United States
Environmental Protection
Agency
Office of Health and
Environmental Assessment
Washington DC 20460
EPA/600/8-82/005F
July 1985
Final Report
Research and Development
&EPA
Health Assessment
Document for
Tetrachloroethylene
(Perchloroethylene)
Final
Report
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PREFACE
The Office of Health and Environmental Assessment, in consultation with an
Agency work group, 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; however, at the request
of the Agency 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 rela-
tionships are discussed, where appropriate, so that the nature of the adverse
health responses are placed in perspective with observed environmental levels.
Any information regarding sources, emissions, ambient air concentrations,
and public exposure has been included only to give the reader a preliminary
indication of the potential presence of this substance in the ambient air.
While the available information is presented as accurately as possible, it is
acknowledged to be limited and dependent in some instances on assumption
rather than specific data. This information is not intended, nor should it be
used, to support any conclusions regarding risks to public health.
If a review of the health information indicates that the Agency should
consider regulatory action for this substance, a considerable effort will be
undertaken to obtain appropriate information regarding sources, emissions, and
ambient air concentrations. Such data will provide additional information for
drawing regulatory conclusions regarding the extent and significance of public
exposure to this substance.
In view of the pending release of National Toxicology Program reports on
long-term animal studies with tetrachloroethylene, the carcinogenicity conclu-
sions of this document are considered interim and will be updated when the
reports can be evaluated.
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CONTENTS
LIST OF TABLES vi i
LIST OF FIGURES ix
AUTHORS AND REVIEWERS. xi
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 3-3
3.5 ENVIRONMENTAL FATE AND TRANSPORT 3-4
3.5.1 Ambient Air 3-4
3.5.2 Water „. 3-7
3. 6 LEVELS OF EXPOSURE 3-9
3.6.1 Mixing Ratios 3-9
3.7 ANALYTICAL METHOD. 3-18
3.7.1 Ambient Air.. 3-18
3.7.2 Water 3-22
3.7.3 Biological Media 3-24
3.7.4 Cal ibration 3-24
3.7.5 Storage and Stability of PCE 3-25
3.8 REFERENCES 3-26
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. MAMMALIAN METABOLISM AND PHARMACOKINETICS. 5-1
5.1 ABSORPTION AND DISTRIBUTION 5-1
5.1.1 Dermal Absorption 5-1
5.1.2 Oral Absorption 5-2
5.1.3 Pulmonary Absorption 5-3
5.1.4 Tissue Distribution and Concentrations 5-13
5.2 EXCRETION 5-15
5.2.1 Pulmonary Elimination in Man 5-15
5.2.2 Urinary Metabolite Excretion in Man 5-18
5.2.3 Chronic Exposure 5-20
5.2.4 Excretion Kinetics in the Rodent 5-21
5.3 MEASURES OF EXPOSURE AND BODY BURDEN 5-23
1v
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CONTENTS (continued)
5.4 METABOLISM 5-27
5.4.1 Known Metabolites 5-27
5.4.2 Enzymic Pathways of Metabolism 5-28
5.4.3 Magnitude and Dose-Dependency on Metabolism 5-32
5.4.4 Covalent Binding 5-37
5.4.5 Interactions with Metabolism 5-40
5.5 SUMMARY 5-43
5.6 REFERENCES 5-45
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., 6-3
6.1.3 Effects on Other Organs/Tissues 6-3
6.1.4 Effects on CNS and Behavior 6-4
6. 2 LABORATORY ANIMAL STUDIES 6-7
6.2.1 Lethality and Anesthesia 6-7
6. 2.2 Behavioral Effects 6-18
6.2.3 Effects on the Liver and Kidney 6-19
6.2.4 Effects on the Heart 6-23
6.2.5 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 6-25
6.4.1 Inhalation Exposure 6-25
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 7-2
7.1.3 Rabbits 7-6
7.2 SUMMARY 7-9
7.3 REFERENCES 7-10
8. MUTAGENICITY 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
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CONTENTS (continued)
8. 6 MUTAGENICITY OF METABOLITES 8-22
8. 7 SUMMARY AND CONCLUSIONS 8-23
8. 8 REFERENCES 8-25
9. CARCINOGENICITY 9-1
9.1 ANIMAL STUDIES 9-1
9.1.1 National Cancer Institute Bioassay (1977a) 9-2
9.1.2 Dow Chemical Company Inhalation Study (Rampy et al.,
1978) 9-12
9.1.3 Intraperitoneal Administration Study (Theiss et al.
1977) 9-14
9.1.4 Skin Painting Study (Van Duuren et al. , 1979) 9-16
9.2 EPIDEMIOLOGIC STUDIES 9-17
9.2.1 Kaplan (1980) 9-18
9.2.2 Blair et al. (1979) 9-22
9.2.3 Katz and Jowett (1981) 9-23
9.2.4 Lin and Kessler (1981) 9-24
9.2.5 Dun and Asal (1984) 9-24
9.2.6 Asal (personal communication, 1985) 9-26
9.3 RISK ESTIMATES FROM ANIMAL DATA 9-27
9.3.1 Selection of Animal Data 9-28
9. 3.2 Interspecies Dose Conversion 9-30
9.3.3 Choice of Risk Model 9-46
9.3.4 Unit Risk for a Compound 9-51
9.3.5 Alternative Low-Dose Extrapolation Models 9-53
9.4 CALCULATION OF UNIT RISK 9-53
9.4.1 Potency (Slope) Estimate on the Basis of Animal Data. 9-53
9.4.2 Risk Associated with 1 ng/L of PCE in Drinking Water. 9-56
9.4.3 Risk Associated with 1 ug/m3 of PCE in Air 9-56
9.5 RISK PREDICTION AGAINST HUMAN EXPERIENCE 9-63
9.6 COMPARISON OF POTENCY WITH OTHER COMPOUNDS 9-65
9. 7 SUMMARY 9-71
9. 7.1 Qualitative 9-71
9.7.2 Quantitative 9-72
9.8 CONCLUSIONS 9-73
9.9 REFERENCES 9-75
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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 3-10
4-1 Levels of PCE in tissues of marine organisms, birds, and
mammals '. 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 Physical properties of PCE and other chloroethylenes 5-2
5-2 Estimated uptake of six individuals exposed to PCE
at rest and after exercise 5-10
5-3 Disposition in rats of 14C-PCE radioactivity 72 hr following
oral or inhalation administration 5-12
5-4 Disposition in mice of 14C-PCE radioactivity 72 hr following
oral or inhalation administration 5-13
5-5 Rat organ content of PCE after daily inhalation exposure
of 200 ppm for 6 hours per day 5-14
5-6 Disposition in rats of 14C-PCE radioactivity 72 hr following
drinking water ingestion 5-23
5-7 Reported metabolites of PCE (other than TCA) 5-24
5-8 Metabolism of PCE by rat hepatic microsomes and the
effect of various inducers 5-31
5-9 Disposition of 36C1-PCE after oral administration to
Wistar rats 5-35
5-10 Irreversible hepatic binding of 14C-PCE in rats 72 hr
after exposure 5=39
5-11 Comparison of irreversible hepatic binding of 14C-PCE in
Sprague-Dawley rats and B6C3F1 mice after inhalation
(6 hr) and oral exposures 5-40
5-12 Effects of chronic oral administration of PCE on hepatic
DNA content and DNA synthesis in mice and rats 5-41
6-1 Summary of the effects of PCE on animals 6-8
6-2 Toxic dose data 6-14
7-1 Summary of reproductive/teratogenic effects of PCE in
laboratory animals 7-7
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 9-4
9-2 Cumulative survival of Sprague-Dawley rats exposed to PCE
for 12 months 9-13
9-3 Pulmonary tumor response to PCE 9-15
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LIST OF TABLES (continued)
Table
9-4 Incidence of hepatocellular carcinomas in male and female
B6C3F1 mice fed PCE by gavage 9-29
9-5 Disposition of 14C-PCE radioactivity for 72 hours after sin-
gle oral doses to Sprague-Dawley rats and B6C3F1 mice...... 9-33
9-6 Disposition of 14C-PCE radioactivity for 72 hours after
inhalation exposure for 6 hours to Sprague-Dawley rats and
B6C3F1 mice 9-39
9-7 Total hepatic macromolecular binding following oral exposure
of rats and mice to 500 mg/kg 14C-PCE 9-43
9-8 Lifetime average daily exposure (LAE) for B6C3F1 mice in the
NCI bioassay, assuming 17% of dose is metabolized 9-45
9-9 Lifetime average daily exposure (LAE) for B6C3F1 mice in the
NCI bioassay, using a dose-metabolism (in urinary excretion)
relationship. 9-45
9-10 Dose-response data and potency (slope) estimates 9-54
9-11 Relative carcinogenic potencies among 54 chemicals evaluated
by the Carcinogen Assessment Group as suspect human
carcinogens 9-67
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LIST OF FIGURES
Figure Page
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 First-order excretion curves for PCE in blood of rats after
exposure to 600 ppm for 6 hr or to 500 mg/kg gavage doses 5-4
5-2 PCE concentrations in blood and exhaled air following
i nhal ati on exposure for 4 hr. 5-5
5-3 Predicted uptake and distribution of PCE to tissue groups
during and after an 8-hr exposure to 100 ppm.................. 5-6
5-4 PCE alveolar air concentration during exposure of 5 subjects
for 8 hr at 100 ppm 5-9
5-5 Mean exhaled breath concentrations of PCE for five volunteers
exposed to 100 ppm PCE for 7 hr per day for 5 days 5-11
5-6 Daily (8-hr) occupational inhalation exposure to PCE............ 5-17
5-7 Trichloroacetic acid blood concentrations following inhalation
exposure to PCE for 4 hr 5-19
5-8 Urinary excretion of trichloroacetic acid during and following
inhalation exposure to PCE for 4 hr. 5-19
5-9 Relationship between PCE occupational inhalation exposure and
urinary concentration of total trichloro-compounds at end of
work shift for 36 male and 25 female workers. 5-20
5-10 Daily (8-hr) occupational inhalation exposure to PCE 5-21
5-11 Relationship between PCE and dose and the amount of total
urinary metabolite excreted per day by mice. 5-25
5-12 Direct linear relationship between the time-weighted average
occupational exposure to PCE over the last 4 hr of a work
day and the concentration of PCE in exhaled air 15-30 min
after the end of exposure. 5-26
5-13 Postulated scheme for the metabol i sm of PCE. 5-29
5-14 Production of TCA from PCE by hepatic microsomas from differ-
ently pretreated rats as a function of time. 5-31
5-15 Relationship between hepatotoxicity parameters from PCE oral
administration and total urinary metabolite excreted per
day by mice at increasing dose levels of PCE 5-38
6-1 Concentration-time curves for various effects of PCE 6-16
6-2 Concentration-time curves for PCE-induced anesthesia 6-17
8-1 Dose response curves of Perchlor 200 and Perchlor 230 using
Salmonella typhimurium tester strains TA100 and TA1535.... 8-7
8-2 Induction of mitotic recombination by PCE in Saccharomyces
cerevisiae D7 8-19
IX
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LIST OF FIGURES (continued)
Figure ' Page
9-1 Growth curves for male and female mice in the PCE chronic
study 9-5
9-2 Survival comparisons of male and female mice in the PCE
chronic study 9-7
9-3 Growth curves for male and female rats in the PCE chronic
study 9-8
9-4 Survival comparisons of male and female rats in the PCE
chronic study 9-9
9-5 Relationship between the PCE dose and the amount of total
urinary metabolite excreted per day by mice in each group 9-36
9-6 Histogram representing the frequency distribution of the
potency indices of 54 suspect carcinogens evaluated
by the Carcinogen Assessment Group 9-66
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AUTHORS AND REVIEWERS
The principal authors of this document are:
Vernon Benignus, Health Effects Research Laboratory, U.S. Environmental
Protection Agency, Research Triangle Park, N.C.
Chao W. Chen, Carcinogen Assessment Group, U.S. Environmental Protection
Agency, Washington, D.C.
I.W.F. Davidson, Department of Physiology and Pharmacology, The Bowman Gray
School of Medicine, Wake Forest University, Winston-Salem, N.C.
Herman J. Gibb, Carcinogen Assessment Group, U.S. Environmental Protection
Agency, Washington, D.C.
Mark M. Greenberg, Environmental Criteria and Assessment Office, U.S. Environ-
mental Protection Agency, Research Triangle Park, N.C.
Charalingayya B. Hiremath, Carcinogen Assessment Group, U.S. Environmental
Protection Agency, Washington, D.C.
Jean C. Parker, Carcinogen Assessment Group, U.S. Environmental Protection
Agency, Washington, D.C.
Vicki Vaughan-Dellarco, Reproductive Effects Assessment Group, U.S. Environ-
mental Protection Agency, Washington, D.C.
The following individuals reviewed earlier drafts of this document and submitted
valuable comments.
All Members of the
Interagency Regulatory Liaison Group
Subcommittee on Organic Solvents
Dr. Joseph Borzelleca
Dept. of Pharmacology
The Medical College of Virginia
Virginia Commonwealth University
Richmond, VA 23298
Dr. Mildred Christian
Argus Laboratories
Perkasie, PA 18944
Dr. Herbert Cornish
Dept. of Environmental and Industrial Health
University of Michigan
Ypsilanti, MI 48197
XI
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Dr. John Egle
Dept. of Pharmacology
Virginia Commonwealth University
Richmond, VA 23298
Dr. Lawrence Fishbein
National Center for lexicological Research
Jefferson, AR 72079
Dr. Thomas Haley
National Center for Toxicology Research
Jefferson, AK 72079
Dr. Rudolf Jaeger
Institute of Environmental Medicine
New York, NY 10016
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
Dr. Norman Trieff
Dept. of Preventive Medicine
University of Texas Medical Branch
Galveston, TX 77550
Dr. Benjamin Van Duuren
Institute of Environmental Medicine
New York University Medical Center
New York, NY 10016
Dr. James Withey
Food Directorate
Bureau of Food Chemistry
Ottawa, Canada
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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.
Robert P. Beliles, Ph.D.
Chao W. Chen, Ph.D.
Margaret M.L. Chu, Ph.D.
I.W.F. Davidson, Ph.D
Herman J. Gibb, B.S. ,
Bernard H. Haberman, D.V.M., M.S
Charalingayya B. Hiremath, Ph.D.
Robert E. McGaughy, Ph.D.
Charles H. Ris, M.S., P.E.
Dharm W. Singh, D.V.M., Ph.D.
Todd W. Thorslund, Sc.D.
(consultant)
M.P.H.
Participating Members of the Reproductive Effects Assessment Group
John R. Fowle III, Ph.D.
K.S. Lavappa, Ph.D.
Sheila Rosenthal, Ph.D.
Carol Sakai, Ph.D.
Daniel S. Straus, Ph.D. (consultant)
Vicki Vaughan-Dellarco, Ph.D.
Lawrence R. Valcovic, Ph.D.
Peter E. Voytek, Ph.D. (Chairman)
Members of the Agency Work Group on Solvents
Elizabeth L. Anderson
Charles H. Ris
Jean Parker
Mark Greenberg
Cynthia Sonich
Steve Lutkenhoff
Arnold Edelman
James A. Stewart
Paul Price
William Lappenbusch
Hugh Spitzer
David R. Patrick
Lois Jacob
Josephine Brecher
Mike Ruggiero
Charles Delos
Jan Jablonski
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 Tcxic Substances
Office of Drinking Water
Consumer Product Safety Commission
Office of Air Quality Planning and Standards
Office of General Enforcement
Office of Water Regulations and Standards
Office of Water Regulations and Standards
Office Water Regulations and Standards
Office of Solid Waste
Office of Pesticide Programs
Office of Emergency and Remedial Response
XI11
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1. EXECUTIVE SUMMARY
Tetrachloroethylene (PCE) is a moderately volatile chlorinated hydro-
carbon which has important commercial applications in the dry cleaning of
fabrics and in the degreasing of fabricated metal parts. It is estimated that
approximately 265,000 metric tons were produced in the United States in 1982.
Approximately 90 percent of production is estimated to be released eventually
to the atmosphere. Because PCE is relatively insoluble in water (150 mg/L)
and has a vapor pressure of 19 torr at 25°C, PCE in natural waters would be
conveyed 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/m^ 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/m^. PCE has been detected less frequently in water;
it has been monitored in surface and drinking waters, generally at levels
between 1 and 2 ppb. In certain instances involving contamination of ground-
water, 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
degradation 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
concentrations of ozone or other oxidants; thus, PCE is not believed to be a
significant factor in production of photochemically induced pollution often
experienced 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
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basis, PCE levels fluctuate considerably.
Inhalation is the principal route of concern by which PCE enters the body.
Ingestion of drinking water contaminated by PCE is another important concern.
PCE is rapidly and virtually completely absorbed following oral administration,
while pulmonary uptake of PCE during inhalation exposure is linearly propor-
tional to exposure duration and the air concentration; pulmonary uptake is
also influenced by physical activity and body mass. The metabolism and
pharmacokinetics of PCE are highly contingent on its physicochemical proper-
ties. Controlled inhalation studies with human volunteers (at 100 ppm)
suggest that whole-body, steady-state conditions are not established within
short exposure periods (e.g., 8 hours). Because partitioning into adipose
tissue is slow, steady-state (the rate at which whole body uptake is balanced
by clearance) may require considerably longer periods of exposure. PCE
distributes widely into body tissues. PCE is eliminated from the body mainly
by the pulmonary excretion of unchanged parent compound. Limited metabolism
of PCE occurs; metabolism is dose-dependent and saturable in mice, rats, and
humans; in humans saturation would not be expected until air exposure levels
approximate 100 ppm (678 mg/m3) or greater. The principal site of metabolism
is in the liver where PCE is oxidized to PCE oxide, which rearranges to
trichloroacetic acid. Controlled studies with humans have demonstrated that
PCE metabolism (urinary trichloroacetic acid) represents 1 to 3 percent of
the amount of PCE absorbed during 8-hour exposures to between 100 to 400 ppm
(678 to 2,112 mg/m3)-. While the metabolic profiles of PCE are not as yet
fully established in mice, rats, and humans, there is no convincing evidence
of differences in the pathways in these species. PCE metabolites are known
to covalently bind, in vitro and in vivo, to cellular macromolecules such as
protein and lipid. Since tissue-bound metabolites have a slow rate of turn-
over, cumulative cellular changes may occur in humans with chronic exposure.
Indices of hepatotoxicity of PCE in the rat and mouse have been shown to be
highly correlative with the dose-dependent nature of PCE metabolism.
Excluding carcinogenicity as an end point, toxicity testing in experi-
mental animals, coupled with limited human data derived principally from
overexposure situations, suggest that long-term exposure of humans to
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environmental levels found or expected (10 ppb or less) of PCE is not likely
to present a health concern.
Decrements in task performance and coordination are the first gross signs
of central nervous system (CNS) and behavioral alterations observed in con-
trolled human studies in which individuals were exposed to about 100 ppm (678
mg/m^) for up to 7 hours. More sensitive tests, however, would have to be per-
formed to determine if PCE affects the nervous system at even lower concentra-
ti ons.
Transient liver damage in humans is generally associated with short-term
exposures greatly in excess of 100 ppm (678 mg/m^). In rodent species tested,
intermittent or prolonged exposure to PCE has been observed to result in liver
and kidney damage at levels exceeding 200 ppm (1,356 mg/m^). Since ambient
air levels are generally orders of magnitude lower than that associated with
liver damage, long-term exposure of humans to ambient air or water levels
would not be expected to cause adverse liver or kidney effects.
Similarly, ambient air and water levels of PCE are unlikely to cause
adverse effects upon the heart.
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 primarily reflect delayed
development and generally 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 irre-
versible manner. The teratogenic potential of PCE for humans is unknown.
PCE has been evaluated for its ability to cause gene mutation, chromoso-
mal aberrations, unscheduled DNA synthesis, and mitotic recombination. These
tests were conducted using bacteria, Drosophi1 a, yeast, cultured mammalian
cells, whole mammal systems, and cytogenetic analyses of exposed humans. Cer-
tain 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 were observed at high concentrations that were cyto-
toxic; dose-dependent relationships were not established. The positive
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findings may be the result of mutagenic contaminants and/or added stabilizers.
Several other tests of commercial and technical samples of PCE 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.
In a gavage bioassay, PCE has induced a statistically significant increase
of malignant liver tumors in both male and female B6C3F1 mice. No carcinogenic
effects were observed in lifetime studies of rats exposed to PCE in gavage and
inhalation bioassays; however, these latter studies had diminished sensitivity
to detect a response due to excessive dose-related mortality in the gavage
study and a low dose level in the inhalation study.
