
one degree of freedom, which corresponds to a 95% upper-limit (one-sided). This
approach of computing the upper confidence limit for the extra risk P^(d) is an
improvement on the Crump et al . (1977) model. The upper confidence limit for
the extra risk calculated at low doses is alway linear. This is conceptually
consistent with the linear nonthreshold concept discussed earlier. The slope,
q*, is taken as an upper-bound of the potency of the chemical in inducing cancer
at low doses. (In the section calculating the risk estimates, Pt(d) will be
abbreviated as P.)
In fitting the dose-response model, the number- of terms in the polynomial
is chosen equal to (h-1), where h is the number- of dose groups in the experiment,
including the control group.
Whenever the multistage model does not fit the data sufficiently well, data
at the highest dose is deleted and the model is refit to the rest of the data.
This is continued until an acceptable fit to the data is obtained. To
determine whether or not a fit is acceptable, the chi-square statistic
2
X =
N,P,
is calculated where N^ is the number of animals in the i"1 dose group, X^ is
the number of animals in the ith dose group with a tumor- response, Pn- is the
probability of a response in the itn dose group estimated by fitting the
multistage model to the data, and h is the number of remaining groups. The
fit is determined to be unacceptable whenever X2 is larger than the cumulative
99% point of the chi-square distribution with f degrees of freedom, where f
equals the number of dose groups minus the number of non-zero multistage
coefficients.
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9.3.1.2 Selection of Data—For- some chemicals, several studies in different
animal species, strains, and sexes, each run at several doses and different
routes of exposure, are available. A choice must be made as to which of the
data, sets from several studies to use in the model. It may also be appropriate
to correct for metabolism differences between species and for absorption factors
via different routes of administration. The procedures used in evaluating
these data are consistent with the approach of making a maximum-likely risk
estimate. They are as follows:
1. The tumor incidence data are separated according to organ sites or
tumor types. The set of data (i.e., dose and tumor incidence) used in the
model is the set where the incidence is statistically significantly higher
than the control for- at least one test dose level and/or where the tumor
incidence rate shows a statistically significant trend with respect to dose
level. The data set that gives the highest estimate of the lifetime car-
cinogenic risk, q*, is selected in most cases. However-, efforts are made to
exclude data sets that produce spuriously high risk estimates because of a
small number of animals. That is, if two sets of data show a similar dose-
response relationship, and one has a very small sample size, the set of data
having the larger sample size is selected for calculating the carcinogenic
potency.
2. If there are two or more data sets of comparable size that are
identical with respect to species, strain, sex, and tumor- sites, the geometric
mean of q*, estimated from each of these data sets, is used for risk assessment.
The geometric mean of numbers A]_, A2» ••., Am is defined as
(Aj x A2 x ... x A,,,)1/"1.
3. If two or more significant tumor sites are observed in the same study,
and if the data are available, the number of animals with at least one of the
specific tumor sites under consideration is used as incidence data in the model.
9.3.1.3 Calculation of Human Equivalent Dosages—Following the suggestion of
Mantel and Schneiderman (1975), it is assumed that mg/surface area/day is an
equivalent dose between species. Since, to a close approximation, the surface
area is proportional to the two-thirds power of the weight, as would be the case
for a perfect sphere, the exposure in mg/day per two-thirds power of the weight
is also considered to be equivalent exposure. In an animal experiment, this
equivalent dose is computed in the following manner.
9-25
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Let
Le = duration of experiment
le = duration of exposure
m = average dose per day in mg during administration of the agent (i.e.,
during le), and
W = average weight of the experimental animal
Then, the lifetime exposure is
l_e x W2/3
9.3.1.3.1 Oral. Often exposures are not given in units of mg/day, and it
becomes necessary to convert the given exposures into mg/day. Similarly, in
drinking water studies, exposure is expressed as ppm in the water. For example,
1n most feeding studies exposure is given in terms of ppm in the diet. In
these cases, the exposure in mg/day is
m = ppm x F x r
where ppm is parts per million of the carcinogenic agent in the diet or water,
F is the weight of the food or water- consumed per- day in kg, and r is the
absorption fraction. In the absence of any data to the contrary, r is assumed
to be equal to one. For a uniform diet, the weight of the food consumed is
proportional to the calories required, which in turn is proportional to the
surface area, or- two-thirds power- of the weight. Water demands are also
assumed to be proportional to the surface area, so that
m « ppm x W x r
or
jn a ppm.
rW2/3
As a result, ppm in the diet or water is often assumed to be an equivalent
exposure between species. However, this is not justified for the present
study, since the ratio of calories to food weight is very different in the
diet of man as compared to laboratory animals, primarily due to differences
in the moisture content of the foods eaten. For the same reason, the amount
9-26
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of drinking water requited by each species also differs. It is therefore
necessary to use an empirically-derived factor, f = F/W, which is the
fraction of an organism's body weight that is consumed per day as food,
expressed as follows:
Fraction of body
weight consumed as
Species W ffood fwater
Man
Rats
Mice
70
0.35
0.03
0.028
0.05
0.13
0.029
0.078
0.17
Thus, when the exposure is given as a certain dietary or water concentration in
ppm, the exposure 1n mg/W^/3 1s
rc = PP") * F = ppm x f x W = ppm x f x W1/3
rW2/3
When exposure is given in terms of mg/kg/day = m/Wr = s, the conversion is
simply
= s x
Wl/3.
9.3.1.3.2 Inhalation. When exposure is via inhalation, the calculation of dose
can be considered for two cases where 1) the carcinogenic agent is either a
completely water-soluble gas or an aerosol and is absorbed proportionally to the
amount of air breathed in, and 2) where the carcinogen is a poorly water-soluble
gas which reaches an equilibrium between the air breathed and the body compart-
ments. After equilibrium is reached, the rate of absorption of these agents
is expected to be proportional to the metabolic rate, which in turn is propor-
tional to the rate of oxygen consumption, which in turn is a function of surface
area.
9.3.1.3.2.1 Case 1. Agents that are 1n the form of particulate matter or
virtually completely absorbed gases, such as sulfur dioxide, can reasonably be
9-27
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expected to be absorbed proportionally to the breathing rate. In this case
the exposure in mg/day may be expressed as
m = I x v x r
where I = inhalation rate per day in m3, v = mg/m3 of the agent in air, and
r- = the absorption fraction.