Intraperitoneal injection of PCE in Strain A mice did not produce an
increased incidence of pulmonary adenomas,; mouse skin initiation-promotion
experiments also did not show a tumor response. However, because of inherent
limitations in these assays, the negative results do not detract from the
positive findings of liver tumors in mice.
A cohort study of dry-cleaning workers exposed to PCE showed that workers
were at an elevated risk of colon cancer mortality; however, the elevated risk
cannot be totally assigned to PCE since as many as one-half of the workers may
have been exposed to petroleum distillates in their working history. Other
studies that found an association between cancer and employment in the dry-
cleaning industry did not identify the dry-cleaning solvents to which the
employees were exposed.
The positive response in both male and female mice constitutes a signal
that PCE and/or its reactive metabolites might be a carcinogen for humans. In
terms of strength of evidence in animal test systems, the mouse bioassay
constitutes limited evidence.
According to the Agency's Proposed Guidelines for Carcinogen Risk
Assessment, the cancer evidence of PCE in animal test systems is limited, and
the cancer evidence in epidemiologic studies is inconclusive. The overall
weight-of-evidence classification for PCE would be Group C, i.e., a possible
human carcinogen.
When the criteria of the International Agency for Research on Cancer
(IARC) are applied, the animal data supporting the carcinogenicity of PCE is
classified as limited. Because existing epidemiologic data for PCE is
inconclusive, its overall IARC ranking should be Group 3, meaning that the
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evidence is inadequate to evaluate the carcinogenic potential of PCE.
The NTP is completing PCE inhalation cancer bioassays for rats and mice.
These findings should be available for scientific review in the fall of 1985.
When the results of these bioassays are made available, consideration will be
given to updating the carcinogenicity evaluation in this assessment document.
Although PCE has only limited carcinogenicity evidence, a carcinogenic
potency and related upper-bound estimate of incremental lifetime cancer risk
can be estimated from the NCI male and female mouse bioassay.
The development of these risk estimates is for the purpose of evaluating
the "what if" question: If PCE is carcinogenic in humans, what is the possible
magnitude of the public health impact? Any use of the risk estimates should
include a recognition of the weight-of-evidence likelihood for the carcinogenic
potential of PCE in humans. The upper-bound incremental cancer risk is
calculated to be 5.0 x 10~2 (mg/kg/day )"•'- for a continuous lifetime exposure
to PCE, under the presumption that PCE is carcinogenic in humans. The upper-
bound nature of this estimate is such that the true risk is not likely to
exceed this value, and may be lower. Expressed in terms of relative potency,
PCE ranks in the lowest quartile among the 54 suspect or known human carcino-
gens evaluated by EPA's Carcinogen Assessment Group.
1-5
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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, tetra-
chloroethene, and 1,1,2,2-tetrachloroethylene.
PCE, though a water and solid waste contaminant, is primarily 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, princi-
pally 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 chem-
istries, and pharraacokinetic parameters of PCE exposure.
Since epidemiologic 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 increases the confidence that similar
effects may occur in humans.
This document is intended to provide an evaluation of the scientific data
base concerning PCE. The publications 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.
The basic literature search that supports this assessment is current up
to 1984. On-going literature searches have been conducted through 1985,
resulting in the inclusion of selected references in all chapters of this
document.
2-1
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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-1ike 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/L)
(Handbook of Chemistry and Physics, 1976; Chemical and Process Technology
Encyclopedia, 1974). Its specific gravity at 20°C is 1.624. Its CAS registry
number is 127-18-4. In air, at 25°C and standard pressure, 1 part per million
(ppm) is equivalent to 6.78 mg/m3.
PCE has negligible photochemical reactivity (Dimitriades et a!., 1983)
and, in the troposphere, is decomposed via free radical mechanisms. When in
contact with water for prolonged periods, PCE slowly decomposes to yield tri-
chloroacetic and hydrochloric acids. Upon prolonged storage in light, it was
reported to decompose slowly to trichloroacetyl chloride and phosgene by auto-
oxidation (Hardie, 1966). At 700°C, it decomposes, when in contact with acti-
vated charcoal, to hexachloroethane and hexachlorobenzene (Gonikberg, 1956).
PCE has a boiling point of 121.1°C at 760 mm Hg and a vapor pressure 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
PCE may be produced by several processes:
1. Chlorination of trichloroethylene:
onop
CHC1 = CC12 + C12 frFf* CHC12CC13
3-1
-------
2CHC12CC13 + Ca(OH)2 ^-H-C12C = CC12 + CaCl2 + 2H20
2. Dehydrochlorination of S-tetrachloroethane:
CHC12-CHC12 + C12 » CC12 - CC12 + 2HC1
3. Oxygenation of S-tetrachloroethane:
2CHC12CHC12 + 02 > 2CC12 = CC12 + 2H20
4. Chlorination of acetylene:
CC12 - CC12 + C12 200^cc-,3-CC-|3
CH - CH + 3 CC1 CC1 2QO-400°C.4CC12 = CC12 + 2HC1
CH - CH + 3 CU3LU3 Cata1yst
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 produced in the United States is derived from the
oxychlorination of 1,2-dichloroethane (reaction 6) or via Chlorination of
hydrocarbons (reaction 5) (Lowenheim and Moran, 1975).
In 1980, 329,000 metric tons of PCE were consumed in the United States
(SRI, 1982). This figure represents production plus imports, minus exports.
According to the U.S. International Trade Commission (1983), 265,770 metric
tons were produced in the United States in 1982. Production in 1983 was esti
mated at 263,000 metric tons (Chemical and Engineering News, 1984).
3-2
-------
TABLE 3-1. MAJOR U.S. PRODUCERS OF PCEC
Organization
Yearly 1981 capacity, MTC
Dow Chemical
PPG
Vulcan
Diamond Shamrock
Ethyl Corporation
E. I. du Pont de Nemours
Stauffer Chemical
145
91
91
75
73
32
Adapted from Chemical Economics Handbook (SRI, 1983). MT = Metric tons.
Terminated production (Halogenated Solvents Industry Alliance, 1983).
C0utput for captive use only.
3.3 USE
PCE has the following uses (Gosselin et al., 1976; Fishbein, 1977):
(1) dry cleaning solvent; (2) textile scouring solvent; (3) dried vegetable
fumigant; (4) rug and upholstery cleaner; (5) stain, spot, lipstick, and rust
remover; (6) paint remover; (7) heat transfer media ingredient; (8) chemical
intermediate in the production of other organic compounds; and (9) metal de-
greaser.
It is estimated that the use of PCE in the dry cleaning industry repre-
sents about 42 percent of 1980 consumption in the United States (SRI, 1982).
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. Data
available concerning discharges to water are discussed in Section 3.6.1.2.
Emissions estimates reflect a diversity of sources throughout the country.
Dry cleaning operations are located primarily in urban areas. Approximately
3-3
-------
26,000 establishments are estimated to exist, according to Bureau of Census
data (U.S. EPA, 1979). There are approximately 18,000 retail establishments
plus much smaller numbers of industrial and coin-operated facilities (Inter-
national Fabricare Institute, 1984).
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 ap-
proximately 90 percent of the amount of PCE produced 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 03 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 ac-
cording 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 106 molecules cm 3 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
adjusted 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 sup-
ported by the field measurements of Campbell et al., 1979) and a weighted global
3-4
-------
average temperature (265°K), an average residence time of PCE would be calcu-
lated to be 292 days or 0.8 year (Singh et al. , 1979).
Estimations of a residence time of PCE of one year or less have been re-
ported by a number of investigators (Dilling, 1982; 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 having high reactant concentrations, reacts to form 03 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 contribute less to 03 production than equal
concentrations of ethane. Ethane is regarded by the authors to be a boundary
species separating the reactive volatile organics from the unreactive ones.
Studies on PCE reactions with 03, 0, and -OH have indicated that rate con-
stants are lower than with Cl (Dimitriades et al. , 1983).
Gay et al. (1976) had determined that trichloroacetyl chloride was a photo-
oxidation product of PCE in smog chamber studies. While the studies of Gay et
al. (1976) indicate that trichloroacetyl chloride may be formed through chlorine
atom migration in an epoxide intermediate, evaluation of -OH and oxygen atom
rate constants indicate that less than 1 percent of PCE in ambient air reacts
with atomic oxygen and, of the activated epoxides formed, only a small percent-
age undergo rearrangement (Graedel, 1978). Dimitriades et al. (1983) found
3-5
-------
that, on irradiation, the only product observed was CO. Phosgene was not de-
tected. When 2 ppm (13.6 mg/m3) PCE and 20 ppb trichloroacetyl chloride were
.irradiated together, the phosgene level reached 0.1 ppm.
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- + 02 »• HOC(C1)2C(C1)202-
HOC(C1)2C(C1)202- oxygen »• HOC(C1)2C(C1)20- (3-2)
abstraction
HOC(C1)2C(C1)20- > HOC(C1)2- + COC12
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)
CC12CC12OH- + 02 > 02CC12CC12OH
02CC12CC12OH + NO > COC1CC12OH- + N02 + Cl•
•OH + COC1CC12OH > COC1COC1 + H20 + Cl•
3-6
-------
Compared to other ethylene compounds studies, Howard (1976) reported that PCE
exhibits unusually low reactivity toward hydroxyl radicals.
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. 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 hydrol-
ysis 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 reac-
tivity. The effect of solar flux was calculated by Altshuller (1980) who esti-
mated 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.
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 L of sea water, PCE levels decreased 50 percent within
3-7
-------
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.
Billing 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
is lost from water.
Parsons et al. (1984) obtained evidence that PCE can be converted to other
chlorinated compounds following incubation of PCE in microcosms containing
muck from an aquifer recharge basin. Results suggested that PCE can be con-
verted to trichloroethylene, chloroethene, cis- and trans-l,2-dichloroethene.
and dichloromethane (methylene chloride).
3-8
-------
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
6 uM. 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 evapo-
ration is the dominant elimination process from surface waters. The annual re-
lease 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.
Wood et al. (1981) demonstrated that PCE in sediment-water samples can be
degraded under anaerobic conditions. PCE was not degraded in autoclaved con-
trols. Under experimental conditions, the resulting half-life for PCE was 34
days.
3.6 LEVELS OF EXPOSURE
3.6.1. Mixing Ratios
3.6.1.1 Ambient Air--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, tropos-
pheric 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 10~3
±0.02 x 10 3 mg/m3), respectively. Globally, the average background level of
3-9
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TABLE 3-2. AMBIENT AIR MIXING RATIOS OF PCE MEASURED AT SITES AROUND THE WORLD
Location
Alabama
Birmingham
Arizona
Grand Canyon3
Phoenix
California
San Bernardino Mtns.
oo Badger Pass
h"—*
O Point Arena
Stanford Hills
Point Reyes
Dominguez
El Cajon
La Jolla
Los Angeles
Menlo Park
Mill Valley
Mt. Cuyamaca
Date of Reported Concentration, ppb (mg/m3)
Measurement Maximum Minimum Average
April 12-22, 1977 0.008 0 0.001 ± 0.003
Nov. 28-Dec. 5, 1977 0 00
Apr. 23-May 6, 1979b 3.696 (0.025) 0.129 0.9938± 0.7155
(0.0008) (0.0067 ± 0.0048)
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 10~4)
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
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)
CO
1
1— >
1 — •
Location
California
Oakland
Palm Springs
Riverside
San Jose
Santa Monica
Upland
Colorado3
Denver
Delaware
Delaware City
Louisiana
Baton Rouge, Geismar,
and Plaquemine
Maryland
Baltimore
Michigan
Detroit
New Jersey
Bayonne
Date of
Measurement
June 30-July 8, 1979
May 5-11, 1976
April 25, 1977-
July 12, 1980
Aug. 21-27, 1978
April 6, 1974
Aug. 13-Sept. 23, 1977
Jun 16-26, 1980
July 8-10, 1974
Fall, 1978
July 11-12, 1974
Oct 27-Nov 5, 1978b
March, 1973-Dec. , 1973
Reported Concentration, ppb (mg/m3)
Maximum Minimum Average
0.64 0.12 0.31 ± 0.17
1.1 0.12 0.28 ± 0.084
0.98 0.37 0.49 + 0.13
1.1 ± 0.036
2.3
1.1 0.01 0.19 ± 0.35
0.47 0.24 0.39 ± 0.077
0.51 (0.0034) <0.02 (<0.0001) 0.24 (0.0016)
0.18 (0.001) .001 (0.007 x 10~3) 0.017 (0.118 X 10~~3)
0.29 (0.0019) <0.02 (<0.0001) 0.18 (0.0012)
2.16 (0.004) <0.1 (<0.001) 0.35 (0.002)
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; 1980
Singh et al. , 1979
Su and Goldberg, 1976
Pellizzari , 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
New Jersey
New Brunswick
New Brunswick
Seagirt
Sandy Hook
Boundbrook, Rahway,
Edison and Passaic
Batsto3
Bridgeport3
Burlington3
Camden3
Carlstadt3
Edison3
Elizabeth3
Fords3
Middlesex3
Newark3
Date of
Measurement
-
-
June 18-19, 1974
July 2, 1974
Sept 18-22, 1978
Feb. 26-Dec. 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-
Dec. 29, 1979
Mar. 26, 1976-
Sept. 27, 1978
July 23-28, 1978
Mar. 23, 1976-
Dec. 29, 1979
Reported Concentration,
Maximum
-
-
0.88 (0.059)
1.4 (95 x 10~4)
58 (0.394)c
0.53
0.041
30
5.5
8.7
14
5.6
0.21
32
ppb (mg/m3)
Minimum
0.5 (0.003)
0.12 (0.0081)
0.10 (0.067)
0.15 (10 x 10"«)
trace
0
0
0
1.1
0.11
0
0
0
0
Average
-
-
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 + 0.090
1.3 ± 3.1
Reference
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
Bozzelli and Kebbekus, 1979
Bozzelli and Kebbekus, 1979;
Bozzelli et al . , 1980;
Pellizzari, 1977
-------
TABLE 3-2. (continued)
CO
1
I—1
CO
Location
New Jersey
Rahway8
Rutherford9
Somerset3
South Amboy
New York
New York City
Niagara Falls and
Buffalo
Whiteface Mtn.
Ohio9
Wilmington
Texas
Houston
Aldinea
Deer Park8
El Pasoa
Freeport8
Houston3
Date r'
Measurement
SepL. 20-22, 1978
May 1, 1978-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,
Maximum
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~")
4.52 (0.030)
0.037
0.15
0.39
1.3
ppb (mg/ma)
Minimum
2.7
0
0
0
1.0 (0.006)
0.16 (0.001)
0,02 (0.122 x 1Q~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)
CO
-pi
Location
Texas
LaPorte3
Pasadena3
Utah3
Magna
Washington3
Auburn
West Virginia3
Charleston
St. Albans
aOata obtained
24-hour value.
Sampling time:
Sampling time:
Date °f Reported Concentration, ppb (mg/m3)
Measurement Maximum Minimum Average Reference
Aug. 12-13, 1976 0.49 Pellizzari et al., 1979
July 28, 1976 0.003 Pellizzari et al . , 1979
Oct. 24-Nov. 3, 1977 0.012 0 0.004 ± 0.006 Pellizzari, 1979
Jan. 10-11, 1977 0.76 0.18 0.59 ± 0.27 Battelle, 1977
Sept. 27-Nov. 20, 1977 0.16 0 0.004 ±0.008 Pellizzari, 1978
Sept. 27-Oct. 25, 1977 0.064 0 0.016 ±0.032 Pellizzari, 1978
from summary report of Brodzinsky and Singh, 1982. Values reported are 24-hour sampling concentrations.
14 minutes.
100 minutes.
-------
PCE was 0.026 ± 0.007.7 ppb (1.7 x 10 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 0.8 ppb.
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).
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.
However, Tenax is reported to generate artifacts (Singh, 1982).
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 di-
sulfide 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. In the U.S. Environmental
Protection Agency's STORET System for the period from August 1977 to September
1984, the mean concentration of PCE in all measured water supplies was 2 ug/L.
Maximum observed concentrations were about 20 ug/L. The information was com-
piled from 66 measurement stations in 10 states. With respect to streams and
other surface waters, analysis of data from 1,102 measurement stations in 45
states indicates that, from August 1975 to September 1984, the mean PCE concen-
tration was 1 ug/L.
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
3-15
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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
Mississippi River water as well as in treated water. An approximate six-fold
reduction in concentration occurred after sedimentation and chlorination. PCE
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
limits were not reported. The highest level reported was 45 (jg/L. In all
samples taken in California, Oregon, and Washington, PCE was either not detec-
ted or was found at a concentration of 1 pg/L or less. Sampling 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.
PCE was among a number of halogenated organics found in community drinking
water supply wells in Nassau County, New York. Because of contamination, 16
of these wells were closed by the New York State Health Department (Ewing et
al., 1977). The maximum detected level of PCE in the contaminated wells was
375 (jg/L. 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 detec-
ted 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).
3-16
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Lochner (1976) found that levels of PCE in Bavarian lake waters ranged
from 0.015 to 3 ppb (0.015 x 10~3 to 2.7 x 10~3 mg/L). European surface waters
were reported to have uniform PCE concentrations ranging from 0.2 x 10"3 to
0.002 mg/L. 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 a!., 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 |jg/L. The treated water before
chlorination contained 3.9 |jg/L PCE. After chlorination, the effluent contained
4.2 [jg/L PCE.
PCE has been detected in the drinking water of a number of U.S. cities.
These include Evansville, Indiana (Keith et al., 1977); Kirkwood, 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 pg/L. An
exception was Jefferson Parish, which had a measured concentration of 5 ppb (5
ug/L). Keith et al. (1977) did not detect PCE in the drinking water of Phila-
delphia. PCE 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. PCE is used as a solvent during
the application of the vinyl coating to the pipe during manufacturing. It was
suggested that residual solvent leaches into the water carried in these pipes.
3-17
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Using a charcoal trap with flame ionization detection, Wakeham and co-
workers (1980) detected PCE 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 Lochner (1976). Samples
taken at various sampling points and times gave a range of 1.1 x 10 3 to 2.4 x
10"3 mg/L. Raw sewage at Munich contained 0.088 mg/L PCE. On mechanical clari-
fication, the 24-hour average concentration of PCE was 0.0068 mg/L.
3.6.1.3 Sediments--Sediment levels for PCE documented in the U.S. Environmental
Protection Agency's STORE! system indicate that mean levels from 1,102 measure-
ment stations in 45 states are about 33 M9/kg (dry weight). Maximum reported
levels have been as high as 3,700 ug/kg. Measurement period covers June, 1977
to November, 1984.
3.7 ANALYTICAL METHODOLOGY
PCE has been analyzed in air and in water, as well as in biological fluids,
by a variety of methods. Separation of PCE from other compounds 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 detection 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).
3-18
-------
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 p.g/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
trap. The estimated detection limit, when high resolution GC/MS is used, is
0.38 ppt (2.5 x 10 6 mg/m3). 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 reproducibi1ity 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
3-19
-------
hydrogen at 1000°C. Carbowax 20M, reacted with nitroterephthal ic 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/m3) 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.
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.
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.
3-20
-------
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
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 ampere. 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 ECD 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 is 75 percent or greater with compounds
3-21
-------
having ionization efficiencies exceeding 50 percent. 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 re-
producibility 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
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, a level unattainable under nearly all
conditions. Precision was reported to be within 2 percent.
To measure PCE at levels expected to occur in occupational air, flame
ionization detection has been used. The analytical method S335, suggested by
the National Institute for Occupational Safety and Health (1977) for organic
solvents in air, utilizes adsorption onto charcoal, followed by desorption with
carbon disulfide (CS2). 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, interfer-
ences are minimal, and those that do occur can be eliminated by altering chroma-
tographic 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
3-22
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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
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 ug/L 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/L with ECD, and 1 ug/L 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/L
can be sampled (Mieure, 1980). Mieure (1980) reported a detection limit for
PCE (analyzed by ECD) of 0.01 ug/L. With flame ionization detection, the limit
was 32 ug/L. 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 Caste!lo 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/L 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.
3-23
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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
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 CS2 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.
3-24
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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--Eva1uation 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
indicated that glass tubes can be acceptable containers for breath samples if
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.
3-25
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3.8 REFERENCES
.Altshuller, A. P. 1980. Lifetimes of organic molecules in the troposphere and
lower stratosphere. Adv. Environ. Sci. Technol. 10:181-219.
Bellar, T. A., and J. J. Lichtenberg. 1974. Determining volatile organics at
the microgram per litre levels by gas chromatography. J. Am. Water Works
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4. ECOSYSTEM CONSIDERATIONS
4.1 EFFECTS ON AQUATIC ORGANISMS AND PLANTS
Tetrachloroethylene (PCE) has been tested for acute aquatic toxicity in
12 species. The information presented in this chapter presents observed
levels reported to result in adverse effects under laboratory conditions. It
is recognized that such parameters of toxicity are not easily extrapolated to
environmental situations. Test populations themselves may not be representa-
tive of the entire species, in which susceptibility of various lifestages to
the test substance may vary considerably. Guidelines for the utilization 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 Methor^
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/L (18.4 ppm); the 95 percent confidence limits
were 14.8 to 21.3 mg/L (14.8 to 21.3 ppm). In comparison, the static experi-
ments gave a 96-hour LC50 of 21.4 mg/L (21.4 ppm); the 95 percent confidence
limits were 16.5 to 26.4 mg/L (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.
4-1
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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/L (14.4 ppm).
The 96-hour LC50, in a static test with the bluegill (Lepomis macrochirus).
was reported as 12.9 mg/L (12.9 ppm) (U.S. EPA, 1978, 1980). The most sensitive
species tested is the rainbow trout (Salmo gairdneri). The LC50 determined by
a flow-through measured procedure was 5.28 mg/L (5.28 ppm) (U.S. EPA, 1980).
With embryo-larval test procedures, a chronic value of 0.840 mg/L (0.840 ppm)
was obtained by the U.S. EPA (1980) for the fathead minnow.
With the freshwater invertebrate Daphnia magna, a 48-hour EC50 value of
17.7 mg/L (17.7 ppm) was obtained (U.S. EPA, 1980). The midge Tanytarsus
dissimilis was more resistant, with a 48-hour LC50 value of 30.84 mg/L (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/L (816 ppm).
For the saltwater species Skeletonema costatum a 96-hour EC50 of about
500 mg/L (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/L (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 limanda), barnacle larvae (Barnacle nauplii), and on unicell-
ular algae (Phaeodactylum tricornutum). The LC50 was 5 mg/L (5 ppm) for the
dab. The 48-hour LC50 for barnacle larvae was 3.5 mg/L (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/L (10.5 ppm).
4-2
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Data collected by the U.S. Environmental Protection Agency (1980) indi-
cate that, for mysid shrimp (Mysidopsis bahia), the LC50 was 10.2 mg/L (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/L (0.45 ppm)
(U.S. EPA, 1980). The chronic value is 0.044 times the 96-hour LC50.
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 bioconcentra-
tion factor (K, f) is a constant of proportionality between the concentration
of the chemical in fish (Cf) and in the water (C ). This can be more clearly
expressed as
Kl
at steady state. (4-1)
Cw
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
partition 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). According to the
American Society for Testing and Materials (ASTM), a log partition coefficient
exceeding a value of three is considered an indication of a high probability
of measurable bioaccumulation in aquatic species (ASTM, 1978). Compounds that
exhibit a large log coefficient generally are those with low water solubility
and high solubility in organic solvents. Although a compound may demonstrate
4-3
-------
a high BCF or log partition coefficient, other environmental factors that act
to reduce this potential often exist. The compound may be rapidly hydrolyzed
or degraded by other mechanisms. Measurable uptake by the organism may be
precluded 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
partition 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 LC50 level for this species or
whether a steady-state was achieved was not reported. A direct measurement
for BCF of 49 for the bluegill (whole body) is cited in the water quality
criteria document (U.S. EPA, 1980). 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 1 to 9 ppb. The highest concentration (8 to 9 ppb) found was in
the crab (Cancer pagurus). Higher levels were found in marine algae (13 to 20
ppb). In tissues of fish, a range of 0.3 to 41 ppb 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, 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 concentra-
tions measured in seawater, indicating little or no bioconcentration and biomag-
nification (U.S. EPA, 1981).
4-4
-------
TABLE 4-1. LEVELS OF PCE IN TISSUES OF MARINE ORGANISMS, BIRDS, AND MAMMALS
Species
Invertebrates
Plankton
Plankton
Ragworm (Nereis diversicolor)
Mussel (Mytilus edulis)
Cockle (Cerastoderma edule)
Oyster (Ostrea edulis)
Whelk (Buccinum undatum)
Slipper limpet (Crej) \dula
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
Liverpool Bay
Firth of Forth
Firth of Forth
Firth of Forth
Thames Estuary
Trichloro-
ethylene
Tissue (ppb by mass
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
on wet tissue)
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 flesus)
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 (ppb by mass
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
!•;..»•/. no on
PCE
on wet tissue)
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
n c _ -an
-------
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 lus"us)
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
(ppb by mass
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
on wet tissue)
1
not detected
5.1
3
3
4
1
1
2
4
2
1
Not detected
1.6
<0.1
3.6
1.5 - 3.2
4.5-26
Shag (Phalacrocerax aristotelis) Irish Sea
eggs
2.4
1.4
-------
TABLE 4-1. (continued)
oo
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
1 iver
muscle
eggs
eggs
bl ubber
1 iver
-
Trichloro-
ethylene
(ppb by mass
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
on wet tissue)
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
Tissue
flesh
1 i ver
flesh
1 i ver
flesh
1 i ver
Period of
Exposure (days)
3-35
3-35
3-35
3 - 35
10
10
Mean Exposure
Concentration (ppm)
0.3
0.3
0.03
0.03
0.2
0.2
Mean
Concentration in
Tissue (ppm)
2.8a (13)
113 (14)
0.16 (9)
7.4b (9)
1.3 (7)
69 (7)
Accumu-
lation
Factor
x 9
x 400
x 5
x 200
x 6
x 350
Numbers in parentheses are number of specimens analyzed.
-, 6
One fish had a flesh concentration of 29.7/10 and was omitted from calcula-
tions.
h 6
One fish had flesh concentration of 50.3/10 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 limanda) 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
4-9
-------
100
O LIVER ACCUMULATION
D LIVER LOSS
A FLESH
EXPOSURE TIME, rlays
Figure 4-1. Accumulation and loss of PCE by dabs.
Source: Alexander et al. , 1978.
4-10
-------
100
10
E
Q.
0.
I
CO
CJ
CL
0.1
0.01
T
EXPOSURE LEVELS, ppm
O 0.3
D
A
0.2
0.02
O
10 100
PCE IN LIVER, ppm
1000
Figure 4-2. Relation between flesh and liver concentration of PCE in dabs
Source: Alexander et al., 1978.
4-11
-------
TABLE 4-3. CONCENTRATION OF PCE AND TRICHLOROETHYLENE IN MOLLUSKS AND
FISH NEAR THE ISLE OF MAN
Species
Eel (Conger conger)
brain
gill
gut
1 i ver
muscle
Cod (Gadus morhua)
brain
gill
heart
1 iver
muscle
skeletal tissue
stomach
Coalfish (Pollachius birens)
alimentary canal
brain
PCE
(mg x
6
2
3
43
1
3
3
8
2
-
6
-
TrichloroethyTene
lOVg dry weight tissue)
62
29
29
43
70
56
21
11
66
8
-
7
306
71
gill
heart
1 iver
muscle
Dogfish (Scyl1iorhinus canicula)
brain
gin
gut
heart
1 iver
muscle
spleen
Bib (Trisopterus luscus)
brain
gut
1 iver
muscle
skeletal tissue
6
2
12
13
4
0.3
70
40
176
41
274
479
41
307
143
187
185
4-12
-------
TABLE 4-3. (continued)
PCE Trichloroethylene
Species (mg x 106/g dry weight tissue)
Baccinum undatum
digestive gland 33 2
muscle 39
Modiolus modiolus
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.
in seawater, there was only a slight enrichment in the tissues (< 25 times).
PCE had one of the lowest mean bioconcentration factors.
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 I 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. Dilling 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
4-13
-------
more nearly like those found in the environment. Evaporation of PCE from the
hydrosphere is a rapid process.
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)
kc
•v
where k is the volatilization rate constant.
Using the data provided by Smith et al. (1980), the t1/2 (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/L 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 Schwarzeir
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
4-14
-------
of organic carbon in the sorbent. A high degree of correlation was found
between the partition coefficient and organic carbon content when the fraction
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 pg/L,
respectively. For saltwater aquatic life, the acute and chronic toxicity
values are 10,200 and 450 M9/L, respectively.
PCE 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.
4-15
-------
4.5 REFERENCES
Alexander, H. C., W. M. McCarty, and E. A. Bartlett. 1978. Toxicity of
perchloroethylene, trichloroethylene, 1,1,1-trichloroethane, and
methylene chloride to fathead minnows. Bull. Environ. Contam. Toxicol.
20:344-352.
American Society for Testing and Materials. 1978. Estimating the hazard of
chemical substances to aquatic life. J. Cavins, K. L. Dickson, and A.
W. Maki, eds., STP 657. Committee D-19 on Water.
Bouwer, E. J., P. L. McCarty, and J. C. Lance. 1981. Trace organic behavior
in soil columns during rapid infiltration of secondary wastewater.
Water Res. 15(1):151-160.
Chiou, C. T., V. H. Freed, D. W. Schmedding, and R. L. Kohnert. 1977.
Partition coefficient and bioaccumulation of selected organic chemicals
Environ. Sci. Techno!. 11:475-478.
Committee on Methods for Toxicity Tests with Aquatic Organisms. 1975.
methods for acute toxicity tests with fish, macroinvertebrates, and
amphibians. Ecol. Res. Series, EPA 600/3-75-009.
Dickson, A. G., and J. P. Riley. 1976. The distribution of short-chain
halogenated aliphatic hydrocarbons in some marine organisms. Marine
Pollut. Bull. 7(9):167-169.
Dilling, W. L., N. B. Tefertiller, and G. J. Kallos. 1975 Evaporation
rates and reactivities of methylene chloride, chloroform, 1,1,1-tri-
chloroethane, trichloroethylene, tetrachloroethylene, and other chlori-
nated compounds in dilute aqueous solutions. Environ. Sci. Technol.
9(a):833-838.
Neely, W. B., D. R. Branson, and G. E. Blau. 1974. Partition coefficient
to measure bioconcentration potential of organic chemicals in fish.
Environ. Sci. Technol. 8:1113.
Pearson, C. R., and G. McConnell. 1975. Chlorinated Ca and C2 hydrocarbons
in the marine environment. Proc. Roy. Soc. London B 189:305-332.
Schwarzenbach, R. P. and J. Westall. 1981. Transport of nonpolar organic
compounds from surface water to groundwater. Laboratory sorption
studies. Environ. Sci. Technol. 15:1360-1367.
Smith, J. H., D. C. Bomberger, Jr., and D. L. Haynes. 1980. Prediction of
the volatilization rates of high-volatility chemicals from natural
water bodies. Environ. Sci. Technol. 14(11):1332-1337.
U.S. Environmental Protection Agency. 1978. In-depth studies on health and
environmental impacts of selected water pollutants. Contract No.
68-01-4646, Duluth, MN.
U.S. Environmental Protection Agency. 1979. Tetrachloroethylene: water
quality criteria. Federal Register 44(52):15966-15969.
4-16
-------
U.S. Environmental Protection Agency. 1980. Tetrachloroethylene: ambient
water quality criteria. Office of Water Regulations and Standards,
EPA 440/5-80-073, October.
U.S. Environmental Protection Agency. 1981 Environmental risk assessment
of tetrachloroethylene, draft report. Office of Toxic Substances. 14
September.
Veith, G. D., D. L. DeFoe, and B. V.
ing the bioconcentration factor
Board Canada 36:1040-0145.
Bergstedt. 1979. Measuring and estimat-
of chemicals in fish. J. Fish. Res.
Zoeteman, B. C. J., K. Harmsen, J. B. H. J. Linders, C. F. H. Morra, and W.
Slooff. 1980. Persistent organic pollutants in river water and ground
water of the Netherlands. Chemosphere 9:231-249.
4-17
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5. MAMMALIAN METABOLISM AND PHARMACOKINETICS
The metabolism and pharmacokinetics of PCE are highly contingent on the
physicochemical properties of this compound. It is a volatile liquid at room
temperature with a relatively low vapor pressure (19 torr at 25°C), is nearly
insoluble in water but is highly lipophilic. Table 5-1 compares the physical
properties of PCE with other familiar chloro-substituted ethylenes. It is of
interest to note that with increasing chloro-substitution, water solubility
decreases and lipid solubility increases. These properties determine to a
considerable extent the absorption, distribution, and routes of elimination of
PCE from the body, i.e., the pharmacokinetics of PCE. All chloroethylenes are
thought to share as common metabolic steps in the body a) epoxidation of the
ethylene double bond, b) rearrangement to the halogenated acetaldehyde or acyl
halide, and c) partial transformation of the latter to a halogenated acetv
acid (Bolt et al., 1982).
5.1 ABSORPTION AND DISTRIBUTION
5.1.1 Dermal Absorption
Studies have shown that absorption of PCE through the skin, from vapor
exposure or from partial body immersion, is minimal in comparison to oral and
inhalation routes of exposure. Riihimaki and Pfaffli (1978) exposed human
volunteers (lightly clad and with respirators to prevent pulmonary absorption)
to 600 ppm (4068 mg/m ) PCE for 3.5 hours and^estimated percutaneous absorption
of the vapor to be about 1 percent of pulmonary absorption. Stewart and Dodd
(1964) and Hake and Stewart (1977) have also experimentally estimated skin
absorption in volunteers by noting PCE appearance and concentration in exhaled
air after immersion of the thumbs in liquid PCE. The mean peak concentration
3
of PCE in exhaled air 40 minutes after immersion was only 0.31 ppm (2.1 mg/m ),
3
and after 2 hours, 0.23 ppm (1.6 mg/m ), indicating that the rate of absorption
through the skin is very slow, even after allowing for storage and metabolism.
Jakobsen et al. (1982) experimentally quantified transport of PCE and other
chlorinated hydrocarbons across guinea pig skin. Liquid contact (skin area,
p
3.1 cm ) was maintained for up to 6 hours and solvent concentration monitored in
5-1
-------
TABLE 5-1. PHYSICAL PROPERTIES OF PCE
AND OTHER CHLOROETHYLENES
Vapor
Pressure at
25°C, 760 torr
Ostwald Solubility, 37°C
Water/ Blood/ Olive oil/
air air air
Vinyl chloride
(1-chloroethylene)
Vinylidene chloride
(1,1 dichloroethylene)
Trans,1,2 dichloroethylene
Cis 1,2 dichloroethylene
Trichloroethylene
(1,1,2 trichloroethylene)
Tetrachloroethylene
(1,1,2,2 tetrachloroethylene)
gas
600
350
250
90
19
2.1
2.9.
1.3
5.8
9.2
9.5
0.43 13.1
69
189
270
718
1917
Adapted from Sata and Nakajima, 1979.
Conversion factor: 1 ppm vapor in air = 6.78 mg/m3 at 25°C, 760 torr.
blood during and following dermal applications. Blood concentrations (reflect-
ing absorption rate) increased rapidly, peaking at 0.5 hour (1.1 ug/mL) and
then decreased with time, suggesting that changes in dermal permeability occur
with time. This pattern was found to be general and characteristic of chlori-
nated hydrocarbon solvents like PCE. Similar observations have been made by
Tsuruta (1975), who estimated the percutaneous absorption rate of PCE through
2
mouse skin at 24 nmol/min/cm of skin. Kronevi et al. (1981) have observed
histological changes in guinea pig derma within 15 minutes of topical exposure
to PCE. Stewart and Dodd (1964), Hake and Stewart (1977), Morgan (1969), and
Ling and Lindsay (1975) have reported severe and prompt erythema after dermal
contact with PCE in man.
5.1.2 Oral Absorption
PCE is rapidly and virtually completely absorbed into the body from the
gastrointestinal tract, presumably because of its high lipid solubility. Pegg
et al. (1979) and Schumann et al. (1980) have administered 14C-PCE, 1 and
5-2
-------
500 mg/kg (in corn oil) to both rats and mice and recovered better than 90
percent of the radioactivity in urine, feces, exhaled air, etc. For the rat,
peak blood concentration occurred within 1 hour or oral dosing (Figure 5-1).
Similar results have been found by Daniel (1963) after oral dosing of Cl-PCE
to rats (1.1 to 6.1 g/kg). Frantz and Watanabe (1983) found virtually complete
14
absorption of C-PCE in drinking water (saturated; approximately 150 ppm w/v;
150 ug/mL) by rats ingesting ad 1 ibitum over a 12-hr period with an average
consumption equivalent to an 8 mg/kg dose.
5.1.3 Pulmonary Absorption
5.1.3.1 Man. PCE in vapor form in air is readily absorbed through the lungs
into blood by first-order diffusion processes. Pulmonary uptake of a volatile
compound like PCE during inhalation exposure is largely determined by the
ventilation rate (about 4 to 8 L/min for man at rest), duration of exposure at
a given air concentration, solubility in blood and other body tissues, and its
metabolism. When body tissue concentrations (body burden) are at steady-state
with inspired air concentration, the rate of uptake is equal to the rate of
metabolism plus nonpulmonary excretion of PCE (i.e., by definition uptake per
unit time must equal excretion by all routes at steady state). Since there
are no known significant routes of excretion of PCE except pulmonary and
metabolism (Section 5.2), the steady-state uptake rate approximates the meta-
bolism rate.
The blood/gas (air) partition coefficient (A. b/g) at 37°C expresses the
solubility of a solvent such as PCE as the ratio of the concentrations in
blood (mg/L) and in air (mg/L). When there is no impairment of pulmonary
diffusion, circulation, or ventilation, this partition coefficient exists
between the arterial blood concentration (C.) and the inspired air concentra-
tion (C ) (or exhaled air concentration postexposure), so that C./C = A b/g.
cl Ma
As illustrated in Figure 5-2, blood concentration parallels air concentration
in a fixed ratio. The blood/gas partition coefficient for PCE is about 13 to
15. It is considerably higher than other solvents such as trichloroethylene
(TCI), 9.5 and methylchloroform (MC), 5 (Monster, 1979; Sato and Nakajima,
1979; Table 5-1). Thus PCE has the potential for a greater pulmonary uptake
rate than TCI or MC.
5-3
-------
100
i10
UJ
O
0.1
• 600 ppm, 6 hr
• 500 mo/kg GAVAGE
10
20
TIME, hours
30
40
Figure 5-1.
First-order excretion curves for PCE in blood of rats after
exposure to 600 ppm for 6 hr (—) or to 500 mg/kg gavage doses
(—). The animals were adult male Sprague-Dawley rats weighing
approximately 250 g. The body burdens of PCE were about 77.5 mg
per animal from inhalation exposure and 123.2 mg per animal after
oral dosage. (From Pegg et al., 1979).
The rate of uptake of PCE by body tissues from blood for a given concen-
tration in inhaled air is determined by tissue volume, the relative solubili-
ties in blood and tissue (tissue/blood partition coefficients), and by the
tissue blood flow. In accordance with the physiological model for PCE of
Guberan and Fernandez (1974), uptake by body tissues is a blood flow-dependent
process which can be grouped into four physiological compartments for conven-
ience in analysis: the blood vessel-rich group of tissues (VRG) (corresponding
to brain, heart, hepatoportal system, kidneys, and endocrine glands), the
muscle group (MG) (muscle and skin), the fat group (FG) (adipose tissue and
yellow marrow), and the vessel-poor group (VPG), composed of connective tissue
(bone, cartilage, etc.). The predicted uptake and distribution of PCE to
these tissue groups during and after 8-hr exposure to 100 ppm (678 mg/m3) is
shown in Figure 5-3. For an exposure of this magnitude, body burden is calcu-
lated as about 1000 mg PCE for a 70-kg man, with half this amount distributed
into fatty tissue.
With prolonged inhalation exposure to PCE, when a whole-body steady-state
condition is reached with inspired air concentration, the concentration of PCE
5-4
-------
= _ 1000
§ .5
o
0.
100
10
0.1
0.01
i A
i 1 1 | I I I 1 | I T I I
O 72 ppm PCE AT REST
A 144 ppm PCE AT REST
D 142 ppm PCE AT REST AND WORKLOAD
EXHALED AIR
50 100
POST EXPOSURE, hours
150
Figure 5-2. PCE concentrations in blood and exhaled air following inhalation
exposure for 4 hr (means of 6 subjects). From Monster et al.,
1979.