The inhalation rates, I, for- various species can be calculated from the
observations of the Federation of American Societies for Experimental Biology
(FASEB 1974) that 25 g mice breathe 34.5 liters/day and 113 g rats breathe 105
liters/day. For- mice and rats of other weights, W (in kilograms), the surface
area proportionality can be used to find breathing rates in m3/day as follows:
For mice, I = 0.0345 (W/0.025)2/3 m3/day
For rats, I = 0.105 (W/0.113)2/3 m3/day
For humans, the value of 30 m3/day* is adopted as a standard breathing rate
(International Commission on Radiological Protection 1977). The equivalent
exposure in mg/W2/3 for these agents can be derived from the air intake data
in a way analogous to the food intake data. The empirical factors for the air-
intake per kg per- day, i = I/W, based upon the previously stated relationships,
are tabulated as follows:
Species VI i = I/W
Man 70 0.29
Rats 0.35 0.64
Mice 0.03 1.3
Therefore, for- particulates or- completely absorbed gases, the equivalent
exposure in mg/W2/3 is
m = Ivr = iWvr = iv/l/3vr
"
In the absence of experimental information or a sound theoretical argument
to the contrary, the fraction absorbed, r, is assumed to be the same for all
species.
9.3.1.3.2.2 Case 2. The dose in mg/day of partially soluble vapors is pro-
portional to the 02 consumption, which in turn is proportional to W2'3 and is
also proportional to the solubility of the gas in body fluids, which can be
*From "Recommendation of the International Commission on Radiological
Protection," page 9. The average breathing rate is 107 cm3 per 8-hour workday
and 2 x 107 cm3 in 24 hours.
9-28
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expressed as an absorption coefficient, r, for the gas. Therefore, expressing
the 62 consumption as 02 = k W^/3, where k is a constant independent of species,
it follows that
m = k W2/3 x v x r
or
d = m = kw
W2/3
As with Case 1, in the absence of experimental information or a sound theoretical
argument to the contrary, the absorption fraction, r, is assumed to be the same
for all species. Therefore, for these substances a certain concentration in
ppm or ug/m3 in experimental animals is equivalent to the same concentration
in humans. This is supported by the observation that the minimum alveolar-
concentration necessary to produce a given "stage" of anesthesia is similar
in man and animals (Dripps et al. 1977). When the animals are exposed via
the oral route and human exposure is via inhalation or- vice versa, the
assumption is made, unless there is pharmacokinetic evidence to the contrary,
that absorption is equal by either- exposure route.
9.3.1.4 Calculation of the Unit Risk from Animal Studies—The risk associated
with d mg/kg2/3/day is obtained from GLOBAL79 and, for- most cases of interest to
risk assessment, can be adequately approximated by P(d) = 1 - exp (-q*d). A
"unit risk" in units X is simply the risk corresponding to an exposure of X = 1.
This value is estimated simply by finding the number of mg/kg2/3/day that
corresponds to one unit of X, and substituting this value into the above
relationship. Thus, for example, if X is in units of ug/m3 in the air, then
for case 1, d = 0.29 x 701/3 x lO'3 mg/kg2/3/day, and for case 2, d = 1,
when ug/m3 is the unit used to compute parameters in animal experiments.
If exposures are given in terms of ppm in air-, the following calculation
may be used:
1 ppm = 1 2 x molecular weight (gas) mg/m3
molecular weight (air)
Note that an equivalent method of calculating unit risk would be to use mg/kg
for the animal exposures, and then to increase the jtn polynomial coefficient
by an amount
(Wh/Wa)J/3 j = 1, 2, ..., k,
and to use mg/kg equivalents for the unit risk values.
9-29
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9.3.1.4.1 Adjustments for Less Than Llfespan Duration of Experiment. If the
duration of experiment Le is less than the natural lifespan of the test animal
L, the slope q*, or more generally the exponent g(d), is increased by multiplying
a factor (L/Le) . We assume that if the average dose d is continued, the age-
specific rate of cancer will continue to increase as a constant function of the
background rate. The age-specific rates for humans increase at least by the
third power of the age and often by a considerably higher power, as demonstrated
by Doll (1971). Thus, it is expected that the cumulative tumor rate would increase
by at least the third power of age. Using this fact, it is assumed that the
slope q*, or more generally the exponent g(d), would also increase by at least
the third power of age. As a result, if the slope q* [or g(d)] is calculated at
age Le, it is expected that if the experiment had been continued for the full
lifespan L at the given average exposure, the slope q* [or g(d)] would have been
increased by at least (L/Le) .
This adjustment is conceptually consistent with the proportional hazard
model proposed by Cox (1972) and the time-to-tumor model considered by Daffer et
al. (1980), where the probability of cancer by age t and at dose d is given by
P(d,t) = 1 - exp [-f(t) x g(d)].
9.3.1.5 Model for Estimation of Unit Risk Based on Human Data—If human epidemic-
logic studies and sufficiently valid exposure information are available for the
compound, they are always used in some way. If they show a carcinogenic effect,
the data are analyzed to give an estimate of the linear dependence of cancer
rates on lifetime average dose, which is equivalent to the factor B^. If they
show no carcinogenic effect when positive animal evidence is available, then it
is assumed that a risk does exist, but it is smaller than could have been observed
in the epidemiologic study, and an upper-limit to the cancer incidence is calculated
assuming hypothetically that the true incidence is below the level of detection
in the cohort studied, which is determined largely by the cohort size. Whenever
possible, human data are used in preference to animal bioassay data.
Very little information exists that can be utilized to extrapolate from high
exposure occupational studies to exposures at low environmental levels. However,
if a number of simplifying assumptions are made, it is possible to construct a
crude dose-response model whose parameters can be estimated using vital statistics,
epidemiologic studies, and estimates of worker exposures.