5-5
-------
1000
4 8
• EXPOSURE -
4 8 12
POST EXPOSURE «-
TIME, hr
20
24
26
Figure 5-3. Predicted uptake and distribution of PCE to tissue groups during
and after an 8-hr exposure to 100 ppm. See text for details.
From Guberan and Fernandez, 1974.
5-6
-------
in the individual tissue groups relative to blood is determined by the tissue/
blood partition coefficients. For highly 1ipid-soluble solvents such as PCE,
the adipose tissue has the highest partition coefficient and hence the highest
tissue concentration at steady state. For the three solvents, PCE, TCI, and
MC, the adipose tissue/blood partition coefficients at 37°C are about 105, 75,
and 50, respectively (Monster, 1979; Sato and Nakajima, 1979; Table 5-1). The
capacity of a tissue to take up PCE is the product of volume and solubility.
Since the volume of adipose tissue in a 70-kg man is about 10L, then adipose
tissue capacity for PCE is about 1500 mg; but because of the small rate of
perfusion of adipose tissue (about 0.4L blood/min; Eger, 1963), the time
needed to equilibrate the adipose tissue is large in comparison to that of
other tissues. Figure 5-3 indicates that a steady-state plateau concentration
is not achieved for adipose tissue within the period of an 8-hr exposure to
PCE, although it is for muscle tissue and other vessel-rich tissues. For a
given concentration in blood (or air), the half-time (TjJ (i.e., the time
necessary to equilibrate the adipose tissue to 50 percent of its final con-
centration) is about 25 hours for PCE (Monster, 1979; Fernandez et al., 1976).
This means that during a single 8-hr exposure, adipose tissue does not reach
steady-state equilibrium, which requires about five times TI , or about 125
^
hours.
At whole-body steady-state condition, the net pulmonary uptake of PCE is
balanced with excretion of PCE by other routes, principally by metabolism. The
difference between the inspired air concentration and end alveolar air concen-
tration (exhaled air concentration) provides a measure of PCE uptake, so that:
Q = (Cins - Caly) V • T (5-1)
where Q is the amount absorbed during time T (min), C is the air concentration
in mg/L, and V is the ventilation rate in L/min. The percent retention (R) is
defined as:
D _ (Cins - Ca1v) (5-2)
R - c:
ins
and therefore
Q = R (C. • V • T)/100 (5-3)
5-7
-------
Figure 5-4 shows the change in end alveolar concentration during exposure of
5 subjects for 8 hours at 100 ppm (678 mg/m ) (Fernandez et al., 1976). The
data show a rapid rise in alveolar concentration during the first hour of expo-
sure, followed by a slower but continuous rise during the subsequent 7 hours of
exposure (retention value decreasing with exposure) towards a fixed retention
value at whole-body equilibrium, although it is apparent from Figure 5-4 that
equilibrium or steady-state retention is not reached during an 8-hr exposure.
Thus, rate of uptake, at first rapid, diminishes with continuing exposure to a
constant rate at steady state. Retention values of 60 to 80 percent have been
reported at the end of short exposures (4 to 6 hours), and lower retention
values would be expected with longer exposures (Fernandez et al., 1976; Monster
et al., 1979; Bolanowska and Golacka, 1972). These values indicate pulmonary
uptake of PCE is rather high. Since the difference in retention, 20 to 40
percent, represents excretion by other than pulmonary routes (principally
metabolism), these retention values suggest that metabolism of PCE may be
considerably greater than indicated by other measures of metabolism (Section
5.2.2).
Monster et al. (1979) have subjected volunteers to controlled inhalation
exposures for 4 hours and determined experimentally PCE uptake kinetics.
Monster et al. found that net pulmonary uptake decreased with exposure duration,
with about 25 percent less PCE absorbed in the fourth hour as compared with the
first hour of exposure. Table 5-2 gives the amounts (mg) PCE uptake at 72 and
144 ppm (488 and 976 mg/m ) exposure for 4 hours. For 6 subjects, the average
uptake was 455 mg and 945 mg for the two exposure concentrations, respectively,
showing a direct proportionality of uptake with inspired air concentration at
least up to 144 ppm (976 mg/m ). Their data (Table 5-2) also demonstrate the
marked increase in uptake rate (and total uptake) as a result of an increase
in ventilation rate provoked by exercise. Interperson variations in PCE
uptake were also influenced by total body mass and adipose tissue mass.
Fernandez and co-workers (Guberan and Fernandez, 1974; Fernandez et al.,
1979) and Stewart and co-workers (Hake and Stewart, 1977; Stewart et al. ,
1961, 1970, 1974) have also conducted controlled exposures with human volun-
teers. Their results are in accord with those of Monster and co-investigators.
Hake and Stewart (1977) exposed five subjects for 5 consecutive days to 100 ppm
(678 mg/m ) PCE for 7 hours daily. As shown in Figure 5-5, the concentration
of PCE in exhaled breath of the 5 subjects increased as the 5-day week pro-
gressed, indicating an increase in the body burden with repeated daily exposure. , , ,
5-8
-------
60
50
a
a
2 40
<
cc
LU
g 30
O
o
cc
o
LU
^ 20
<
10
V
4 5
EXPOSURE T!ME,hr
Figure 5-4. PCE alveolar air concentration during exposure of 5 subjects for
8 hr at 100 ppm. From Fernandez et al., 1976.
5-9
-------
TABLE 5-2. ESTIMATED UPTAKE OF SIX INDIVIDUALS EXPOSED TO PCE AT REST
AND AFTER EXERCISE
Subject
A
B
C
D
E
F
Average
Body
Mass
(kg)
70
82
82
86
67
77
77.3
Lean
body
mass
(kg)
62
71
71
74
61
61
66.7
Ventilation
Min volume
at rest
(L/m)
7.6
11.6
10.0
11.3
12.3
8.8
10.3
72 ppm
(at rest)
(mg)
370
490
530
500
390
450
455
Uptake*
144 ppm
(at rest)
(mg)
670
940
1000
1210
880
970
945
144 ppm
+ exercise
(mg)
1060
1500
1400
1510
1320
1120
1318
^Exposures for 4 hr; exercise two 30-min periods during 4-hr exposure.
Source: Monster et al. , 1979.
Fernandez et al. , using their mathematical model, predicted that chronic
occupational exposure to 100 ppm (678 mg/m3) PCE (8 hr/day, 5 day/wk) would
result in cumulation of PCE in adipose tissue. As uptake by fatty tissue
during the working hours of the week equals the elimination during nights and
weekends, an equilibrium is eventually established but which requires a time
period of 3 to 4 work weeks.
5.1.3.2 Rodents. Pegg et al. (1979) and Schumann et al. (1980) have deter-
mined the overall pulmonary uptake of rats and mice exposed to PCE by inhala-
tion for 6 hours. The animals were exposed in a 30-liter glass chamber to 10 or
600 ppm (67.8 or 4068 mg/m3) 14C-PCE. At termination of exposure (6 hours), the
cumulative pulmonary uptake was estimated by determining the radioactivity of
carcass, and radioactivity in expired air, urine, and feces. The data for
rats and mice are given in Tables 5-3 and 5-4, respectively. For mice, the
body burden resulting from a 10-ppm exposure was 0.40 mg/animal (16.5 rag/kg),
and for rats, 1.48 and 77.5 mg/animal (5.9 and 310 mg/kg) for 10 and 600 ppm
(67.8 or 4068 mg/m ), respectively. These data indicate that pulmonary uptake
5-10
-------
LJJ
O
CC
O
I
O
§: 100
z
O
10
Z
LLJ
O
Z
O
O
I I I I I I
7 hr VAPOR EXPOSURES
1 I
5 6 7 8
TIME, days
9 10 11 12 13 14
Figure 5-5. Mean exhaled breath concentrations of PCE for five volunteers
exposed to 100 ppm PCE for 7 hr per day for 5 days. Note the
rising concentration of PCE toward equilibrium with inspired air
concentration with daily exposure, and the long decay of PCE
concentration in expired air after the last exposure. (From
Hake and Stewart, 1977; Stewart et al., 1970).
5-11
-------
TABLE 5-3. DISPOSITION IN RATS OF 14OPCE RADIOACTIVITY 72 HR FOLLOWING
ORAL OR INHALATION ADMINISTRATION
Recovery, pmol equ. PCE
1 mg/kg*
500 mg/kg
Expired air
Unchanged PCE
Metabol ized
14C02
Uri ne
Feces
Carcass
1.05 ± 0.00
0.04 ± 0.01
0.24 ± 0.00
0.09 ± 0.02
0.05 ± 0.00
71.5
2.5
16.5
6.2
3.3
667.31 ± 22.72
3.44 ± 0.36
34.48 ± 2.40
29.08 ± 4.01
8.50 ± 0.66
89.9
0.5
4.6
3.9
1.2
Total
Expired air
Unchanged PCE
1.46 ± 0.02
10 ppm, 6 hr*
6.08 ± 0.44
68.1
742.79 ± 27.75
600 ppm, 6 hr
412.38 + 7.60
88.0
Metabol ized
14C02
Urine
Feces
Carcass
Total
0.
1.
0.
0.
8.
32 ± 0.
66 ± 0.
46 ± 0.
38 ± 0.
91 ± 0.
05
12
02
06
33
3.
18.
5.
4.
6
7
2
3
3.
27.
14.
10.
467.
25
40
24
07
05
± 0.
± 1.
± 0.
± 0.
± 8.
17
68
75
54
11
0.
6.
3.
2.
7
0
1
2
*Mean ± SE for 3 rats; gavage doses in corn oil in volume of 1.0 mL/kg;
Sprague-Dawley rats.
Source: Pegg et al. , 1979.
of PCE by the rat at least is approximately proportional to exposure concentra-
tion.
On a comparative species basis, the pulmonary uptake (comparing 10 ppm,
6-hr exposures) for the rat and mouse more closely relates to the ratio of
2/3 ?/^
their surface areas (b.w. rat /b.w. mouse ' = 4.7) than to their relative
body weights (ratio 10.2). Recently, Landry et al. (1983) have developed
methods to measure simultaneously respiratory frequency, tidal volume, minute
volume, and net uptake of an inhaled vapor in rats. During steady state, if
metabolism is the only significant route of elimination, then net uptake rate
of the inhaled vapor is equal to the rate of metabolism. This new approach
5-12
-------
TABLE 5-4. DISPOSITION IN MICE OF 14OPCE RADIOACTIVITY 72 HR FOLLOWING
ORAL OR INHALATION ADMINISTRATION
Expired air
Unchanged PCE
Metabol ized
14C02
Urine
Feces
Carcass
Cagewash
Total
*
10 ppm, 6 hr.
0.29 ± 0.06
0.19 ± 0.01
1.53 ± 0.20
0.16 ± 0.04
0.07 ± 0.01
0.19 ± 0.11
2.44 ± 0.27
Recovery
%
12.0
7.9
62.5
6.8
3.0
7.7
, pmol equ. PCE
500 mg/kg
53.69 ± 4.96
0.85 ± 0.67
6.58 ± 3.19
0.80 ± 0.05
0.32 ± 0.07
2.67 ± 1.53
64.92 ± 3.09
%
82.6
1.3
10.3
1.2
0.5
4.1
*Mean ± SD for 3 rats; gavage doses in corn oil; BgC^f^ mice.
Source: Schumann et al., 1980.
appears to be useful for inhalation dosimetry and evaluation of metabolic
rates in rats.
5.1.4 Tissue Distribution and Concentrations
PCE is expected to distribute, by first-order diffusion processes, into
all body tissues, as is known for other chloroethylene compounds with lower
lipid solubility (Table 5-1; Waters et al., 1977). PCE readily crosses the
blood-brain barrier and the placental barrier. The partition coefficients for
various tissues, relative to blood or air, have not been firmly established,
although approximations can be calculated from the rat tissue distribution
data of Savolainen et al. (1977) (Table 5-5); rat tissue/blood: adipose, 60;
liver, 5; brain, 4. Absolute tissue concentrations are directly proportional
to the body burden or exposure dose.
5.1.4.1 Rodents. Savolainen et al. (1977) analyzed tissue levels of PCE
3
after exposure of rats 6 hours daily for 5 days to 200 ppm (1356 mg/m ) PCE in
air. Values are shown in Table 5-5. Seventeen hours after exposure on day
four, blood and other tissues still contained significant concentrations of
PCE with highest concentrations in adipose tissue (103 pg/g) reflecting the
high solubility of PCE in this tissue and its long storage or half-time of
5-13
-------
TABLE 5-5 RAT ORGAN CONTENT OF PCE AFTER DAILY INHALATION EXPOSURE
OF 200 PPM FOR 6 HOURS PER DAY. MEASUREMENTS MADE ON FIFTH DAY.
Time (hr of
exposure
on fifth day)
0
(17 hr after
day 4
exposure)
2
3
4
6
Cerebel
18.4
±3.7
89.7
+16.7
108.7
±4.9
101.5
±17.2
142.5
±0.2
1 urn Cerebrum
nmol/g (mean of
13.1
±4.4
62.0
±7.1
72.0
±4.5
68.1
±5.5
92.3
±1.3
Lungs
Liver
two determinations
9.5
±1.8
45.6
±10.1
52.4
+9.8
59.5
±13.4
74.0
+3.6
35.2
±9.3
107.4
±27.3
134.9
±1.0
133.8
±0.7
160.7
±24.2
Peri renal
Fat
+ range)
622.2
±20.0
977.6
±174.0
807.1
±38.0
1105.3
±109.7
1724.8
±422.4
Blood
4.3
±1.1
21.3
±4.0
25.3
±1.5
24.4
±3.4
30.3
+6.6
Source: Savolainen et al., 1977.
desaturation. With exposure on day five, tissue cumulation occurred in brain,
lungs, liver, fat, and blood, and steady-state level was not achieved by the
end of the 6-hr exposure.
5.1.4.2 Man. Except for blood concentrations (Figure 5-1). no direct analy-
tical information of tissue levels after various exposure concentrations of
PCE is available. As for the rat, tissue concentrations are expected to be
proportional to exposure concentrations and to duration of exposure. Guberan
and Fernandez (1974) developed a mathematical model of PCE uptake, distribution,
and excretion (based on experimental data). Their physiological model predic-
ted the distribution of PCE in the three major physiological compartmental
tissue groups--vessel-rich group (VRG), muscle group (MG), and adipose tissue
(FG)--during and after an 8-hr inhalation exposure of 100 ppm. Figure 5-2
shows that by the end of exposure, 50 percent of the solvent taken up by the
body has been distributed to the fatty tissue, due to its high tissue-blood
partition coefficient. The distribution of PCE to the other tissue groups -is
5-14
-------
related to their volume and partition coefficient and therefore is higher for
the MG than for the VRG and the vessel-poor group (VPG). After exposure, the
depletion of these three groups of tissues is almost complete after about
20 hours, and the continuing elimination in alveolar air is then related to the
slow release of PCE from adipose tissue. Beyond about 20 hours of elimination,
the rate of discharge of PCE from the adipose tissue, and therefore from the
whole body, is an exponential function of time with a predicted half-life of
about 71 hours. Because of this long residence of PCE in adipose tissue,
repeated daily exposure (6 hr/day, 100 ppm) results in an accumulated concen-
tration, as PCE from new exposures adds to residual concentration from previous
exposures, until steady state is reached. Adipose tissue levels of PCE reach
a steady-state "plateau" (dependent on inhalation concentration and daily
exposure duration) and thereafter remain relatively constant. For a 6-hr dc
100 ppm (678 mg/m ) exposure, Guberan and Fernandez (1974) predicted that
plateau concentrations in adipose tissue occur about 22 to 29 days after
initial daily exposure.
5.2 EXCRETION
The total elimination of an absorbed body burden of PCE involves two
major processes—pulmonary excretion of unchanged PCE and metabolism to urinary
metabolites. The metabolism of PCE appears to be very limited in man and
animals and is reviewed extensively in Section 5.4. PCE or its metabolites
have not been reported to be excreted by other routes than lung and urine in
any significant amounts. Bolanowski and Golacka (1972) have reported that the
rate of PCE elimination through the skin in man is about 125 ug/hr during a
6-hr exposure to 390 ppm (694 mg/m ), or about 0.02 percent of the uptake/hr.
Thus, pulmonary elimination is the major route of PCE excretion.
In rodents, the principal metabolites of PCE are trichloroacetic acid
(TCA) and oxalic acid. In man, TCA is the predominant metabolite although
oxalic acid may be unrecognized.
5.2.1 Pulmonary Elimination in Man
Pulmonary elimination of unchanged PCE after exposure occurs as a first-
order diffusion process across the lungs from blood to alveolar air, depicted
in a manner as an inverted reproduction of its uptake as illustrated in
5-15
-------
Figure 5-3. From controlled experimental exposures in humans, Stewart and co-
workers (Stewart et al., 1961, 1970, 1974; Hake and Stewart, 1977) have followed
the body desaturation of PCE after exposure by serial breath analysis of
alveolar air. Figure 5-5 shows the results following exposure of 5 subjects
for 5 consecutive days to 100 ppm (678 mg/m3) PCE for 7 hours daily. At least
two first-order phases of pulmonary elimination of PCE are apparent: an
initial fast phase, followed by a slow predominant excretion phase with a
half-time of about 65 hours. Monster et al. (1979) have also studied the
kinetics of pulmonary elimination of PCE in experimental human exposures and
determined from the concentration curves of PCE in exhaled air and blood
(Figure 5-2) that the compound was eliminated from the body through the lungs
at three different first-order rate constants with corresponding half-lives of
12 to 16 hours, 30 to 40 hours, and 55 to 65 hours, in accordance with the
desaturation of three major body compartments, represented by VRG, MG, and FG
compartments, respectively (Monster et al. , 1979; Guberan and Fernandez,
1974). The long half-time (55 to 65 hours) of elimination of PCE from adipose
tissue (FG), which is due to the high adipose tissue/blood partition coefficient
of PCE and the low rate of blood perfusion of the tissue (Eger, 1963), is inde-
pendent of the body burden of PCE as shown by the parallel blood and exhaled air
concentration decay curves of Figure 5-2 for exposure concentrations of 72 and
144 ppm (488 and 976 mg/m3).
However, the exhaled air (or end alveolar air) concentrations and blood
concentrations of PCE after exposure and throughout desaturation are propor-
tional to the acquired body burden or exposure concentration and duration, and
serve as a means for estimating body burdens, as illustrated in Figures 5-2
and 5-6 (see Section 5-3 for discussion). Guberan and Fernandez (1974) have
developed a physiological kinetic model for PCE excretion in man, based on
experimental findings from controlled exposures of volunteers. The alveolar
concentrations during and after exposure in the excretory phase were predicted
for use in estimating mean exposure in occupational and environmental situa-
tions. The predicted half-life of PCE stores in adipose tissue was 71.5
hours, in good agreement with the experimental observations of Stewart et al.
and of Monster et al. The long half-time of pulmonary excretion of PCE shows
that a long period is necessary to completely eliminate PCE, i.e., above five
times the half-life or about 2 weeks.
5-16
-------
300
DAY
Figure 5-6.
Daily (8-hr) occupational inhalation exposure to PCE. Course of
the relative (first Monday morning concentration taken as 100
percent) concentration of trichloroacetic acid (TCA) in blood and
urine (mean of 23 subjects ± S.D.) during the work week. c
Monster et al., 1983.
From
5-17
-------
5.2.2 Urinary Metabolite Excretion in Man
In contrast to first-order pulmonary excretion of PCE, the urinary excre-
tion of TCA (or total trichloro-compounds by Fujiwara reaction) from metabolism
of PCE is dose-dependent, and presumably follows Michaelis-Menten kinetics
[See Section 5.4.3 for discussion]. Most studies have quantitated PCE urinary
metabolite excretion by the Fujiwara reaction for which there is a question of
the true nature of the compound(s) measured, although it is accepted that TCA
is the principal compound. Thus Ikeda and Imamura (1973) found the mean T^ of
total urinary trichloro-compounds for 13 subjects occupationally exposed to
PCE to be 144 hours (range 123 to 190 hours). In contrast, when TCA is directly
administered to men, the half-life has been found to be 51 to 82 hours (Paykoc
and Powell, 1945; Muller et al., 1974). Hence, the longer T^ of TCA from PCE
metabolism is no doubt due to constant metabolic formation of TCA from PCE
cycling to the liver over the period of its long residence time in adipose
tissue (Tj , 17 to 55 hours).
'i
Monster and co-workers (Monster et al. , 1979; Monster and Houtkooper,
1979; Monster et al. , 1983), using gas chromatographic methods for analysis,
3
exposed volunteers to 72 and 144 ppm (488 and 976 mg/m ) PCE for 4 hours and
determined TCA in blood and urine. They found that urinary TCA represented
less than 1 percent of the estimated absorbed dose of PCE. In blood, following
exposure, TCA continued to increase in concentration until about 20 hours
(representing continued metabolism of PCE in the body to TCA) and then declined
exponentially as a first-order elimination process (Figure 5-7). The blood
concentration of TCA was proportional to exposure concentration (i.e., body
burden of PCE) and declined with a half-life of about 65 to 90 hours. Figure
5-8 shows that urinary cumulation paralleled blood TCA disappearance, and also
that the urinary TCA cumulation was proportional to the inhalation exposure
concentration of PCE (and hence body burden). Physical activity increases PCE
pulmonary uptake (body burden), and hence, increased TCA formation and urinary
excretion. However, more recently Ohtsuki et al. (1983) correlated inhalation
exposure concentration with urinary total trichloro-compounds and found that
urinary metabolite excretion was not a linear function of exposure concentra-
tion but instead a hyperbolic function approaching a "plateau" at about 400
ppm exposure (Figure 5-9). Since urinary metabolite(s) excretion reflects PCE
metabolism, these observations indicate that PCE metabolism is dose-dependent
and saturable with high PCE body burdens or exposure.
5-18
-------
O)
E
o
o
O
O
_l
03
0.6 —
0.4 —
0.2 —
0.1 —
£144 ppm PCE AT REST
D142 ppm PCE AT REST AND WORKLOAD
I I I I I I I I I I I I I I I I I I
0 50 100 150
TIME AFTER EXPOSURE, hours
Figure 5-7. Trichloroacetic acid blood concentrations following inhalation
exposure to PCE for 4 hr (means of 6 subjects). From Monster
et al., 1979.
k.
QJ
•a 10
< 8
£ 6
LLJ
Z 4
i 2
n
I I I I
72 ppm PCE , 1
AT REST I
I
— I
:I^i
44 p
AT
rh
1 1
pm PCE
REST
¥1
*
! 1
142 ppm PCE
AT REST AND
WORKLOAD
"MT -
0 22 46 70 0 22 46 70 0 22 46 70
TIME AFTER START EXPOSURE, hours
Figure 5-8. Urinary excretion of trichloroacetic acid during and following
inhalation exposure to PCE for 4 hr (means of 6 subjects).
From Monster et al. , 1979.
5-19
-------
150
100
200
300 400 500 600 700
PCE CONCENTRATION IN AIR, ppm
800
900
Figure 5-9. Relationship between PCE occupational inhalation exposure (time-
weighted ppm average for an 8-hr work shift) and urinary concen-
tration of total trichloro-compounds at end of work shift (correc-
ted to a urine specific gravity of 1.016) for 36 male (squares)
and 25 female workers (circles). From Ohtsuji et al. , 1983.
5.2.3 Chronic Exposure
The long half-time of elimination of PCE by the pulmonary route (~ 65 hr)
and by urinary metabolite(s) (^ 65 to 144 hr) indicates that, with repetitive
daily or chronic inhalation exposure to PCE, sustained or "plateau" levels of
PCE and TCA in blood and of TCA in urine would occur as body burden, metabolism,
and excretion approach a steady state with pulmonary uptake. Monster et al.
(1983) determined blood levels of PCE and TCA and urine TCA concentrations
during occupational exposure (8 to 50 ppm, 8 hours daily) during the course of
the work week. Figures 5-6 and 5-10 show their data expressed in terms relative
to Monday morning concentrations (control; 100 percent) occurring from previous
work exposure for 23 male and female workers. Figure 5-10 shows that exhaled
air concentration (a measure of body burden) and blood concentration (directly
proportional to exhaled air concentration and also a measure of body burden)
both rose 4-fold with the 3-day exposures of Monday, Tuesday, and Wednesday,
and tended to "plateau" with Thursday and Friday exposures. These "plateau"
levels declined to Monday morning levels (100 percent) during exposure-free
days, Saturday and Sunday. Figure 5-6 shows that the concentrations of the
metabolite TCA increased during the work week in blood as well as in urine.
The relative concentrations on Wednesday and Friday after work were higher by
a factor of two than on Monday morning. The increase of the concentration in
5-20
-------
800
MON TUES WED THUR FRI
DAY
SAT SUN
MON
Figure 5-10. Daily (8-hr) occupational inhalation exposure to PCE. Course of
the relative concentration (first Monday morning concentration
taken as 100 percent) of PCE in blood and exhaled air during the
work week (mean of 23 subjects ± S.D.). From Monster et al. ,
1983.
urine was more pronounced than that in blood. At the start of the following
week the relative concentrations had returned to about 100 percent. Thus,
during daily exposure during the work week, the body burden and blood levels
of PCE increased with each exposure and then tended to plateau towards the end
of the week, but the metabolite TCA continued to accumulate in blood and urine
apparently as a result of the longer half-life of TCA. Similar experimental
observations have been reported by Tada and Nakaaki (1969), and predicted from
the mathematical model of Guberan and Fernandez (1974).
5.2.4 Excretion Kinetics in the Rodent
As is the case for man, PCE is excreted from the rodent (rat and mouse)
principally by two routes: pulmonary elimination of unchanged PCE and by
metabolism and renal elimination of PCE metabolites. For the rat, Pegg et al.
(1979) and Frantz and Watanabe (1983) found that the rate of PCE excretion in
expired air is described by apparent first-order kinetics. No significant dif-
ference was observed in the elimination half-time (approximately 7 hours) with
5-21
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either dose or route of administration (inhalation, gavage, or drinking water).
Figure 5-1 shows the disappearance of PCE from whole blood which follows
apparent first-order kinetics for up to 36 hours with elimination rate constant
of 0.10 to 0.12 hr"1.
The percentage of the total body burden excreted by the pulmonary route
in the rat and mouse depends on the dose. As calculated from the data of Pegg
et al. (1979) and Schumann et al. (1980) (Tables 5-3 and 5-4), the body burdens
of PCE in rats after a 6-hr inhalation exposure to 10 and 600 ppm (68 and
4068 mg/m3) are 1.48 and 77 5 mg per animal, respectively; the percentages of
these body burdens excreted by the pulmonary route as unchanged PCE were 68
3
and 99 percent, respectively. For the mouse exposed to 10 ppm (68 mg/m ) PCE
for 6 hours, the body burden was 0.40 mg and the pulmonary excretion only 12
percent, whereas for a body burden of 10.8 mg from oral administration, 83
percent was excreted by the pulmonary route. Similar results can be calculated
for body burdens from gavage and drinking water administration from the data
of Tables 5-3, 5-4, and 5-6 (Pegg et al., 1979; Schumann et al. , 1980; Frantz
and Watanabe, 1983). Hence, as the body burden of PCE is increased in the rat
or mouse, the percentage excreted unchanged increases. Conversely, as metabo-
lism is the other principal route of elimination of PCE, when the body burden
increases, the percentage of the burden metabolized (urinary metabolites)
decreases (although the absolute amount increases) (Tables 5-3, 5-4, and 5-6).
These observations suggest that metabolism of PCE and urinary excretion of
metabolites in the rodent are rate-limited and dose-dependent, following
Michael is-Menten kinetics (whereas pulmonary excretion is a first-'order process
and dose-independent, i.e., with half-time and rate constant independent of
dose).
Other investigators have found evidence of dose-dependent Michaelis-Menten
metabolism kinetics. Filser and Bolt (1979) exposed Wistar rats to PCE in air
in a closed system and determined the pharmacokinetics from disappearance
rate of PCE from the chamber (uptake by rat). In comparison to other halo-
genated ethylenes, they found PCE to be metabolized extremely slowly in rats.
Zero order Vmax (saturation) was <1.16 mg/hr/kg b.w.; for a 250-g rat, about
3.0 mg PCE/hr is metabolized and excreted as urinary TCA and other metabolites.
The principal urinary metabolites in the rat are oxalic acid and TCA (Table 5-7)
Buben and O'Flaherty (1984), using Swiss-Cox mice, found urinary metabolite(s)
(TCA only in these mice) excretion was limited by metabolic capacity. Mice
were chronically administered PCE by gavage at dose levels up to 2000 mg/kg/day.
5-22
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TABLE 5-6. DISPOSITION IN RATS OF 14OPCE RADIOACTIVITY 72 HR FOLLOWING
DRINKING WATER INGESTION*
Expired air
Unchanged PCE
Metabolized
14C02
Urine
Feces
Carcass
Cagewash
Total
Recovery, umol equ.
8.09 mg/kg*
12.47 ± 5.15**
0.29 ± 0.08
0.96 ± 0.26
0.23 ± 0.07
0.15 ± 0.04
0.02 ± 0.02
14.10 ± 5.51
PCE
%
87.9
2.2
7.2
1.7
0.9
0.2
*Rats consumed an average dose of 8.09 ± 3.13 mg/kg, which varied with the
volume of drinking water ingested over a 12-hr exposure period.
**Each value represents mean ± SD from four Sprague-Dawley rats,
Source: Frantz and Watanabe, 1983.
A plot of amount of urinary metabolite (TCA) per day versus dose level fitted
the Michaelis-Menten function showing the metabolism of PCE is capacity-1imi ted
even at relatively low doses (below 100-200 mg/kg) (Figure 5-11), with an
estimated V for urinary metabolite excretion of 136 mg/kg/day (5.7 mg/hr/kg
max
b.w. ) and a K of 660 mg/kg. These values suggest an excretion by metabolism
in mice greater than that estimated in the rat by Filser and Bolt (1979) by
other experimental means.
5.3 MEASURES OF EXPOSURE AND BODY BURDEN
The pharmacokinetic parameters of PCE uptake, metabolism, and pulmonary
excretion obtained from experimental human and animal exposures have been used
to determine reliable estimates of body burdens occasioned by occupational and
environmental exposures (Stewart et al., 1961, 1970, 1974; Boettner and Muranko,
1969; Ikeda et al., 1972; Essing et al., 1973; Ikeda and Imamura, 1973; Guberan
and Fernandez, 1974; Fernandez et al., 1976; May, 1976; Monster et al., 1979;
Monster and Houtkooper, 1979; Tada and Nakaaki, 1969; Ogata et al., 1971;
5-23
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TABLE 5-7. REPORTED METABOLITES OF PCE (OTHER THAN TCA)
Metabolite System Investigator
Inorganic chloride Rat urine Daniel, 1963
Carbon dioxide Rat, mouse Pegg et al., 1979
exhaled air Schumann et al., 1980
Oxalic acid Rat, mouse Yllner, 1961
urine Pegg et al. , 1979
Dmitrieva, 1967
Dichloroacetic acid Mouse urine Yllner, 1961
Ethylene glycol Rat urine Dmitrieva, 1967
Production from PCE-oxide Spontaneous Decomposition
in Aqueous and Non-aqueous Solution
Trichloroacetic acid Bonse et al., 1975
Trichloroacetyl chloride Kline et al., 1978
Verberk and Scheffers, 1980; Ziglio, 1981; Ohtsuki et al. , 1983; Monster
et al., 1983; Monster and Smolders, 1984). From these studies, three approaches
have been developed to estimate body burdens of PCE: 1) concentration of PCE
in exhaled air, 2) blood levels of PCE, and 3) concentrations of trichloroacetic
acid (TCA) in blood and urine. Stewart et al. (1961) pioneered the use of
expired air concentrations as a measure of body burden. In addition to the
advantage of a non-invasive methodology, alveolar PCE concentrations appear to
be a good index of the vapor exposure to which individuals have most recently
been subjected (Figure 5-12). Urine analysis for TCA excretion appears to be
of lesser value in estimating exposure to PCE. The urinary excretion of TCA
increases slowly and gains the maximal level only after 1 to 4 days, and
furthermore, the level is not clearly proportional to inhalation concentration
(Weiss, 1969; Tada and Nakaaki, 1969; Kundeg and Hogger, 1970; Ikeda et al.,
1972; Fernandez et al., 1976; Ohtsuji et al., 1983), owing to the long half-
life of PCE metabolites that appear in urine (144 hr, Ikeda, 1977) and to a
low metabolic rate which is dose-dependent and saturable (Section 5.4.3).
Monster and co-workers (Monster et al., 1983; Monster and Smolders, 1984) find
5-24
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120
200
500
1000
DOSE, mg/kg
1500
2000
Figure 5-11.
Relationship between PCE and dose and the amount of total urinary
metabolite excreted per day by mice. Data points are means of 9
to 11 mice ± SEM. The points fit the Michaelis-Menten equation
with an r2 value of 0.996. The calculated value from the curvi-
linear fit are V = 136 mg/kg/day and K = 660 mg/kg.
Buben and 0'FlaheTty, 1984. m
From
that for measuring environmental exposure to PCE, biological monitoring of
exhaled air is a "simple, efficient, effective and convenient method" of
assessing total ambient exposure of both young and aged subjects. For measuring
occupational exposure, they find that the best parameter to estimate the time-
weighted average exposure to PCE over the whole work week is the exhaled concen-
tration of PCE and TCA in blood at the end of the workday on Friday. The
second best parameter is measurement of blood PCE and exhaled air concentration.
These methods for estimating PCE body burdens after single or chronic
exposure are subject to high inter-individual variations that limit their
predictive value. Some of the factors known to contribute to these variabili-
ties are pulmonary dysfunctional states, individual differences in intrinsic
metabolic capacity for PCE, body mass and adipose tissue mass, modification of
metabolism by drugs and environmental xenobiotics, age, sex, and exercise or
workload. Guberan and Fernandez (1974), using a mathematical model developed
to predict uptake and distribution of PCE in the body and its elimination in
alveolar air, have computer-simulated the effects of age, body weight, height,
and body fat content, both at rest and during physical exercise.
5-25
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5000
1755-
1000
805—
to
-f 500
o
a.
cc
<
Q
LU
-i
<
I
X
IU
z
m 100
50
10
Ln PERC = -2.79+1.345 Ln TWA
R2 = 0.927
A= 1.43
n = 63
• WORKSHOP A
• WORKSHOP B
A WORKSHOP^
• WORKSHOP D
REGRESSION LINE
95% CONFIDENCE LIMIT
/" -,
* /* *
TT
/ ,
/
I I KM I
50
Figure 5-12.
100
500 1000 2000
2050
II
10,000
PCE IN WORKROOM AIR AS 4-HOUR TWA, /zmol/m3
Direct linear relationship between the time-weighted average
occupational exposure to PCE over the last 4 hr of a work day
and the concentration of PCE in exhaled air 15-30 rain after the
end of exposure (data for 32 subjects). Inhalation exposure
concentration of 2,050 umol/m3 is equivalent to 50 ppm. From
Monster et al., 1983.
5-26
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5.4 METABOLISM
5.4.1 Known Metabolites
Trichloroacetic acid (TCA) is the most prominent metabolite of PCE in
animals and man. The principal site of metabolism is the liver. Bonse et al.
(1975) have perfused the isolated rat liver with PCE and directly demonstrated
the capacity of this organ to metabolize PCE to TCA; however, no other meta-
bolite (e.g., oxalic acid) was found. Other tissues, such as lung and kidney,
may also metabolize PCE, as these tissues are known also to contain P-450
metabolizing systems. Nonetheless, other metabolites have also been described
(Table 5-7). The metabolism of PCE was first systematically investigated with
14
C-labelled compound some 25 years ago by Yllner (1961). Yllner established
the principal urinary metabolites of mice to be TCA (52 percent), oxalic acid
(11 percent) with traces of dichloroacetic acid, and the remaining radioactivity
(37 percent) in unknown form. More recently, Dmitrieva (1967) and Pegg et al.
(1979) have confirmed that oxalic acid is a major urinary metabolite of rats,
and, indeed, in this species, may be the principal metabolite. As yet, oxalic
acid has not been determined to be a metabolite of man, for whom TCA is pre-
sently assumed to be the principal metabolite. Buben and 0'Flaherty (1984)
observed no excess oxalic acid excreted by mice after chronic administration
of large oral doses of PCE, and these workers suggest that oxalic acid is not
a significant metabolite in this species.
The identification of trichloroethanol (TCE) in the urine of humans has
been described by some authors (Ogata et al. , 1962; Tanaka and Ikeda, 1968;
Ikeda and Ohtsuji, 1972; Ikeda et al., 1972). However, in all of these studies,
the method used for TCA metabolite quantitation (Fujiwara reaction) is based
on color production before and after oxidation with chromium trioxide and
addition of pyridine. The difference between the color production before
(TCA) and after oxidation is considered an estimate of TCE content. Sakamoto
(1974), from comparative urine analysis by GC and the Fujiwara reaction under
different pH and temperature conditions, expresses doubt that the entire
fraction detected by Fujiwara reaction in the oxidation fraction is truly TCE.
Nonetheless, Monster et al. (1983) and also Weicherd and Lindner (1975) identi-
fied by GC small amounts of TCE (<4 |jmol/mmol creatinine) in the urine of
persons exposed to 10 to 30 ppm (68 and 204 mg/m ) PCE in air, although others
using GC have not in controlled experimental exposures to pure PCE (Hake and
Stewart, 1977; Fernandez et al., 1976; Monster et al., 1979). For mice, Yllner
5-27
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(1961) reported that by his chromatographic method TCE was not detected in
urine as a PCE metabolite. Daniel (1963) reported similarly for the rat,
-using steam distillation combined with isotopic dilution methodology.
Buben and 0'Flaherty (1984) also were unable to find evidence of TCE in
the urine of chronically dosed mice as analyzed by gas chromatography; TCA was
the only metabolite found. Furthermore, Costa and Ivanetich (1980) could not
detect TCE as a product of PCE metabolism by pre-induced rat liver microsomal
preparations i_n vitro.
In mammalian systems, oxalic acid is known to be converted in part to
carbon dioxide, which has been identified in exhaled air of the rat and mouse
after exposure to 14C-labelled PCE (Table 5-4). Dmitrieva (1967) has also
reported substantial amounts of ethylene glycol in urine of rats exposed to
PCE, although this observation has not been confirmed by others (Buben and
0'Flaherty, 1984). As no complete balance study of end-metabolites of PCE has
been reported for the rodent or man, it is likely that all metabolites of PCE
have not been identified in these species. At present, there is insufficient
evidence for assuming a qualitative species difference in PCE metabolism.
5.4.2 Enzymic Pathways of Metabolism
The pathways of PCE metabolism are speculative. The currently accepted
pathway for the production of TCA from PCE is shown in Figure 5-13. This
pathway was initially proposed by Powell in 1945 for trichloroethylene and was
subsequently supported for PCE by the results of Yllner (1961), Daniel (1963),
Liebman and Ortiz (1970, 1977), Costa and Ivanetich (1980) and others. The
first step in this pathway appears to be catalyzed by hepatic cytochrome P-450
system to give 1,1,2,2-tetrachloroethylene oxide, although formation of this
epoxide has not been rigorously verified by chemical isolation or identifica-
tion as an intermediate to TCA formation in microsomal reactions. However,
the evidence is strong as reviewed below. The epoxide has been chemically
synthesized (Frankel et al., 1957; Bonse et al. , 1975; Kline et al. , 1978).
In sodium phosphate buffer at pH 7.4, the oxide hydrolyzes at a pseudo-first
order rate of 6 x 10 min'1 (T^-12 min) to TCA, and thermally decomposes in
nonaqueous solution to trichloroacetyl chloride (Table 5-4). Sakamoto (1976)
also synthesized the epoxide in milligram amounts, essentially by the method
of Frankel et al. (1957), and determined urinary metabolites after i.p.
5-28
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ci ci
\ /
c =c
CI / CI
NADPH, O2
P450
CI
O
\/ \
CI
CI
\
EPOXIDE
HYDRASE
H20,
O O
c - c
/ \
CI OH
DECARBOXYLATION
V2H20
O
'OH
O
~OH
OXALIC ACID
CO2 +
HC
O
'OH
CI
CHLORIDE
MIGRATION
O
CI
H2O
O
OH
TCA
+H CI
Figure 5-13.
Postulated scheme for the metabolism of PCE.
1979; Costa and Ivanetich, 1980).
(After Pegg et al.,
5-29
-------
injection into guinea pigs. By GC analysis, TCA was the only metabolite; the
Fujiwara reaction, however, indicated the presence of a small amount of TCE
which this investigator believed to be an artifact of the Fujiwara method.