9-30
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In human studies, the response is measured in terms of the relative risk of
the exposed cohort of individuals as compared with the control group. The
mathematical model employed for the low-dose extrapolation assumes that for
low exposures the lifetime probability of death from cancer, PQ, may be
represented by the linear equation
P0 = A + BHx
where A is the lifetime probability in the absence of the agent, and x is the
average lifetime exposure to environmental levels in units such as ppm. The
factor BH is the increased probability of cancer associated with each unit
increase of x, the agent in air.
If it is assumed that R, the relative risk of cancer for exposed workers
as compared to the general population, is independent of length of exposure or
age at exposure, and depends only upon average lifetime exposure, it follows that
R = P = A + BH (*i + *2)
FO A + BH xf
or
RP0 = A + BH (X! + x2)
where xj = lifetime average daily exposure to the agent for the general population,
X2 = lifetime average daily exposure to the agent in the occupational setting,
and PQ = lifetime probability of dying of cancer- with no or- negligible exposure.
Substituting PQ = A + BH xj and rearranging gives
BH = PQ (R - D/X2
To use this model, estimates of R and x2 must be obtained from epidemiologic
studies. The value PQ is derived by means of the life table methodology from
the age- and cause-specific death rates for- the general population found in
U.S. vital statistics tables.
9.3.2 Unit Risk Estimates
9.3.2.1 Data Available for Potency Calculation—The data on hepatocellular
carcinomas in B6C3F1 female mice, as reported in the NCI gavage study (1977a),
constitute the only information available for use in estimating the carcinogenic
potency of tetrachloroethylene. Because of the high mortality and early tumor-
occurrence in the animals treated with tetrachloroethylene in the above study,
the data on times-to-death cited therein have been used in the present report
9-31
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for potency calculations. These data, along with the results of the calculations,
are presented in Appendix B. For comparison, the study's incidence of hepato-
cellular carcinomas (Table 9-4) is also utilized, employing four different
extrapolation models.
TABLE 9-4. INCIDENCE RATE OF HEPATOCELLULAR CARCINOMAS IN FEMALE B6C3F1 MICE
(NCI 1977a)
Experimental
dose (mg/kg/day)
0
386
772
Human equivalent
dose (mg/kg/day)a
0
18.02
36.04
Tumor
incidenceb
0/20
19/48 (40%)
19/45 (42%)
aThe human equivalent dose is calculated by d x (5/7) x (78/90) x (0.03/
70)1/3 = 4.67 x 10~2 x d, where d is the experimental dose given to animals 5
days per week for 78 weeks. The lifespan of mice is assumed to be 90 weeks.
The factor (0.03/70)1/3 is the cubic root of the ratio of body weights between
animals and humans. The use of this factor assumes that the doses in mg per
body surface are equally effective between mice and humans. For comparison,
the potency is also calculated when mg/kg/day is assumed to be the equally
effective dosage between mice and humans.
^Denominators are the numbers of animals that survived beyond 40 weeks. The
time to the first tumor is 41 weeks.
There 1s no sound human epidemlologic data presently available that can be
used to calculate the cancer risk of tetrachloroethylene. However, existing
human data can be used for- arriving at rough estimates of the magnitude of
human cancer risk from exposure to tetrachloroethylene, and for comparison
with potency estimates made on the basis of animal data. The colon cancer
mortality data presented in the NIOSH report (Kaplan 1980) have been used for
this purpose. A detailed discussion of that study, focusing particularly on
9-32
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its investigation of colon cancer, was presented above in the epidemiology
section of this chapter.
9.3.2.2 Assumption About Equally Potent Doses—To convert animal doses to
human doses, it is assumed that mg/surface area/day is equivalent among species.
This assumption is suggested by Mantel and Schneiderman (1975) and is supported
empirically in comparing the toxicity of anticancer agents in various animal
species and man (Freireich et al. 1966, Rail 1977). Under this assumption,
the animal dose d, in mg/kg/day, is multiplied by a factor (Wg/W^)1'^ to obtain
the human equivalent dose, where Wa and Wn are, respectively, the animal and
human body weights. Another approach to interspecies extrapolation that has
been used in risk assessment is direct extrapolation on the basis of body
weight (i.e., mg/kg/day). In the present report, the risks calculated on the
basis of body weight are also provided. The carcinogenic potency per- mg/kg/day
calculated on the basis of body surface area is always higher than when
calculated on the basis of body weight. However, the fact that the calculated
potency expressed in a particular unit, (mg/kg/day)"^, is higher- on the
basis of surface area does not necessarily imply that humans are assumed to be
more sensitive than animals. This would be the case if doses in mg/kg/day
were in reality equally potent between humans and animals.
9.3.2.3 Choice of Low-Dose Extrapolation Models--In addition to the multistage
model currently used by the CAG for low-dose extrapolation, three mote models,
referred to as the probit, the Weibull, and the one-hit models, are employed
for- purposes of comparison (Appendix A). These models cover almost the entire
spectrum of risk estimates that could be generated from existing mathematical
extrapolation models. Generally statistical in character, these models are not
derived from biological arguments, except for the multistage model, which has
been used to support the somatic mutation hypothesis of carcinogenesis (Armitage
and Doll 1954, Whittemore 1978, Whittemore and Keller 1978). The main difference
among these models is the rate at which the response function P(d) approaches
zero or P(0) as dose d decreases. For instance, the probit model would usually
predict a smaller risk at low doses than the multistage model because of the
difference of the deer-easing rate in the low-dose region. However, it should
be noted that one could always artificially give the multistage model the same
(or even greater) rate of decrease as the probit model by making some dose
9-33
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transformation and/or by assuming that some of the parameters in the multistage
model are zero. This, of course, is not reasonable without knowing, a priori,
what the carcinogenic process for the agent is. Although the multistage model
appears to be the most reasonable or at least the most general model to use,
the point estimate generated from this model is of limited value because it
does not help to determine the shape of the dose-response curve beyond experi-
mental exposure levels. Furthermore, point estimates at low doses extrapolated
beyond experimental doses could be extremely unstable and could differ-
drastically, depending on the amount of the lowest experimental dose. Since
upper-bound estimates from the multistage model at low doses are relatively
more stable than point estimates, it is suggested that the upper-bound estimate
for the risk (or the lower-bound estimate for- the dose) be used in evaluating
the carcinogenic potency of a suspect carcinogen. The upper-bound estimate
can be taken as a plausible estimate if the true dose-response curve is actually
linear at low doses. The upper-bound estimate means that the risks are not
likely to be higher, but could be lower if the compound has a concave upward
dose-response curve or- a threshold at low doses. Another reason one can, at
best, obtain an upper-bound estimate of the risk when animal data are used is
that the estimated risk is a probability conditional to the assumption that an
animal carcinogen is also a human carcinogen. Therefore, in reality, the
actual risk could range from a value near zero to an upper-bound estimate.