The enzymatic or spontaneous nonenzymatic steps in PCE conversion to TCA
does not explain other reported metabolites of PCE, particularly the well-
established urinary metabolite, oxalic acid (Table 5-7). Furthermore, a
significant portion of PCE is completely metabolized to C02 in a dose-dependent
manner (Tables 5-3, 5-4, and 5-6).
5.4.2.1 P-450 oxidation. That PCE is metabolized by hepatic microsomal
cytochrome P-450 system is strongly supported by several observations. This
enzymic system is known to metabolize analogs of PCE, i.e., vinyl chloride,
•vinylidene chloride, and trichloroethylene (Bolt et al., 1982). PCE has been
shown to bind to the active site of the enzyme system, as evidenced by the
production of a Type I difference spectrum in hepatic microsomes from rats jjn
vitro (Pelkonen and Vaino, 1975; Costa and Ivanetich, 1980). Costa and
Ivanetich (1980) and Liebman and Ortiz (1977) demonstrated with rat microsomes
j_n vitro that PCE also stimulated nicotinamide adenine dinucleotide reduced
(NADPH) oxidation which is characteristic of cytochrome P-450-dependent reac-
tions. Furthermore, TCA was produced during the aerobic incubation of hepatic
microsomes (0?; NADPH required) in a time-dependent manner (Figure 5-14). The
KM (1.1 mmol) and V (0.046 nmol TCA/ min/nmol P-450) were respectively de-
I I IllctX.
creased and increased with microsomes from rats treated with inducers of
cytochrome P-450 (Table 5-8). With respect to phenobarbital induction, the
data of Table 5-8 suggest the induction of one iso-form of cytochrome P-450
that has a greater PCE binding equilibrium and also a greater reactivity,
i.e., a greater ability to metabolize PCE by both increasing 2.5-fold hepatic
microsomal content of P-450 and also increasing 4-fold average P-450 reactivity.
Finally, inhibitors of cytochrome P-450 (SKF 525A, metyrapone, CO) added to
hepatic microsomes with PCE inhibited both spectral binding and metabolism of
PCE to TCA.
The equilibrium constant for the binding of PCE to cytochrome P-450 (1C =
0.43 mmol) is similar to that of trichloroethylene (0.4 mmol). However, the
maximum rates of uninduced metabolism (V ) of PCE by the cytochrome P-450
ni3x
system are approximately 30-fold lower than the corresponding rate for the
metabolism of trichloroethylene (0.046 versus 1.55 nmol TCA/min/nmol P-450)
(Table 5-8; Costa et al., 1980).
5-30
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INDUCER
PHENOBARBITAL
PREGENOLONE-16a-CARBONITRILE
UNINDUCED
fl-NAPHTHOFLAVONE
TIME, min
Figure 5-14.
TABLE 5-8.
Production of TCA from PCE by hepatic microsomes from different-
ly pretreated rats as a function of time. Incubation mixtures
contained in 0.2 M Tris-HCl (pH 7.4), tetrachloroethylene
(3.3 rtiM). NADPH-generating systems, EDTA (0.2 mM), and hepatic
microsomes (2 mg protein/ml) from uninduced (triangle), p-naph-
thoflavone-induced (diamond), pregnenolone-16a-carbonitrile
(square) and phenobarbital-induced (circle) male Long-Evans
rats. (From Costa et al., 1980).
METABOLISM OF PCE BY RAT HEPATIC MICROSOMES AND THE EFFECT
OF VARIOUS INDUCERS
carbonitrile
Phenobarbital
P-450 content
nmoles/mg
microsomal
2.32 ± 0.32
0.2 ± 0.1
max
nmoles TCA/min/
Inducing Agent
None
p-naphthoflavone
Pregnenolone-16 a
fjr u ie i ii
0.89 ± 0.06*
1.03 ± 0.01
1.50 ± 0.11
(mM)
1.1 ± 0.8
0.5 ± 0.5
2.4 ± 0.8
nmole P-450
0.046 ± 0.004
0.055 ± 0.014
0.16 ± 0.09
0.19 ± 0.02
*Mean ±S.D.; measured in presence of 3.3 mM PCE.
Source: Costa et al., 1980.
5-31
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5.4.2.2 Other Pathways. The occurrence of oxalic acid and of C02 as major
metabolites of PCE, at least in rodents (Table 5-7), indicates the existence
of pathway(s) of metabolism other than the primary TCA pathway. Assuming the
primacy of P-450 oxidation to PCE-oxide as the initial step in the metabolism
of PCE, as outlined above, it has been proposed by Pegg et al. (1979) that
PCE-oxide, after conversion to chloroethylene glycol by microsomal hydrase, is
dehydrohalogenated to tetrachlorodiacetyl chloride and then to oxalic acid or
CCL and formic acid (Figure 5-13). Liebman and Ortiz (1977) attempted a test
of the postulated PCE-oxide to diol step by adding inhibitors of hydrase
activity to microsomal preparations metabolizing PCE. These investigators,
however, did not observe the expected increase in the PCE-oxide to TCA pathway
by the addition of cyclohexene to rat microsome-PCE incubations to inhibit the
postulated epoxide-diol pathway. Nonetheless, these experiments were inconclu-
sive, particularly as the data did not include a control demonstrating hydrase
activity in their microsomal preparations, and the effect of the cyclohexene
inhibitor with a known substrate. According to Oesch (1972), there are two
types of epoxide hydrase, one of which is tightly coupled to the cytochrome
P-450 system and one of which is not; the coupled form is thought to be more
important in detoxifying epoxide formed by mixed-function oxidases and is
relatively resistant to inhibitors.
The postulated pathways of PCE metabolism (Figure 5-13) do not include
conjugation reactions as possible metabolic detoxification steps. Presently,
there is little evidence that PCE-oxide might conjugate with glutathione
(GSH), for example, as a detoxification pathway as postulated for vinylidene
chloride-oxide (Jones and Hathway, 1978; Reichert et al. , 1979). Pegg et al.
(1979) found that inhalation exposure of rats to 600 ppm (4068 mg/m3) PCE for
6 hours had no significant effect on total hepatic nonprotein sulfhydryl, an
estimate of hepatic GSH content. GSH concentration was estimated as 5.06 ±
0.02 and 4.93 ± 0.21 umol/mg liver for control and exposed animals, respec-
tively. The pathways for the origin of major metabolites of PCE other than
TCA (Table 5-7) are deserving of further attention and investigation.
5-4.3 Magnitude and Dose-Dependency of Metabolism
Experimental human and animal exposures to PCE have been carried out with
an overall finding that PCE metabolism is very limited, i.e., that saturation
of mammalian capacity to PCE occurs at relatively low levels of exposure.
5-32
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5.4.3.1 Man. Estimates of the extent of metabolism in man, as a percentage
of retained PCE, have been made from balance studies by accounting for a
retained dose after inhalation exposure by measuring metabolites (TCA, total
trichloro-compounds by Fujiwara reaction) excreted in the urine (Stewart
et al., 1961, 1970; Monster et al. , 1979; Monster and Houtkooper, 1970; Monster
et al., 1983; Boettner and Muranko, 1969; Ikeda et al., 1972; Essing et al.,
1973; Fernandez et al., 1979; May, 1976). Only about 1 to 3 percent of the
estimated amounts absorbed are metabolized into TCA and other chlorinated
metabolites. Ikeda and co-workers (Ikeda et al., 1972; Ohtsuki et al. , 1983)
have explored the relationship of inhalation exposure concentration in occupa-
tional settings (time-weighted 8-hr average) metabolite concentration in urine
(total trichloro-compounds; Fujiwara reaction) at the end of an 8-hr work
shift. Under these conditions, these investigators found a saturation of
•3
metabolism occurs between 100 to 400 ppm (678 to 2112 mg/m ) PCE in inspired
air (Figure 5-9). They estimated that at the end of an 8-hr shift with expo-
3
sure to PCE at 50 ppm (339 mg/m ) (below saturation; Figure 5-9), 38 percent
of PCE pulmonary uptake would be exhaled unchanged, and less than 2 percent
would be metabolized to urinary chlorinated compounds, while the rest would
remain in the body to be eliminated in succeeding hours.
The difficulties associated with balance studies in man are the problems
encountered with (1) accurate measurement of the retained dose of PCE from
inhalation exposure, (2) the imprecision of the older methodologies using the
Fujiwara reaction for metabolite quantification, (3) the possibilities that
metabolites other than TCA (e.g., oxalic acid, TCE-glucuronide) may be excreted
in urine or.by biliary excretion, and (4) the very long half-life (144 hours)
of PCE urinary metabolites which necessitates extended collection of samples.
For example, Bolanowski and Golacka (1972) in their series of experimental
exposures to PCE in man found that the elimination of TCA in the urine repre-
sents only a few percent of the dose while the alveolar retention of PCE was
60 to 80 percent (Fernandez et al., 1976; Monster et al., 1979). This suggests
a total metabolism of 20 to 40 percent of the dose. Thus they assume that
another unrecognized pathway exists for PCE which has not yet been taken into
consideration. Conversion of PCE to COp could constitute a significant portion
of this metabolism.
5.4.3.2 Animals. For rodents (rat and mouse), the extent of metabolism after
single doses of PCE has been estimated by balance studies using isotopically
labelled PCE. Yllner (1961) carried out the earliest study in mice. By
5-33
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exposing the mice by inhalation to C-PCE vapor, he found that of the radio-
active PCE absorbed (2 hr-exposure; 0.9 mg/kg b.w.), 70 percent was excreted
unchanged in expired air, 2 percent was metabolized and excreted in the urine,
and 0.5 percent in the feces over a four day post-dosage period. Daniel
(1963) gavaged rats with 36C1-PCE (about 1.12 and 8.29 mg/g b.w.) and found
that 98 percent of these doses were excreted unchanged by the lungs and only
2 percent of radioactivity was metabolized and excreted in the urine over a 2
to 18-day period (Table 5-9). These early observations, which indicate a very
limited metabolism of PCE by rodents, are in accord with the more recent
kinetic studies of Filser and Bolt (1979). These investigators exposed rats
to PCE in a closed system and measured the rate of disappearance of PCE vapor.
The kinetics of metabolism of PCE was dose-dependent and the estimated M
(TldX
was less than 1.2 mg/kg/hr. Filser and Bolt noted that PCE was metabolized
extremely slowly in rats (in comparison to other halogenated ethylenes) and
they were unable to experimentally differentiate between zero order and first
order kinetics and hence to determine the saturation point, i.e., ppm PCE in
air at which metabolism "saturation" occurs.
Other investigators have also found evidence of dose-dependent Michaelis-
Menten kinetics for metabolism of PCE in mice and rats following both oral and
inhalation exposure. Pegg et al. (1979) exposed rats to-single oral doses of
14
C-PCE (1 and 50 mg/kg b.w.) and to single inhalation exposures (10 and
600 ppm for 6 hours) and followed the disposition of radioactivity in exhaled
air, urine, and feces for 3 days post administration. Their data are given in
Table 5-3. For these low doses (1 mg/kg and 10 ppm, 6 hours), 71.5 and 68.1
percent, respectively, of the body burdens were excreted as unchanged PCE in
exhaled air. The remainder, 28.5 and 31.9 percent, was metabolized to C02 and
fecal and urinary metabolites. In contrast, for the higher oral and inhalation
exposure doses (500 mg/kg and 600 ppm, 6 hours), a greater percentage of the
body burdens was exhaled as unchanged PCE (89.9 and 88.0 percent, respectively)
Hence, these data demonstrate that even at very low body burdens of PCE in the
rat (approximately 0.25 to 2.5 mg per animal) only 30 percent is metabolized,
and this percentage decreased with increasing body burden. This indicates
that metabolic disposition of PCE is a saturable, dose-dependent process in
this species after either oral (gavage) or inhalation exposures. Recently,
Frantz and Watanabe (1983) have demonstrated that PCE from saturated solutions
of drinking water, allowed ad libitum over a 12-hr period to rats, provided
5-34
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TABLE 5-9. DISPOSITION OF 36C1-PCE AFTER ORAL ADMINISTRATION TO
WISTAR RATS
Dose
(uc)***
1.75
13
Mg/Kg
Equivalents
1116
6160
No.
Animal s
1
4
Expired air*
(unchanged)
PCE
97.9
--
Metabol
Urine**
2.1
1.6
ized
Feces
0
0
"Collected for 48 hr.
**Collected for 18 days.
***Specific activity of 36C1-PCE was 1.3 uc/mmol.
Source: Daniel, 1963.
average body burdens of 8.09 ± 3.13 mg/kg (2.3 mg per animal), and gave com-
parable results to low dose (1 mg/kg) gavage dosage, i.e., 88 percent of the
body burden of PCE was excreted unchanged in exhaled air and only approximately
12 percent was metabolized. Their results are shown in Table 5-6.
Schumann et al. (1980) have investigated the extent of metabolism in mice
using Olabelled PCE. The experimental conditions were very similar to
those of Pegg et al. (1979) and Frantz and Watanabe (1983) in rats as these
experiments originated from the same laboratory. Schumann and co-workers
determined the disposition in mice of body burdens of PCE of 0.04 mg per mouse
(from 10 ppm, 6 hours inhalation exposure) and of 10.8 mg per mouse (from
500 mg/kg gavage dose) (Table 5-4). For the smaller body burden, 12 percent
was excreted unchanged in exhaled air and the remainder (about 88 percent)
metabolized; for the larger body burden, the proportions were reversed with
83 percent of the body burden excreted unchanged and about 17 percent disposed
of by metabolism. These data indicate that for mice also, metabolism of PCE
is dose-dependent and saturable for this species. When compared to rats
(Tables 5-3 and 5-6 versus Table 5-4), mice were found to metabolize 8.5 and
1.6 times more PCE per kg body weight following inhalation of 10 ppm (68 mg/m )
for 6 hours or a single oral dose of 500 mg/kg respectively. However, on a body
burden per animal basis, the rat metabolized 6.8 and 1.3 times more PCE than the
mouse following inhalation of 10 ppm (68 mg/m ) for 6 hours or a single oral
5-35
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dose of 500 mg/kg respectively. Thus, from these observations, the metabolism
of PCE by these two species is more consistent with a metabolism proportional
-to body surface area than to kg body weight.
Buben and 0'Flaherty (1984) have also investigated the extent of PCE
metabolism in male mice (Swiss-Cox strain) using urinary metabolite excretion
as an index of metabolism. These investigators daily dosed the mice by gavage
with PCE in corn oil, 5 days per week for 6 weeks, at dose levels of 20, 100,
200, 500, 1000, 1500, and 2000 mg/kg/day. TCA was the only metabolite found
in the urine of this strain with less than 5 percent of the dose excreted in
the feces. Nonetheless, total urinary metabolite as an index of metabolism is
only a dose approximation since it is known that some metabolized compound is
bound to cellular macromolecules and some is metabolized to CO^ (Tables 5-3,
5-4, and 5-6). However, urinary excretion is by far the major route of disposi-
tion of metabolized PCE. Therefore, total urinary metabolite should be a
reasonable approximation of the amount of metabolism. Also, with chronic
administration, body burden of PCE is expected to reach steady-state, and in
fact these investigators found no week-to-week trends in daily urinary meta-
bolite output although urinary TCA tended to increase during the week with
plateauing towards the end of the week; urine collections were made at the
weekly plateau. Figure 5-12 shows the relationship found between the amount
of metabolism and the administered dose. The data clearly show a dose-dependent
metabolism consistent with capacity limited kinetics in accordance with Michaelis-
Menten equation, Y = M - X/K + X, where X is the dose (mg/kg) and Y is
maX ITS
mg/day of metabolite resulting. From the computer-fitted equation, Buben and
0'Flaherty estimated Mmo^ (maximum rate of metabolite formed and excreted in
INclX
24 hr) to be 136 mg/kg/day and Km (the dose at which the amount of metabolite
excreted in a 24-hr period is half the apparent maximum amount) at 660 mg
PCE/kg. These workers also estimated that at a very low chronic oral dose
of PCE only 25 percent was metabolized (indicating significant hepatic flow-
limited or first pass effect) and less than 5 percent of a very high dose.
These observations also indicate that, in this strain of mice at least, the
metabolism of PCE deviates from linearity (i.e., from first order) at oral
doses above 100 to 200 mg/kg. Of considerable importance also is the finding
of a clear relationship between toxicity and metabolism of PCE. Figure 5-15
shows a direct linear relationship between hepatotoxicity (serum SGPT activity)
and urinary metabolite excretion. Similar relationships were observed with
5-36
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other indices of hepatotoxicity such as liver weight increase, liver trigly-
ceride accumulation, and glucose-6-phosphatase activity. These observations
are consistent with the formation of a toxic or reactive intermediate(s) in
the primary metabolic pathway of PCE (Figure 5-13).
5.4.4 Covalent Binding
PCE is known to bind to the substrate binding site of hepatic microsomal
cytochrome P-450 i_n vitro (Costa et al., 1980). A direct relation also has
been demonstrated among the level of hepatic microsomal cytochrome P-450,
the extent of metabolism of PCE i_n vivo, and cellular damage (Bonse et al.,
1975; Moslen et al., 1977; Pegg et al., 1979; Schumann et al. , 1980; Buben and
0'Flaherty, 1984). For example, Moslen et al. (1977), in a study designed to
correlate metabolism with hepatocellular toxicity, observed that induction of
microsomal systems in rats by phenobarbital or chlorinated biphenyls increased
metabolism of orally administered PCE (5.7-fold) and increased indices of
toxicity (2-fold). Buben and 0'Flaherty (1984) have also presented evidence
for a strong correlation between hepatocel1ular toxicity and metabolism. As
shown in Figure 5-15, these workers found in mice linear relationships between
the amount of metabolism of PCE (as measured by urinary metabolites) and
hepatotoxicity as measured by several indices of toxicity. In terms of the
proposed metabolic pathway(s) for PCE (Figure 5-13), covalent binding to
cellular macromolecules, leading to cellular damage, may occur with putative
reactive intermediates of PCE metabolism: PCE-oxide, trichloroacetyl chloride,
and chlorodiacetyl chloride. Bonse et al. (1975) and Bolt et al. (1982) have
proposed that the symmetrical PCE oxide metabolite is much less reactive than
the epoxides of unsymmetrical chlorinated ethylenes such as trichloroethylene,
vinylidene chloride, and vinyl chloride. Consequently, covalent binding of
PCE metabolites to cellular macromolecules may result in less severe hepatic
toxicity and genotoxicity, even when taking into account the large differences
in the hepatic abilities to metabolize these compounds (Section 5.4.3). This
correlation between biological activity and chemical stability has been con-
firmed by Jonas and Mackrodt (1982), who studied the relationship between the
mutagenic activity of chlorinated ethylenes and the C-0 bond energy of their
epoxides.
Bonse et al. (1975) perfused isolated rat liver preparations with PCE
(180 ppm concentration) and observed that only 3 to 5 percent of the total
uptake bound to liver tissue; this portion was extractable only after acid
5-37
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40 60
TOTAL URINARY METABOLITE, mg/kg
80
100
Figure 5-15.
Relationship between hepatotoxicity parameters from PCE oral
administration (shown as serum SGPT activity increases) and
total urinary metabolite excreted per day by mice at increasing
dose levels of PCE. The slope of the linear regression line is
0.436 with a coefficient of
Buben and 0'Flaherty, 1984.
regression, r2, of 0.949. From
5-38
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hydrolysis. They proposed that the acyl chloride intermediate in the conversion
to TCA (10 to 15 percent of uptake) reacted with cell functional groups, e.g.,
-OH, -5H, and -NH2- Pegg et al. (1979) have demonstrated i_n vivo rat hepatic
binding after exposure to PCE by both inhalation and oral routes. Their data
are shown in Table 5-10. Radioactivity irreversibly associated with tissue
macromolecules was assayed 72 hr after termination of inhalation exposure to
10 or 200 ppm (68 or 1356 mg/m3) for 6 hr and 72 hr after oral administration
14
of C-PCE. Irreversible binding (primarily to cellular protein and nucleic
acid) was dose-dependent and metabolism-dependent, but independent of the
route of exposure. The ratio of amount bound to total amount metabolized
(about 1 percent; Table 5-10) was not significantly different between high and
low doses of PCE by either route of exposure, suggesting that irreversible
binding was solely a function of metabolism and the availability of reactive
species. The persistence of irreversible bound metabolites of PCE (72 hr
after exposure) indicates that turnover rates of the bound metabolite(s) are
slow enough that accumulation could potentially occur in man with repeated
daily exposures.