9.3.2.4 Calculation of the Carcinogenic Potency of Tetrachloroethylene
9.3.2.4.1 Calculation on the Basis of Gavage Data. As indicated previously,
data on hepatocellular carcinomas in female mice are analyzed by using either
the data of time-to-event or- the data of incidence rate.
The time-to-event data and the calculations of potency are presented in
Appendix B. In this calculation, it is assumed that all animals that died and
were found to have hepatocellular carcinomas died from these tumors. Strictly
speaking, one knows only that the tumor occurred before the death of an animal,
and that a tumor was observed. However, the carcinogenic potency estimates
based on other- assumptions do not differ greatly when the multistage model with
a time factor tk is used for low-dose extrapolation. This conclusion is generally
true for life-threatening tumors, as observed by Krewski et al. (1983) using
simulated data.
9-34
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The data on hepatocellular carcinoma incidence rates are used to fit four
different extrapolation models. The maximum likelihood estimates of the parameters
in each of the four models are presented in Table A-l in Appendix A. Table 9-5
presents the upper-bound and point estimates of the risk at 0.01, 0.05, 0.1,
0.5, and 1 mg/kg/day.
Both the probit and the Weibull models estimate much higher risk at these
dose levels than the multistage and the one-hit models. The potency q* = 3.5 x
10~2 mg/kg/day, calculated by the multistage model with a time factor, will be
used for calculating the unit risk of tetrachloroethylene from drinking water.
This potency is obtained when the dose in mg/surface area/day is assumed to be
equally effective among species. If this assumption is made, then q* = 2.6 x
10-3 mg/kg/day should be used for- this calculation.
9.3.2.4.2 Calculation on the Basis of Human Epidemiologic Data. No sound
epidemiologic data are available that could be used to estimate the carcinogenic
potency of tetrachloroethylene. However-, one study by Kaplan (1980) for- NIOSH
may be used to provide an estimation of the possible range of its potency.
This is the retrospective cohort mortality study on dry-cleaning workers exposed
to tetrachloroethylene. The author of this report concluded: "Based on the
study results, cancer of the colon, which accounted for 11 observed deaths,
appears to be the cause of death most likely to have resulted from occupational
exposure in our cohort." For this reason, data on colon cancer- have been
selected by the CAG for the pur-pose of calculating the carcinogenic potency of
tetrachloroethylene. In the present study, the standardized mortality ratio
(SMR) for colon cancers is estimated to be 182, which is statistically
significant at P < 0.05. There are no data on the concentration of tetra-
chloroethylene to which the cohort workers were exposed. Based on an industrial
hygiene study by NIOSH (Kaplan 1980), Campbell et al. (1980) estimated that
the average annual tetrachloroethylene exposure in dry-cleaning plants that
used the chemical ranged approximately from 1 ppm for "coin-op" workers to 7
ppm for machine operators. The annual average exposure was calculated under
the assumption that a worker- worked 40 hours a week for 50 weeks in a year.
As an approximation, 0.8 ppm and 6 ppm will be used as lower limits and
upper limits of lifetime exposures by the cohort workers. If exposure were
0.8 ppm, the carcinogenic potency would be:
9-35
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TABLE 9-5. UPPER-BOUND (POINT) ESTIMATES OF RISK AT VARIOUS DOSE LEVELS, BASED ON THE DATA OF
HEPATOCELLULAR CARCINOMAS IN FEMALE MICE (NCI 1977a)
Assumption of
human equivalent
dose3
Dose with
surface
corrections
Model s
Multistage with
time factor
Multistage
Problt
Wei bull
One-hit
0.01
3.5 x lO'4
(1.8 x 10-4)
2.6 x lO'4
(2.0 x 10-3)
1
(1.6 x 10-1)
1
(1.6 x 10-M
2.6 x lO'4
(2.0 x 10-3)
Risk at Dose Levels (mg/kg/day)
0.05 0.1 0.5
1.7 x 10-3
(8.8 x lO'4)
1.3 x 10-3
(1.0 x 10-3)
1
(2.0 x 10-1)
1
(2.0 x 10-1)
1.3 x 10-3
(1.0 x 10-3)
3.5 x 10-3
(1.8 x 10-3)
2.6 x 10-3
(2.0 x 10-3)
1
(2.2 x 10-1)
1
(2.1 x 10-M
2.6 x 10-3
(2.0 x 10-3)
1.7 x 10-2
(8.8 x 10-3)
1.3 x ID'2
(1.0 x 10-2)
1
(2.7 x 10-1)
1
(2.5 x 10-1)
1.3 x 10-2
(1.0 x 10-2)
1
3.5 x ID'2
(1.8 x ID'2)
2.6 x 10-2
(2.0 x 10-2)
1
(2.9 x 10-1)
(2.7 x 10-1)
2.6 x ID'2
(2.0 x 10-2)
Dose without
surface
correction
Multistage with
time factor-
Multistage
Problt
Welbull
One-hit
2.6 x 10-5
(1.3 x 10-5)
2.0 x 10-5
(1.5 x lO-5)
1
(1.1 x 10-1)
1
(1.4 x 10-1)
2.0 x 10-5
(1.5 x lO-5)
1.3 x 10-4
(6.6 x lO-5)
1.0 x lO'4
(7.5 x 10-5)
1
(1.4 x 10-1)
1
(1.6 x 10-1)
1.0 x lO'4
(7.5 x 10-5)
2.6 x 10-4
(1.3 x 10-4)
2.0 x 10-4
(1.5 x 10-3)
1
(1.5 x 10-1)
1
(1.8 x 10-1)
2.0 x 10-4
(1.5 x lO'4)
1.3 x 10-3
(6.6 x lO'4)
1.0 x 10-3
(7.5 x lO'4)
1
(1.9 x 10-1)
(2.1 x 10-1)
1.0 x ID'3
(7.5 x 10-4)
2.6 x 10-3
(1.3 x 10-3)
2.0 x lO-3
(1.5 x ID'3)
(2.1 x 10-1)
1
(2.3 x 10-1)
2.0 x 10-3
(1.5 x 10-3)
*Doses with surface correction assume that doses 1n mg/surface area/day are equally potent in producing cancer 1n
humans and animals. Doses without surface correction assume that doses 1n mg/kg/day are equivalent between
humans and animals.
-------
B = (1.82 - 1) x 1.6 x IP"2 = 1>6 x 10-2/ppm.
*• 0.8
If exposure were 6 ppm, the carcinogenic potency would be:
B = (1.82 - 1) x 1.6 x IP"2 = 2.2 x 10-3/ppm.
£ 6
These estimates will be used to evaluate the reasonableness of the tetrachloro-
ethylene potency estimations calculated on the basis of animal data.
9.3.2.5 Pharmacokinetic Data Relevant to Quantitative Risk Assessment—AIthough
a large amount of pharmacokinetic data is available on experimental animals, the
uses of these data for risk extrapolation are very limited because of the lack
of corresponding human data. Based on the information in Table 9-7, taken from
Pegg et al. (1979) and Schumann et al. (1980), it is suggested that the rate of
tetrachloroethylene metabolism for- mice is much higher- than that for rats when
animals are exposed via inhalation. However, the comparative metabolic rates
of mice and rats do not appear to be different when the animals are exposed via
gavage. The percentage of tetrachloroethylene that was expired unchanged from
rats is consistent with that which was observed in humans when tetrachloro-
ethylene was given by the inhalation route. No Information is available on the
human absorption rate of tetrachloroethylene via the oral route. For the oral
route, if it is assumed that humans and rats are similar- with respect to absorption
rate, then the absorption rate for mice would be also similar to that of humans,
because, on the basis of available information, mice and rats have a similar-
rate of absorption by this route of exposure.
TABLE 9-6. EXPIRED TETRACHLOROETHYLENE UNCHANGED AS PERCENTAGE OF
RECOVERED RADIOACTIVITY
(Pegg et al. 1979, Schumann et al. 1980)
Oral Inhalation
1 mg/kg 500 mg/kg 10 ppm 600 ppm
Rats
Mice
72%
NA
90%
83%
68%
12%
88%
NA
NA =OJot avail able.
9-37
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9.3.2.6 Risk Associated With 1 ug/liter of Tetrachloroethylene In Drinking Water—
As discussed above, it Is reasonable to assume that both humans and mice have
similar absorption rates for tetrachloroethylene. Therefore, the carcinogenic
—potency q* = 3.5 x 10~2/(mg/kg/day), estimated from the mice gavage data, is used
herein to calculate the unit risk of tetrachloroethylene in drinking water. Under
the assumption that dally water consumption for a 70-kg person is 2 liters, the
risk from drinking water containing 1 ug/Hter of tetrachloroethylene is estimated
as follows: P = 3.5 x 10-2/(mg/kg/day) x 10-3 (mg/ug) x 2 (liter/day) v 70 kg
= 1 x ID'6.
9.3.2.7 Risk Associated With 1 ug/m3 of Tetrachloroethylene in Air--The
carcinogenic potency q* = 3.5 x 10~2/(mg/kg/day) estimated from the data in the
NCI mice gavage study (1977a) is used herein to calculate the carcinogenic
risk from tetrachloroethylene by inhalation. The relative absorption rate of
tetrachlorothylene by inhalation and by gavage is not known. As an approximation,
it is assumed that the effective dose by inhalation in mice 1s one-sixth of
that by gavage. This assumption is based on the pharmacokinetic data presented
in Table 9-7 and on the observation that absorption rates for tetrachloroethylene
by the oral route are similar- between mice and rats, since humans are assumed to
be similar to rats with respect to absorption rate. To calculate the potency of
tetrachloroethylene 1n terms of ug/m3, it is assumed that the daily air intake
for a 70-kg person is 20 m3. Thus, for 1 ug/m3 of tetrachloroethylene in air,
the corresponding dose in mg/kg/day is
(1/6) x (1 ug/m3) x (20 m3/day) x (1/70 kg) x (10'3 mg/ug) = 4.76 x 10-5 mg/kg/day.
The carcinogenic potency of tetrachloroethylene by inhalation is calculated as
follows:
q* = (3.5 x 10-2) x (4.75 x 10-5) = 1.7 x 10-6/{ug/m3).
The risk associated with 1 ug/m3 of tetrachloroethylene in air is thus estimated
as 1.7 x 10-6. since 1 ppm of tetrachloroethylene is equivalent to 6,908
ug/m3, the potency q* in terms of ppm is estimated as
q* = 1.7 x 10-6 x 6,908 = 1.0 x 10-2/ppm.
9-38
-------
This estimate does not appear to be inconsistent with the crude estimates that
have been calculated on the basis of human data. These range from 2.2 x 10-3/ppm
to 1.6 x 10'2/ppm.
9.3.3 Comparison of Potency With Other Compounds
One of the uses of quantitative potency estimates is to compare the
relative potencies of carcinogens. Figure 9-5 is a histogram representing the
frequency distribution of potency indices for 53 suspect carcinogens evaluated
by the CAG. The actual data summarized by the histogram are presented in Table
9-7. The potency index is derived from q* the 95% upper bound of the linear
component in the multistage model, and is expressed in terms of (mMol/kg/day)'1.
Where no human data were available, animal oral studies were used in preference
to animal inhalation studies, since oral studies have constituted the majority
of animal studies.
Based on available data concerning hepatocellular carcinomas in female
mice, the potency index for tetrachloroethylene has been calculated as 6 x 10^.
-This figure is derived by multiplying the slope q* = 3.5 x 10~2/(mg/kg/day) and
the molecular weight of tetrachloroethylene, 165.8. This places the potency
index for tetrachloroethylene in the fourth quartile of the 53 suspect
carcinogens evaluated by the CAG.