TABLE 5-10. IRREVERSIBLE HEPATIC BINDING OF 14C-PCE IN RATS
72 HR AFTER EXPOSURE
Route
Oral
Inhalation
(6 hr)
Exposure
1 mg/kg
500 mg/kg
10 ppm
600 ppm
Amount
Metabol ized
umol eq
0.41 ± 0.03*
73.92 ± 4.98
2.82 ± 0.16
55.46 ± 0.44
Bound
(jmol eq/
1 i ver protein
0.0035 ± 0.0003
0.4321 ± 0.0477
0.0245 ± 0.0022
0.4029 ± 0.0222
Bound/
Metabolized
x 100
0.83
0.59
0.87
0.72
*Mean ± S.E.
Source: Pegg et al., 1979.
Schumann et al. (1980) have compared metabolism of PCE and irreversible
hepatic binding in the two species, rat (Sprague-Dawley) and mouse (BgC^F^).
Their data are summarized in Table 5-11. The time course of hepatic macro-
molecular binding differs between these two species; the peak binding occurs
in mice significantly more rapidly than in rats after exposure from both
5-39
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TABLE 5-11 COMPARISON OF IRREVERSIBLE HEPATIC BINDING OF 14C-PCE IN SPRAGUE-DAWLEY
RATS AND BgCsFi MICE AFTER INHALATION (6 HR) AND ORAL EXPOSURES
|jg mol-eq bound/ g hepatic protein
Time,
hr post-
exposure
0
1
6
12
24
48
72
A.
1.604
1.355
1.027
0.729
0.481
600 ppm
Mice
± 0.251*
-
± 0.120
-
± 0.138
± 0.232
± 0.127
B.
0.174
0.227
0.279
0.246
0.171
Rats
± 0.064
-
± 0.026
-
± 0.017
± 0.062
± 0.050
A/B
9.2
-
6.0
-
3.7
3.0
2.8
A. Mice
0.
2.
1.
1.
1.
0.
-
964 ±
171 ±
687 ±
386 ±
081 ±
472 ±
0.115
0.523
0.112
0.262
0.290
0.131
500 rag/kg
B.
0.170
0.331
0.402
0.612
0.347
0.321
Rats
-
± 0.025
± 0.027
± 0.086
± 0.027
± 0.060
± 0.060
A/B
-
5.7
8.2
4.2
2.3
3.1
1.5
*Mean ± S.D.
Source: Schumann et al. , 1980.
inhalation and oral routes of administration. There is 3 to 5 times more
binding per gram hepatic protein in mice than rats. Since liver weight of these
two species is proportional to species body weight, the total hepatic-bound
PCE is comparable to the metabolism of PCE by the mouse and rat (Section
5.4.3.2).
Schumann et al. (1980) also have tried to evaluate the extent of binding
to DNA and its relation with the probability of genotoxicity. These workers
utilized the mouse in their experiments because of greater binding per gram of
liver in this species, and because of the susceptibility of this species to
hepatic tumors (NCI, 1977). These investigators did not find a detectable
binding of high specific activity 14C-PCE (1.3 x 105 dpm/nmol) to liver DNA
when mice were exposed by inhalation (600 ppm) or by gastric intubation (500
mg/kg). However, the limit of detection was estimated as 10 to 14.5 alkyla-
tions per 10 nucleotides. Therefore, levels of DNA binding below their
detection limit may have occurred. In parallel experiments, these investi-
gators (Schumann et al., 1980; Watanabe et al., 1980) also chronically admin-
istered PCE (11 to 12 daily doses) by gavage at increasing dose levels up to
1000 mg/kg (Table 5-12). Mice, but not rats, showed a 20-25 percent increase
in liver/body weight ratio. DNA/g liver protein decreased, indicating that
5-40
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TABLE 5-12. EFFECTS OF CHRONIC ORAL ADMINISTRATION OF PCE ON HEPATIC DNA
CONTENT AND DNA SYNTHESIS IN MICE AND RATS3
Liver DNA content
(mg DNA/g liver)
Dose
(mg/kg/day)
12 doses in 16 days
0
500
11 doses in 11 days
0
100
250
500
1000
4.00
3.31
2.51
2.29
2.25
2.23
2.30
Mice
± 0.42
± 0.22*
± 0.11
± 0.10*
± 0.17*
± 0.15*
± 0.10*
Rats
2.22 ± 0.29
2.03 ± 0.17
2.33 ± 0.19
2.24 ± 0.22
2.51 ± 0.16
2.36 ± 0.24
2.18 ± 0.27
Liver DNA synthesis
(dpm [3H]thymidine/Mg DNA)
Mice
65 ± 34
118 + 22*
44 ± 15
56 ± 27
63 ± 26
75 ± 23
82 ± 31*
Rats
63 ± 11
65 ± 17
88 ± 24
46 ± 26
137 ± 103
105 ± 45
108 ± 45
aMean ± SD (n = 3-7); [3H]thymidine 1000 pCi/kg i.p. 6 hr prior to sacrifice.
*
p <0.05.
Source: Schumann et al. (1980).
organ weight gain resulted from hypertrophy rather than hyperplasia. Nonethe-
less, DNA synthesis ( C-thymidine incorporation) tended to increase in both
mice and rats in a dose-related manner, suggesting some activation of "repair"
mechanisms.
5.4.5 Interactions with Metabolism
5.4.5.1 Induction and Inhibition. The metabolism of PCE is known to be
increased by inducers of the cytochrome P-450 oxidative system (microsomal
mixed-function monooxygenases). Moslen et al. (1977) demonstrated that the
metabolism of PCE in intact rats, as measured by urinary metabolites, was
enhanced (4 to 7-fold) by pretreatment with inducers of cytochrome P-450,
phenobarbital and polychlorinated biphenyls, which increased the hepatic
cytochrome content 2.4 and 3.3-fold, respectively. Similar results have been
reported by Ikeda and Imamura (1973) for both intact rats and hamsters. Costa
and Ivanetich (1980) found that pretreatment of rats with the inducers, pheno-
barbital and pregnenolone-16orcarbonitrile, increased cytochrome P-450 content
of microsomes and markedly enhanced the metabolism of PCE to TCA by these
5-41
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microsomes i_n vitro (Figure 5-14). Inhibitors of cytochrome P-450, SKF 525A,
metyrapone, and CO, added to the microsomal reaction medium, significantly
decreased the production of TCA.
Chronic exposure to PCE appears to produce minimal self-induction of
metabolism. Kaemmerer et al. (1982) report that feeding rats 25 and 200 ppm
PCE in their diet resulted in a small but statistically significant induction
of cytochrome P-450 in rat liver homogenates after 14 and 7 days treatment,
respectively. Vainio and co-workers (Savolainen et al., 1977; Vainio et al.,
1976) found a slight depressing effect on rat liver P-450 content 24 hours
after a single high oral dose of PCE (1700 mg/kg) but reported slight increases
3
after PCE inhalation exposure at 600 ppm (4068 mg/m ), 6 hours daily for 5
days. Plevova et al. (1975) and also Kaemmerer et al. (1982) observed that in
3
rats, exposure to 177 ppm (1200 mg/m ) PCE for 6 hours, 20 hours prior to i.p.
administration of pentobarbital (44 mg/kg), lengthened by 30 percent the
pentobarbital sleeping time. These observations suggest that substrate competi-
tion for cytochrome P-450 system for the metabolism of pentobarbital and for
PCE may occur, with a consequent increase of rate of metabolism of phenobarbital
and prolongation of its half-life.
5.4.5.2 Drugs. In addition to the potential for PCE interactions with
clinical drugs, some of which are known to induce P-450 system or compete for
P-450 oxidative metabolism, other types of interactions are possible. For
example, Hake and Stewart (1977) evaluated the effect of diazepam and of
alcohol, CNS depressants, on human exposure to PCE. During an 11-week study
with inhalation exposure to PCE, 5.5 hours per day at 25 and 100 ppm (169 and
678 mg/m ), concomitant administration of ethanol (blood levels 30 to 100 mg
percent) increased PCE blood levels in the 25 ppm exposure but not with 100
ppm; diazepam (blood levels 7 to 30 ug percent) had little effect. Addition-
ally, behavioral and neurological tests revealed no interactive effects under
these conditions. However, May (1976) found that urinary excretion of TCA
was completely inhibited by a blood ethanol concentration of 0.65 mg percent.
5.4.5.3 Other Interactions. PCE has been observed to have a significant
effect on intermediary metabolism. Takano and Miyazaki (1982) and Miyazaki
and Takano (1983) reported that the PCE inhibited glutamate, succinate, and
malate oxidation by rat liver mitochondria. The lesser inhibition of succinate
oxidation suggested to the authors that PCE may act as an uncoupler in the
electron transport system between NADH dehydrogenase and coenzyme Q. Ogata
5-42
-------
and Hasegawa (1981) have reported similar findings for PCE and other halocar-
bons for succinate oxidation by rat liver mitochondria. Since both the K and
Vmax of the reactions were affected by PCE, it was suggested that PCE represents
a nonspecific inhibitor (possible mitochondrial membrane perturbation by this
lipid solvent) rather than a specific inhibitor like cyanide or antimycin A.
5.5 SUMMARY
Perchloroethylene is rapidly and virtually completely absorbed from the
gastrointestinal tract of rats and mice after administration of doses up to
500 mg/kg by intragastric intubation, and of doses up to 8 mg/kg in drinking
water, with peak blood concentrations occurring within one hour of dosing.
Pulmonary uptake of PCE during inhalation exposure by experimental animals and
man is linearly proportional to exposure duration and concentration in air; it
is also influenced by physical activity and body mass. For rats exposed to 10
3
and 600 ppm (68 and 4068 mg/m ) for 6 hours, net uptake of PCE has been measured
as 5.9 and 310 mg/kg b.w.; for man exposed to 72 and 144 ppm (488 and 976 mg/m )
for 4 hours, the net uptake averages 6.5 and 13.5 mg/kg. Steady-state uptake
in man is not established within short (8 hours) exposure periods. With
chronic exposure (100 ppm; 8 hr/day; 5 days/wk), 3 to 4 weeks is required for
steady state. Absorption of PCE during vapor or liquid contact with the skin
of experimental animals or man is very slow and adds less than 1 percent to
the body burden.
PCE distributes widely into body tissues, and readily crosses the blood-
brain barrier and the placental barrier. Highest tissue concentrations are
found in the adipose tissue (60 times blood level), and in brain and liver (4
and 5 times blood level, respectively). Tissue concentrations increase in
direct proportion to the body burden. With chronic exposure in man, adipose
tissue concentrations do not achieve plateau concentrations until 3 to 4 weeks
of exposure.
PCE is eliminated from the body principally by pulmonary excretion of
unchanged compound and also by a rate-limited metabolism of PCE. Pulmonary
excretion in man and experimental animals occurs in three first-order phases
of desaturation of blood vessel-rich tissue group (VRG), muscle tissues (MG),
and adipose tissues (FG). For man, the half-times of elimination from these
groups are 12-16 hr, 30-40 hr, and 55-65 hr, respectively. For rat, the domi-
nant half-time of pulmonary elimination is about 7 hours.
5-43
-------
The capacity of man and rodents to eliminate PCE by metabolism is limited.
For man, as assessed by urinary metabolite excretion, about 1-3 percent of the
2
absorbed amounts of PCE from exposures of 100 to 400 ppm (678 and 2112 mg/m )
exposure for 8 hours are metabolized, with a half-time of renal elimination of
the metabolites of about 144 hours. Urinary metabolite excretion increases as
PCE body burden increases until a plateau is reached at about 400 ppm (2112
mg/m3) exposure (500 to 600 mg body burden). Metabolism is, therefore, dose-
dependent and saturable in man. Similar observations have been made for
rodents. For the rat and mouse, the percentage of the administered dose that
is metabolized decreases with increase of dose. Metabolism in both the rat
and mouse is consistent with saturable, dose-dependent Michaelis-Menten kinetics
with a maximum rate (V ) in the order of 1.6 mg/hr/kg b.w. for the rat and
max
5.7 mg/hr/kg for the mouse.
Indices of hepatotoxicity of PCE in the rat and mouse are highly correl-
ative with the dose-dependent nature of the metabolism of PCE in these species.
The principal site of metabolism appears to be hepatic P-450 monooxygenase
system where PCE is oxidized to PCE oxide, which rearranges to trichloroacetic
acid (TCA). TCA has been identified as a major urinary metabolite in man,
rat, and mouse. Metabolism to TCA, i_n vitro and i_n vivo, is known to be
increased by inducers of the P-450 oxidative system. Secondary pathways of
metabolism may include hydration of PCE oxide with subsequent dehydrohalogena-
tion of the glycol to chlorodiacetyl chloride and further rearrangement to
yield oxalic acid or formic acid and carbon dioxide. In rodents, oxalic acid
is a urinary metabolite while carbon dioxide is a pulmonary metabolite. There
is no evidence that conjugation reactions with glutathione or cysteine are
important in the metabolism of PCE. While the metabolite profiles of PCE in
man, rat, and mouse are not as yet fully established, there is no convincing
evidence of qualitative or quantitative differences of pathways in these
species.
PCE metabolite(s) are known to covalently bind, i_n vitro and in vivo, to
cellular macromolecules such as protein and lipids. Tissue binding appears to
be solely a function of metabolism and availability of reactive metabolites.
Since tissue-bound metabolites have a slow rate of turnover and accumulate,
cumulative cellular damage may occur with chronic exposure. Covalent binding
to hepatic DNA i_n vivo is minimal; however, as the limit of detection for
these experiments was only 10 to 14.5 alkylations/10 nucleotides, very low
levels of DNA binding cannot be completely excluded.
5-44
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6. TOXIC EFFECTS
6.1 HUMANS
The known effects of tetrachloroethylene (perchloroethylene, PCE) on
humans have been established primarily from Individuals accidentally or occupa-
tional^ exposed to very high (in nearly all cases, unknown) concentrations of
PCE. Exposure to high concentrations of PCE causes a variety of toxicological
effects in humans. Effects upon the central nervous system (CNS) are generally
the most noticeable following acute or excessive occupational exposures.
Effects upon the liver and kidney usually are observed after an elapsed period
of exposure to high concentrations.
6.1.1 Effects on the Liver
6.1.1.1 Acute Exposure Situations—Associations between liver damage and
acute exposures of humans to high air levels of PCE have been reported by
several investigators (Stewart et al. , 1961a; Stewart, 1969; Sal and, 1967;
Hake and Stewart, 1977; Levine, 1981).
Clinical evidence of impaired liver function was found by Stewart et al.
(1961a) in a worker rendered semi-comatose by overexposure to a petroleum-based
solvent estimated to contain about 50 percent PCE. Simulation of the exposure
suggested that during the 3.5-hr exposure, the last 30 minutes probably averaged
•2
about 1000 ppm (6780 mg/m ) PCE; the average estimated concentration during
the initial 3 hours was about 275 ppm (1864 mg/m3). At such levels, metabolic
saturation was exceeded. A neurological examination conducted 1 hour after
the worker's collapse indicated no abnormality of function; also, the liver
was not palpable. During a 6-week period subsequent to the incident, neither
clinical jaundice nor neurological deficits were observed. Beginning 9 days
postexposure, urinary urobilinogen and total serum bilirubin were elevated.
Serum glutamic-pyruvic transaminase (SGPT) and serum glutamic-oxaloacetic
transaminase (SCOT) were in the normal range throughout the postexposure
period. On the 18th day, SGPT was slightly elevated. On the 16th postexposure
day, the PCE level in expired air was sharply elevated. The investigators
suggested that such an acute overexposure may represent a continuing insult to
the liver since PCE had an exceedingly long exponential decay in expired air,
indicating slow release from body tissues.
In another case report, Stewart (1969) diagnosed transient and mild
hepatitis in a worker exposed to anesthetic levels of PCE for 30 minutes.
6-1
-------
Urinary urobilinogen was elevated on the 9th postexposure day. SCOT was
slightly elevated on the 3rd and 4th postexposure days. Stewart concluded
this patient had experienced marked depression of the central nervous system
followed by transient, minimal liver injury.
Hake and Stewart (1977) reported mild liver injury, as indicated by
elevated serum enzymes, in a 60-yr-old male overcome by PCE vapors. The
individual was reported to have fully recovered.
In the reports by Levine et al. (1981) and Sal and (1967), a cause-effect
relationship between PCE and the effects observed cannot be made.
An enlarged liver and obstructive jaundice were diagnosed by Bagnell and
Ellenberger (1977) in a 6-wk-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 estab-
lishment 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
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.
6.1.1.2 Chronic Exposure Situations—Alterations in liver function in persons
exposed to 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 have been altered as a result of excessive PCE exposure
include sulfobromophthalein retention time, thymol turbidity, serum bilirubin,
serum protein patterns, cephalin-cholesterol flocculation, serum alkaline
6-2
-------
phosphatase, SCOT, and serum lactic acid dehydrogenase (LDH). However, these
parameters may result from other causes that are completely dissociated with
PCE.
Effects attributed to exposure to PCE at high or unknown levels included
cirrhosis of the liver (Coler and Rossmiller, 1953), toxic hepatitis (Hughes,
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).
Chielewski et al. (1976) found that SCOT and SGPT levels were significantly
elevated in a group of 16 of 25 workers compared to non-exposed controls. The
group of 16 had been exposed to PCE vapors in the range of 59 to 442 ppm (400
to 3000 mg/m ) for periods ranging from 2 months to 27 years. Serum enzyme
levels in the remaining group of 9 who were exposed to levels of PCE at or
3
below 29 ppm (200 mg/m ) were normal. Low urinary excretion of 17-ketosteroids
and abnormal EEG tracings was reported for members of both exposed groups, but
no cause-effect relationship to PCE exposure can be inferred.
6.1.2 Effects on Kidneys
Associations between high air levels of PCE and symptoms of renal dysfunc-
tion have been reported by Larsen et al. (1977) and Hake and Stewart (1977).
Symptoms included diminished excretion of urine, uremia, elevated serum creati-
nine (Larsen et al., 1977), and proteinuria and hematuria (Hake and Stewart,
1977). A cause-effect relationship with PCE exposure cannot be inferred from
the report by Larsen et al. (1977). Similarly, other factors that may have
been responsible for the observation of renal damage in the case report by
Hake and Stewart (1977) cannot be precluded.
6.1.3 Effects on Other Organs/Tissues
6.1.3.1 Effects Upon the Pulmonary System and Skin--0verexposure to high but
unknown concentrations of PCE have been associated with pulmonary damage
(Patel et al., 1977; Levine et al., 1981).
Effects of PCE upon the skin range from a mild to moderate burning sensa-
tion upon direct contact for 5 to 10 minutes, to a marked erythema after
prolonged exposure, and finally., blistering if PCE is trapped under clothing
6-3
-------
or in shoes (Hake and Stewart, 1977). Observations of this nature have been
reported also by Stewart and Dodd (1964), Morgan (1969), and Ling and Lindsay
(1971).
6.1.3.2 Effects Upon the Heart—In the case report of Abedin et al. (1980),
work-related exposure of an individual to PCE was apparently the principal
factor leading to dizziness, headaches, and premature ventricular beats. The
individual's prior history showed no other possible contributing factors. The
individual's symptoms were more pronounced when the PCE plasma level was high
(3.8 ppm). Removal from exposure to PCE was reported to relieve the symptoms
as well as the premature ventricular beats.
6.1.4 Effects on CNS and Behavior
6.1.4.1 Effects of Short-Term Exposure—Reports of accidental short-term
exposure to PCE have implied that such exposure produces temporary central
nervous system (CNS) effects (Coler and Rossmiller, 1953; Lob, 1957; Eberhardt
and Freundt, 1966; Gold, 1969; Stewart, 1969; Bagnell and Ellenberger, 1977).
Effects most prominently mentioned were dizziness, confusion, headache, nausea,
and irritation of eyes and mucous tissue. Higher-level exposures intensified
the above symptoms and, at sufficiently high levels, produced unconsciousness
and eventually death.
An early report of short-term experimental exposures to PCE of four
subjects (Carpenter, 1937) described subjective effects similar to those
listed above. He also observed signs of increased autonomic activity such as
palmar sweating, salivation, and nasal secretion. Such symptoms were observed
3
at 500 ppm (3390 mg/m ) PCE exposure for about 2. hours. Concentrations of
1000 to 1500 ppm (6780 to 10,170 mg/m3) PCE for up to 2 hours produced faint-
ness. Higher concentrations produced faintness more quickly.