The ranking of relative potency indices is subject to the uncertainties
involved in comparing a number- of potency estimates for- different chemicals
based on varying routes of exposure in different species, by means of data from
studies whose quality varies widely. All of the indices presented are based on
estimates of low-dose risk, using linear- extrapolation from the observational
range. These indices may not be appropriate for the comparison of potencies if
linearity does not exist at the low-dose range, or if comparison is to be made
at the high-dose range. If the latter is the case, then an index other than
the one calculated above may be more appropriate.
9.4 SUMMARY AND CONCLUSIONS
9.4.1 Qualitative
Tetrachloroethylene induced a statistically significant increase in the
incidence of hepatocellular carcinomas in both high- and low-dose male and female
B6C3F1 mice when administered by gavage for a period of 78 weeks. The tetra-
chloroethylene used was over 99% pure, but was estimated to contain epichlorohydrin
9-39
-------
4th
quartile
•4-
3rd
quartile
-»-
2nd
quartile
-»•
1st
quartile
1 x 10*' 4x 10+i 2x10
NO
co
-2
i
0
246
Log of Potency Index
r
8
Figure 9-5. Histogram representing the frequency distribution of
the potency indices of 53 suspect carcinogens evaluated
by the Carcinogen Assessment Group.
9-40
-------
TABLE 9-7. RELATIVE CARCINOGENIC POTENCIES AMONG 53 CHEMICALS EVALUATED BY
THE CARCINOGEN ASSESSMENT GROUP AS SUSPECT HUMAN CARCINOGENS1.2.3
Slope
Compounds (mg/kg/day)"l
Acrylonitrile
Aflatoxin B^
Aldrin
Allyl Chloride
Arsenic
B[a]P
Benzene
Benzidene
Beryllium
Cadmium
Carbon tetrachloride
Chlordane
Chlorinated ethanes
1,2-dichl oroethane
hexachl oroethane
1,1, 2 ,2- tetrachl oroethane
1 , 1 , 1-trichl oroethane
1 , 1 ,2-trichl oroethane
Chloroform
Chromium
DDT
Dichlorobenzidine
1,1-dichloroethylene
Dieldrin
0.24(W)
2924
11.4
1.19x10-2
15(H)
11.5
5.2xlO-2(W)
234(W)
4.86
6.65(W)
1.30x10-1
1.61
6.9x10-2
1.42x10-2
0.20
1.6x10-3
5.73x10-2
7x10-2
41
8.42
1.69
1.47x10-1(1)
30.4
Molecular-
weight
53.1
312.3
369.4
76.5
149.8
252.3
78
184.2
9
112.4
153.8
409.8
98.9
236.7
167.9
133.4
133.4
119.4
104
354.5
253.1
97
380.9
Potency
i ndex
1x10+1
9xlO+5
4x10+3
9x10-1
2x10+3
3x10+3
4x10°
4xlO+4
4x10+1
7x10+2
2x10+1
7x10+2
7x10°
3x10°
3x10+1
2x10-1
8x10°
8x10°
4x10+3
3x10+3
4x10+2
1x10+1
lxlO+4
Order of
magnitude
(Iog10
index)
+1
+6
+4
0
+3
+3
+1
+5
+2
+3
+1
+3
+1
0
+1
-1
+1
+1
+4
+3
+3
+1
+4
9-41
-------
TABLE 9-7. (continued)
Slope Molecular-
Compounds (nig/kg/day)'! weight
Dinitrotoluene
Diphenylhydrazine
Epichlorohydrin
B1s(2-chloroethyl )ether
B1s(chloromethyl ) ether
Ethyl ene di bromide (EDB)
Ethyl ene oxide
Heptachlor
Hexachlorobenzene
Hexachlorobutadiene
Hexachl or ocycl ohexane
technical grade
alpha Isomer
beta Isomer
gamma Isomer
Methyl ene chloride
Nickel
N1trosam1nes
Dimethyl ni tr osami ne
D1 ethyl nitr osami ne
Di butyl nitrosamine
N-ni tr osopyr rol 1 di ne
N-ni troso-N-ethyl urea
N-ni troso-N-methyl urea
N-ni troso-di phenyl ami ne
0.31
0.77
9.9xlO-3
1.14
9300(1)
8.51
0.63(1)
3.37
1.67
7.75x10-2
4.75
11.12
1.84
1.33
6.3xlO-4
1.15(W)
25.9(not by q*)
43.5(not by q*)
5.43
2.13
32.9
302.6
4.92xlO-3
182
180
92.5
143
115
187.9
44.0
373.3
284.4
261
290.9
290.9
290.9
290.9
84.9
58.7
74.1
102.1
158.2
100.2
117.1
103.1
198
Potency
index
6xlO+1
1x10+2
9X10'1
2xlO+2
lxlO+6
2x10+3
3X10+1
1x10+3
5x10+2
2xlO+1
1x10+3
3x10+3
5x10+2
4x10+2
5x10-2
7xlO+1
2x10+3
4xlO+3
9x10+2
2x10+2
4x10+3
3xlO+4
1x10°
Order of
magnitude
nog10
index)
+2
+2
0
+2
+6
+3
+1
+3
+3
+1
+3
+3
+3
+3
-1
+2
+3
+4
+3
+2
+4
+4
0
PCBs
4.34
324
1x10+3
+3
9-42
-------
TABLE 9-7. (continued)
Compounds
Phenols
2 ,4 ,6-tHchl orophenol
Tetrachlorodloxin
Tetr achl oroethyl ene
Toxaphene
Trichloroethylene
Vinyl chloride
Slope
(mg/kg/day)'1
1.99xlO-2
4.25x10$
3.5xlO-2
1.13
1.9x10-2
1.75x10-2(1)
Order- of
magnitude
Molecular Potency (logjo
weight index index)
197.4 4x10° +1
322 lxlO+8 +8
165.8 6x10° +1
414 5xlO+2 +3
131.4 2.5x100 0
62.5 IxlOO 0
Remarks:
1. Animal slopes are
95% upper-limit si
lopes based on the linearized multi-
2.
3.
stage model. They are calculated based on animal oral studies, except
for those indicated by I (animal inhalation), W (human occupational
exposure), and H (human drinking water- exposure). Human slopes are
point estimates based on the linear non-threshold model.
The potency index is a rounded-off slope in (mMol/kg/day)'1 and is cal-
culated by multiplying the slopes 1n (mg/kg/day)'1 by molecular weight
of the compound.
Not all of the carcinogenic potencies presented in this table represent
the same degree of certainty. All are subject to change as new evidence
becomes available.
9-43
-------
concentrations of less than 500 ppm. It is unlikely that this response could be
attributed to the low concentration of epichlorohydrin.
No carcinogenic effect was observed in lifetime studies of Osborne-Mendel
rats given tetrachloroethylene by gavage or in Sprague-Dawley rats exposed via
inhalation for 12 months followed by 12 months of observation. However, because
of excessive dose-related mortality in the gavage experiment, and because of
the low dose level in the inhalation study, no conclusions can be made about
the carcinogenicity of tetrachloroethylene in rats. A gavage study in four rat
strains, now in progress at the National Toxicology Program (NTP), should
clarify the nature of the rat response.
Intraperitoneal injection of tetrachloroethylene in strain A mice induced
no statistically significant incidence of pulmonary adenomas. In mouse skin
initiation experiments, tetrachloroethylene did not initiate skin tumors, nor
did it induce skin tumors when applied alone three times per week for the
lifetime of the animals. However, because of inherent limitations in these
assays, the negative results they showed do not detract from the positive
findings of the National Cancer Institute (NCI) mouse experiment.
In a cohort study of dry-cleaning workers exposed to tetrachloroethylene,
1t was found that these workers were at an elevated risk of colon cancer
mortality; however, it was also the case that as many as one half of these
workers may have been exposed to petroleum distillates earlier in their working
history. Other studies either completed or currently in progress of dry-cleaners
or of cancer cases for which employment in the dry-cleaning occupation was
found to be a risk factor- did not attempt to Identify workers by tetrachloro-
ethylene exposure.
9.4.2 Quantitative
Data on hepatocellular carcinomas in female mice from the NCI (1977a)
gavage study on tetrachloroethylene have been used herein to calculate the risk
associated with drinking water contaminated with tetrachloroethylene. Although
the available pharmacokinetic data suggest that humans absorb less tetrachloro-
ethylene than mice via inhalation exposure, the same conclusion cannot be made
concerning exposure via the oral route.
Under the assumption that absorption rates for- humans and mice are similar,
the upper-bound cancer risk due to drinking water containing 1 ug/liter of
tetrachloroethylene is estimated to be 1 x 10~6.
9-44
-------
The upper-bound cancer risk associated with 1 ug/m3 of tetrachloroethylene
in air is estimated to be 2 x 10~6. This is calculated on the basis of data
obtained from the NCI gavage study in mice (1977a), under the assumption that
the effective dose by inhalation in mice is one-sixth of that by gavage, and
the consequent assumption that tetrachloroethylene in air is about six times
more potent in mice than in humans.
The resultant estimate of the carcinogenic potency of tetrachloroethylene
in air is 1.0 x 10~2/ppm. This does not appear to be inconsistent with the
potency estimate based on the NIOSH epidemiologic study (Kaplan 1980), which is
a crude risk estimate ranging from 2.2 x 10"3/ppm to 1.6 x 10~2/ppm.
9.4.3 Conclusions
Tetrachloroethylene has been demonstrated to induce malignant tumors of
the liver in both male and female mice of the B6C3F1 strain. This constitutes
limited evidence that tetrachloroethylene may be carcinogenic in humans. It
should be recognized that there is a substantial body of opinion in the scientific
community to the effect that the mouse liver- overreacts to chlorinated organic
compounds in contrast to that of the rat, and that the induction of liver-
cancer in the mouse represents only a promoting action for- spontaneous liver-
tumors which normally occur with substantial incidence.
According to the criteria of the International Agency for- Research on
Cancer (IARC), the data supporting the carcinogenicity of tetrachloroethylene
must be classified as "limited." Since existing human epidemiologic data for
tetrachloroethylene is inconclusive, its overall IARC ranking should be Group 3,
corresponding to the conservative scientific view that tetrachloroethylene is
probably carcinogenic in humans.
9-45
-------
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Blair, A., P. Drople, and D. Grarrman. 1979. Causes of death among laundry
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Campbell, G.L., D. Cohan, and D.W. North. 1980. The application of decision
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Cox, C.B. 1972. Regression model and life tables. J. Roy. Stat Soc. B
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Crump, K.S., and W.W. Watson. 1979. GLOBAL79: A Fortran program to extrapo-
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Crump, K.S., H.A. Guess, and L.L. Deal. 1977. Confidence intervals and test
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Daffer, P., K. Crump, and M. Master-man. 1980. Asymptotic theory for
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Doll, R. 1971. Weibull distribution of cancer: Implications for models of
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Dripps, R.D., J.E. Eckenhoff, and L.D. Vandam. 1977. Introduction to
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Philadelphia PA: W.B. Saunders Company, 121-123.
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Freireich, E., E. Gehan, D. Rail, L. Schmidt, and E. Skipper. 1966. Quantitative
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Howe, R.B., and K.S. Crump. 1982. GLOBAL 82: A computer- program to extrapo-
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International Commission on Radiological Protection (ICRP). 1977. Recommendation
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9-46
-------
Replacement Page 9-45 for EPA-600/8-82-005B, "Health Assessment Document
for Tetrachloroethylene (Perchloroethylene)
The upper-bound cancer risk associated with 1 ug/m3 of tetrachloroethylene
in air is estimated to be 2 x 10~6. This is calculated on the basis of data
obtained from the NCI gavage study in mice (1977a), under the assumption .that
the effective dose by inhalation in mice is one-sixth of that by gavage, and
the consequent assumption that tetrachloroethylene 1n air is about six times
more potent in mice than in humans.
The resultant estimate of the carcinogenic potency of tetrachloroethylene
in air is 1.0 x 10'2/ppm. This does not appear to be inconsistent with the
potency estimate based on the NIOSH epidemiologic study (Kaplan 1980), which is
a crude risk estimate ranging from 2.2 x 10~3/ppm to 1.6 x 10'2/ppm.
9.4.3 Conclusions
Tetrachloroethylene has been demonstrated to induce malignant tumors of
the liver in both male and female mice of the B6C3F1 strain. This constitutes
limited evidence that tetrachloroethylene may be carcinogenic in humans. It
should be recognized that there is'a substantial body of opinion in the scientific
community to the effect that the mouse liver overreacts to chlorinated organic
compounds in contrast to that of the rat, and that the induction of liver
cancer in the mouse represents only a promoting action for spontaneous liver
tumors which normally occur with substantial incidence.
According to the criteria of the International Agency for Research on
Cancer (IARC), the data supporting the carcinogeniclty of tetrachloroethylene
must be classified as "limited." Since existing human epidemiologic data for
tetrachloroethylene is inconclusive, its overall IARC ranking should be Group 3,
w:*:*:*:*x*x*:-xtt^^
meaning that there is inadequate evidence for classifying tetrachloroethylene :*:*:*
9-45 1/25/84
-------
Innes, J.R.M., B.M. Ulland, M.G. Valeric, L. Petrocelli, L. Fishbein, E.R. Hart,
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1114.
Kaplan, S.D. 1980. Dry-cleaners workers exposed to perch!oroethylene: A retro-
spective cohort mortality study. NIOSH Contract No. 210-77-0094.
Katz, R.M., and D. Jowett. 1981. Female laundry and dry-cleaning workers in
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Krewski, D., K.S. Krump, J. Farmer, D. Gaylor, R. Howe, C. Portier, D. Salsburg,
R. Sielhen, and J. VanRyzin. 1983. A comparison of statistical methods
for low-dose extrapolation utilizing time-to-tumor data. Fundamental and
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Laskin, S., A.R. Sellakumar, M. Kuschner, N. Nelson, S. La Mendola, G.M. Rusch,
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Lin, R.S., and I.I. Kessler. 1981. A multifactorial model for pancreatic cancer
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Mantel, N., and M.A. Schneider-man. 1975. Estimating "safe" levels, a hazardous
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Health, Education and Welfare, Public Health Service, National Institutes
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National Cancer- Institute (NCI). 1977b. Bioassay of trichloroethylene for
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-------
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9-48
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APPENDIX A
COMPARISON AMONG DIFFERENT EXTRAPOLATION MODELS
Four models used for low-dose extrapolation, assuming the independent
background, are:
Multistage: P(d) = 1 - exp [-(q1d+ ... + qkdk)]
where qj are non-negative parameters.
A + B ln(d)
Probit: p(d) = / f(x) dx
oo
where f(.) is the standard normal probability density function
Weibull: P(d) = 1 - exp [-
where b and k are non-negative parameters
One-hit: P(d) = 1 - exp [-bd]
where b is a non-negative parameter.
The maximum likelihood estimates (MLE) of the parameters in the multi-
stage and one-hit models are calculated by means of the program GLOBAL82, which
was developed by Howe and Crump (1982). The MLE estimates of the parameters in
the probit and Weibull models are calculated by means of the program RISK81,
which was developed by Kovar and Krewski (1981).
Table A-l presents the MLE of parameters in each of the four models.
A-l
-------
TABLE A-l. MAXIMUM LIKELIHOOD ESTIMATE OF THE PARAMETERS FOR EACH OF THE FOUR EXTRAPOLATION MODELS BASED ON
HEPATOCELLULAR CARCINOMAS IN FEMALE MICE (doses in mg/kg/day)
Basis of interspecies
extrapolation
Multistage
Probit
Weibull
One-hit
Body surface area
q1 = 2.00 x 10
q2 = 0
-2
A = - 0.55
B = 9.80 x 10
-2
b = 0.33
k = 0.12
b = 2.00 x 10
-2
Body weights
q1 = 1.5 x 10
q2 = 0
-3
A = -0.80
B = 9.80 x 10
-2
b = 0.25
k = 0.12
b = 1.51 x 10
-3
-------
APPENDIX B
TIME-TO-EVENT DATA AND CALCULATIONS OF POTENCY
TABLE B-l. TIME-TO-DEATH IN WEEKS (H INDICATES THAT AN ANIMAL DIED OR WAS
SACRIFICED AND WAS FOUND TO HAVE A HEPATOCELLULAR CARCINOMA. ALL
ANIMALS WERE SACRIFICED AT 91 WEEKS).
Control group (vehicle): 20 animals
90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90
90, 90, 90, 90, 90, 90, 69, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90
90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 90, 68.
Low-dose group: 50 animals
91(H),
58,
66,
57,
65(H),
50,
91(H),
49(H),
62,
57,
66,
50(H),
63,
50,
91(H),
45,
61(H),
47,
66(H),
91(H),
62,
48,
87(H),
44
61,
23,
64,
91(
62,
48,
52,
60, 59
91(H), 83
62, 62
68, 67
55, 21
52,
41(H), 91(H)
51, 51(H)
High-dose group: 50 animals
91(H),
49,
50(H),
87(H),
50,
56,
56(H),
48,
91(H),
46,
49,
56,
49,
56,
53,
24,
91(H),
46,
46,
55(H),
46,
54(H).
53,
26,
63,
91(H),
45,
55(H),
91,
50(H),
52(H),
17
53,
91(H),
23,
52,
58(H),
49,
50(H),
50,
72(H),
22,
50,
56,
49,
49,
50(H)
50
91(H)
50
56(H)
The dose-response function at a time t is assumed to have the form
P(d/t) = 1 - exp [-q(d) x f(t)]
where
q(d) = q0 + q^
and
f(t) = tk.
B-l
-------
The lifetime cancer risk 1s calculated when t 1s set equal to 90 weeks. When
the dose unit 1s mg/kg/day, the maximum likelihood estimates of the parameters
are qQ = 0, qj = 1.6 x 10'9, q2 = 2.3 x NT11, and k = 3.6. At t = 90 weeks, the
95% upper-bound estimate of the linear component of the dose-response model 1s
q* = 3.5 x 10~2 mg/kg/day. These estimates are calculated using the computer
program WEIBULL82, developed by Howe and Crump (1982). In the calculation, only
the tumors that were observed before week 91 are considered to be the causes of
death for those animals in which the tumors were observed.
B-2
------- |