Rowe et al. (1952) collected subjective reports from six subjects exposed
to PCE concentrations ranging from about 100 to 1000 ppm (678 to 6780 mg/m3).
No effects were observed at 100 ppm (678 mg/m ), but the odor was noticeable.
The threshold for eye irritation, dizziness, and sleepiness was less than 200
3 3
ppm (1356 mg/m ) for 20 to 30 minutes. Exposure to 280 ppm (1898 mg/m )
intensified the symptoms and also produced "lightheadedness" and numbness
about the mouth. Subjects reported feelings of motor incoordination at 600
q
ppm (4068 mg/m ) exposure levels after 10 minutes. One minute of exposure at
1000 ppm (6780 mg/m ) produced all of the above symptoms and was considered
intolerable by the subjects.
6-4
-------
Stewart et al. (1970) collected subjective reports and a few objective
measures of CMS function in 11 subjects exposed to 100 ppm (678 mg/m3) for 7
hours. This single-day study was part of a long-term exposure experiment.
Symptoms reported were headache, mucous tissue and eye irritation, autonomic
symptoms such as flushed skin, dizziness, sleepiness, and speech difficulty.
Objective data collected were performance of the Romberg test of balance, the
Flanagan test of coordination, an arithmetic test, a visual inspection task, a
visual acuity test, and a test of depth perception. Of the objective tests,
none showed any abnormality except for the Romberg test, which was abnormal in
three of the 11 subjects. No control group was used and data were compared to
known norms. Stewart et al. (1977) executed another long-term repeated exposure
study of PCE using 12 subjects exposed to 0, 25, or 100 ppm (0, 169, or 678
3
mg/m ) PCE. Data collected concerned the Romberg test of balance, the Michigan
eye-hand coordination test, rotary pursuit performance, Flanagan coordination
test, eye saccade velocity measurement, the Lorr-McNair mood scale, and a
divided attention task involving the simultaneous detection of one of an array
of peripheral red lights while simultaneously counting the occurrence of a
white central light. EEC spectra were also collected. No effect was noted on
any of these tests on the first day of exposure to 100 ppm (678 mg/m ) PCE.
In this study, data were compared to control (0 or 25 ppm PCE) days. This
should have made the tests more sensitive. It is difficult to explain why
3
this experiment using 100 ppm (678 mg/m ) PCE did not show effects on the
Romberg test while in Stewart et al. (1970) effects were reported. The fact
that only three of the 11 subjects showed an abnormal Romberg test in Stewart
et al. (1970) might indicate abnormally sensitive subjects or the occurrence
of chance results.
Only three of the above publications (Rowe et al., 1952; Stewart et al.,
1970; Stewart et al. , 1977) reported sufficiently quantitative data at lower
exposure limits so that CNS threshold effects can be estimated. It appears
that such effects as dizziness, eye and mucous tissue irritation, headache,
and sleepiness all appear in mild form when humans are exposed to between 100
and 200 ppm (678 and 1356 mg/m ) PCE. Higher exposure concentrations produce
intensified symptoms leading to confusion and nausea. Feelings of motor
incoordination and faintness occur between 600 and 1000 ppm (4068 and 6780
mg/m ). The data are too scant to hazard a generalization about the durations
required to produce symptoms. The above symptoms were reported at times.
ranging from 10 minutes to 2 hours. All symptoms due to experimental exposures
6-5
-------
were reversed in a matter of hours after cessation of exposure. These conclu-
sions should be treated with extreme skepticism, however, due to the extremely
small data base, the few subjects studied, and the largely qualitative methods
employed.
6.1.4.2 Effects of Long-Term Exposures—Data about long-term (greater than a
few weeks) exposure to PCE in humans originate entirely from case studies
(Grossdorfer, 1952; Coler and Rossmiller, 1953; Lob, 1957; Gold, 1969; Weichardt
et al., 1975). Extended exposure to PCE appears to make many of the temporary
symptoms produced by short-term exposure more continuous or of a longer dura-
tion. Reports of frequent dizziness, headache, nausea, fatigue, and disorien-
tation are common even for extended periods of time after cessation of exposure.
New symptoms which have no clear analogs in the short-term case appear, however,
during long-term exposure. Long-term exposed subjects are reported to have
short-term memory deficits, ataxia, irritability, disorientation, sleep distur-
bances and decreased alcohol tolerance. Such symptoms are sometimes reported
to be irreversible.
Stewart et al. (1970) and Stewart et al. (1977) performed repeated experi-
mental exposures to 100 ppm (678 mg/m ) PCE. While these were not long-term
studies with respect to frequent industrial exposures, these studies could
provide data about cumulative effects and adaptation over a moderate period of
time. Subjective reports indicated that perceived odor intensity decreased
over repeated exposures as well as within days. Subjective complaints also
decreased over the course of repeated exposures. No systematic trend was
found in objective test results (see Section 6.1.4.1 for details), and indeed
it may be argued that no effects on objective test results could be demonstrated
due to PCE exposure at all.
Stewart et al. (1977) conjectured that repeated PCE exposure effects
might be exacerbated by simultaneous administration of either alcohol or
Diazepam. While both alcohol and Diazepam produced decrements in the various
objective tests (see Section 6.1.4.1 for details), the effects were no worse
when these substances were combined with PCE exposure. It would appear that
repeated 100-ppm (678 mg/m ) PCE exposure is not close to the threshold for
objective test effects.
There is only limited useful information concerning the effects of long-
term exposure to PCE. Case studies provide poor exposure data and are poten-
tially confounded with exposure to other similar substances which would be
expected to produce similar results, e.g., Tuttle et al. (1977). Control
6-6
-------
groups are frequently absent In such studies as well. From repeated exposures
in a laboratory setting (Stewart et al., 1970; Stewart et al. , 1977) it is
apparent that some sensory and symptom adaptation to low-level PCE exposure
occurs, but the extent to which such adaptation occurs at higher levels and
the sequelae of such adaptation are not known.
6.2 LABORATORY ANIMAL STUDIES
Reported effects associated with PCE exposure of laboratory animals include
effects on the CMS, cardiovascular system, liver, kidney, skin, eyes, and the
immune system. A summary of effects and dose data are provided in Tables 6-1
and 6-2.
6.2.1 Lethality and Anesthesia
The lethality and anesthesia levels for single exposures to PCE in rats
were studied by Carpenter (1937) and Rowe et al. (1951). Data reported by the
above articles were used to construct the curves of Figure 6-1. Linear regres-
sion analyses were used to predict the time required to produce an effect for
various concentrations of PCE in a log-log plot. The line labeled "LC100"
represents the time required to just produce death at various concentrations
of PCE. The line labeled "ANESTHESIA" is the time required to just produce
anesthesia at various concentrations of PCE. The "LCD" line is the prediction
line for the time required to produce effects just short of death at various
concentrations of PCE. In order to depict the time by concentration curve for
anesthesia in linear space, Figure 6-2 was plotted. It appears that anesthesia
may be produced in rats by sufficiently long single exposures even by concen-
trations of PCE as low as 2500 ppm (16,950 mg/m ), even though nearly 16 hours
of exposure is required. While the curves in Figures 6-1 and 6-2 are probably
of the correct form, the exact values are likely to be in error. Data were
fitted to reported means from groups of unequal size. There is likely to be
variation due to strain of rats. Log transformed data were antilog transformed
to produce Figure 6-2, a process which does not necessarily produce a least
squares fit in linear space.
Repeated exposure to higher concentrations of PCE appear to produce con-
siderable tolerance in rats. Carpenter (1937) reported that 2,750 ppm (18,645
o
mg/m ) PCE produced anesthesia in rats on the first exposure day. After six
such exposures, however, rats did not become anesthetized even though 10,000
ppm (67,800 mg/m ) exposures were tried. This finding corresponds to reports
6-7
-------
TABLE 6-1. SUMMARY OF THE EFFECTS OF PCE ON ANIMALS
Animal Dose
species concentration
Route of
administration
Exposure variables
Effects
Reference
Rabbit N/A
(female)
Rabbit
Mouse
i
CO
Mouse
Guinea
pig
Guinea
pig
Guinea
pig
13 mmol/kg
200 ppm
200 ppm
100 ppm
200 ppm
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
Duprat 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
administration
Exposure variables
Effects
Reference
Guinea
pig
Rabbit
Rat
Monkey
Rabbit
01
Rat
Rat
Rat
Rabbit
Rabbit
2500 ppm
inhalation
100-400 ppm inhalation
2500 ppm
2500 ppm
1600 ppm
3000-6000
inhalation
inhalation
inhalation
inhalation
15 ppm
2212 ppm
(15 mg/L)
inhalation
inhalation
18 7-hour
exposures
7 hours/day
5 days/week
6 months
28 7-hour
exposures
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-
1ium 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 liyer weight, in-
crease in total lipid content
of liver accompanied by a few
.diffusely distributed fat
globules
depressed agglutinin
formation
1iver damage
indicated by elevated
SGPT, SGOT, SGLDH:
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
Mazza, 1972
Mazza, 1972
-------
TABLE 6-1. (continued)
Animal Dose Route of
species concentration administration Exposure variables
Rat 70 ppm inhalation 8 hours/day
5 days/week
150 exposures
(7 months)
Rat 230 ppm inhalation 8 hours/day
5 days/week
150 exposures
(7 months)
Rat 470 ppm inhalation 8 hours/day
5 days/week
150 exposures
(7 months)
en
o
Rat 2750-9000 inhalation single exposure
Rat 19000 ppm inhalation 30-60 minutes
Rabbit 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 urobilinogen,
Rabbit
2211 ppm
(15 mg/L)
inhalation
45 days
pathomorphologi cal
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)
Animal Dose
species concentration
Mouse 2.5 ml/kg
(Swiss
male,
10 animal's)
Mouse 5. 0 mL/kg
(10
Animals)
Rabbit 2211 ppm
(15 rng/L)
Mouse N/A
Mouse N/A
Dog
Dog
Dog
Rat 300 ppm
Route of
administration Exposure variables
intraperitoneal
intraperitoneal
(urine samples were collected 24 hours
inhalation 45 days
intraperitoneal
intraperitoneal
intraperitoneal
intraperitoneal
intraperitoneal
inhalation 7 hours/day
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
LD50
elevated SGPT
caused phenolsulfo-
nephthalein retention
indicating kidney dysfunction
LD50
decreased maternal
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 . ,
days 6-15 of
gestation
weight gains,
increased fetal
reabsorptions
1975
-------
TABLE 6-1. (continued)
Animal
species
Mouse
Dose Route of
concentration administration Exposure variables
300 ppm inhalation '-7 hours/day
days 6-15 of
gestation
Effects Reference
maternal liver weights Schwetz et al.,
increased relative to 1975
body weight; increased
incidences of fetal
subcutaneous edema,
delayed ossification of
skull bones, and split
sternebrae
en
i
Rat
Mouse
Rat
Rat
Dog
(male
beagles)
44.2 ppm
15-74 ppm
15 ppm
73 and
147 ppm
0.5-1.0%
v/v
5000 &
10000 ppm
inhalation entire gestation
period
inhalation 5 hours/day
3 months
inhalation 4 hours/day
5 months
inhalation 4 hours/day
4 weeks
inhalation 7 minutes house air
followed by 10
minutes of tetrachloro-
ethylene 8 ug/kg
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
decreased levels of DNA
and total nucleic acids
in the liver, brain,
ovaries, and placenta
decreased electroconductance
of muscle and "amplitude"
of muscular contraction
EEG changes and proto-
plasmal swelling of
cerebral cortical cells,
some vacuolated cells and
signs of karyolysis
EEG and electromyogram
changes; decreased
acetylcholi nesterase
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,
Aninina, 1972
Dmitrieva, 1968
Dmitrieva, 1966
Dmitrieva, 1966
Reinhardt et al., 1973
-------
TABLE 6-1. (continued)
cr>
i
Animal
species
Cat
Cat
Mouse
Mouse
Rabbit
Cat
Dog
Dose
concentration
3000 ppm
14,600 ppm
40 mg/L
5,900 ppm
4-5 mL/kg
5 ml/kg
4 mg/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
01
I
Description
of exposure3
LDEO
LD50
ED50
EDso
LD50
ED50
ED50
LD50
ED50
LD50
LD50
LD50
LCLo
LD50
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
8100 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 mg/kg
0.74 ml/kg
7.2 mM/kg
1.4 mL/kg
390 mM/kg
27 mM/kg
0.109 ml
0.134 ml
8850 mg/kg
23000 mg/m3
Toxic effect
endpoint
death
death
1 iver
dysfunction
death
liver toxicity
death
liver damage
ki d'fie'y
dysfunction
death
liver toxicity
death
death
death
death
death
Time Reference
36 hr Wenzel and Gibson, 1951
24 hr Klaassen and Plaa, 1966
24 hr Gehring, 1968
24 hr Klaassen and Plaa, 1967
24 hr Klaassen and Plaa, 1967
24 hr Klaassen and Plaa, 1967
10 days Plaa et al. , 1958
Plaa et al. , 1958
unknown Dybing and Dybing, 1946
unknown Dybing and Dybing, 1946
unknown Handbook of Toxicology, 1959
2 hr
Withey and Hall, 1975
-------
TABLE 6-2. (continued)
Description
of exposure3
LCLo
LCLo
LDLo
LDLo
LDLo
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 mg/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 LC50, in air which has been reported to have caused death in
humans or animals.
LDL - the lowest dose of a substance, other than an LD50, that is introduced by any route other than inhalation over any
givin period of time and reported to have caused death in humans or animals introduced in one or more divided
portions.
-------
1000
O LCD (ROWE etal. 1951)
D ANESTHESIA (CARPENTER. 1937)
A LC100 (Rows et al. 1951)
10
345 10
PCE CONCENTRATION (x 103), ppm
Figure 6-1. Concentration-time curves for various effects of PCE
Source: Carpenter (1937) and Rowe et al (1951).
30 40
6-}6
-------
1000
5 10 15
PCE CONCENTRATION (x103), ppm
Figure 6-2. Concentration-time curve for PCE-induced anesthesia.
20
6-17
-------
of tolerance in humans (Stewart et al., 1970; Stewart et al., 1977). Behavioral
adaptation, although in a less obviously systematic way, has also been reported
in rats by Rowe et al. (1951) and Goldberg et al. (1964) (see Section 2.2.2).
Effects of high-level PCE exposure in rats appeared to be similar to
those of other anesthetics. When death occurred due to a single PCE exposure,
the process was highly similar to narcotic overdose (Carpenter, 1937). Although
rats dying from repeated exposures to high-level PCE showed some liver involve-
ment, the death was still attributable to CNS impairment (Rowe et al., 1951).
6.2.2 Behavioral Effects
Behavioral effects of high-level and repeated exposures were reported
qualitatively by Carpenter (1937) and Rowe et al. (1952). Most often noted
were ataxia, somnolence, and eventually anesthesia. A group of 18 rats exposed
to 1600 ppm (10,848 mg/m ) PCE for 7 hr/day for 18 days were reported to be
stuporous during the first week and to have shown increased salivation, motor
activity, "reflexive" aggression and some ataxia, but the latter symptoms
subsided toward the end of the second week (Rowe et al., 1951). Brief recur-
rence of the symptoms was reported at the beginning of each week, but overall
reduction was a function of repeated exposures. Some of the symptoms are
reminiscent of those reported in high-level, long-term industrial exposure of
humans (see Section 2.1.4.2).
Goldberg et al. (1964) studied the effects of 0, 1150, and 2300 ppm (0,
3
7797, and 15,594 mg/m ) PCE exposure for 4 hr/day, 5 days/week for 2 weeks, in
groups of 8 to 10 rats. Rats were tested before and after exposure on an
escape/avoidance task in which the response was pole climbing. An 80 percent
loss of both escape and avoidance responses due to ataxia was reported after
3
the first 2300 ppm (15,594 mg/m ) exposure period. Upon repeated exposures,
however, the effect was diminished. No data were presented on the adaptation
effect other than that it occurred.
Low-level repeated PCE exposure (200 ppm, 6 hr/day, for 4 days) was
studied in groups of 10 rats by Savolainen et al. (1977). Many dependent
measures were collected, including brain protein, RNA, glutathione, acid
proteinase, and nonspecific cholinesterase and six aspects of open field
behavior. The behavioral variables were collected and a separate significance
test was done on each measure after the last exposure day and 17 hours later.
Of the 12 t-tests performed, only the preening rate and preening time were
significant and only immediately after the last exposure. The authors also
6-18
-------
discussed reduced RNA and increased nonspecific cholinesterase content in brain
due to PCE exposure. If these results are replicable, they represent a near
threshold effect of PCE in rats.
Ataxia and somnolence due to PCE ingestion have been reported in dogs
given 0.22 mg/kg for worm treatment (Snow, 1973; Myer and Jones, 1954).
However, such side effects are not universally found at this dosage (Miller,
1966; Lamson et al., 1929; Schlingman, 1925). Occurrence of neurotoxic effects
probably depends upon the extent of absorption of PCE via ingestions which is
in turn likely to be affected by diet (Snow, 1973) as well as by the strain of
dog. In the absence of tissue levels, ingestion data are difficult to interpret,
Extremely low levels of PCE exposure were reported to have affected
electrophysiological responses in rats (Dmitrieva, 1966, 1968, 1973; Dmitrieva
et al., 1968). Documentation is sparse, and the effective dose levels are so
low compared to the rest of the literature that the credibility of the reports
must be questioned until independently replicated.
6.2.3 Effects on the Liver and Kidney
PCE is generally regarded as being both hepatotoxic and nephrotoxic in
laboratory animals 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-mo 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
reveal any damaging effects in any of the chronically exposed rats. Functional
6-19
-------
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-hr 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-hr 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),
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.
6-20
-------
Klaassen and Plaa (1967) showed that near-lethal i.p. doses of PCE in
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-hr LD50 value (lethal dose in 50
percent of the animals treated). The ED50 values were measured by sulfobromo-
phthalein, SGPT, glucose, protein, and phenolsulfonephthalein (PSP), indica-
tors of liver or kidney dysfunction. The LD50 was 2.1 ml/kg (21 mmol/kg)
and the ED50 for elevation of SGPT was 0.74 mL/kg (7.4 mmol/kg). The ED50 for
diminution of PSP excretion was 1.4 ml/kg (14 mmol/kg). After administration,
effects on the liver and kidneys were determined by microscopic examination.
At near-lethal doses, PCE produced moderate neutrophilic infiltrations in the
sinusoids and portal areas; necrosis was not observed. Near the ED5Q, vacuoli-
zation of centralobular hepatocytes in about half the animals was observed.
After a single i.p. injection at 0.75 x LD50, 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
hr after interperitoneal injection). At near PSP excretion--ED5Q 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 LD5Q was 2.9 ml/kg (28 mmol/kg). The ED5Q for BSP
retention was 2.9 mL/kg (32 mmol/kg). The ED5Q for elevation of SGPT was 2.9
mL/kg (28 mmol/kg). When ethanol (5 g/kg) was administered by gavage for 3
days prior to injection of 1.0 mL/kg PCE, neither PSP excretion nor BSP reten-
tion was significantly altered. At ESP-ED™, mice exhibited predominantly
inflammatory changes with trace to marked quantities of lipid accumulation.
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.
Effects upon the livers of rats exposed continuously for 3 months to 0.7
o
and 2.7 ppm (4.5 and 19 mg/m ) were reported by Bonashevskaya (1977). No
significant histomorphological changes in the liver were reported at either
6-21
-------
exposure level. Slight alterations in succinic dehydrogenase and glucose-6-
phosphate dehydrogenase were reported.
Kyi in et al. (1963) noted moderate fatty degeneration of the liver with a
3
single 4-hr exposure to 200 ppm PCE (1356 mg/m ) 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/m3) for 4 hr.
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 ). Kyi in et al. (1965) exposed four groups of 20 albino mice to
200 ppm PCE (1356 mg/m ). Each group of 20 mice 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 upon histologic examination.
Mazza (1972) exposed 15 male rabbits for 4 hr/day, 5 days a week, for 45
days, to 2790 ppm PCE (18,924 mg/m3). Mazza looked at the effect of PCE on
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 cytoplasm!c injury in the liver.
6-22
-------
Mazza and Brancaccio (1971) exposed 10 rabbits for 4 hr/day, 5 days per
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
hr/day, 6 days/week, for 45 days, to 2280 ppm PCE (15,465 mg/m3) to look at
effects on kidney function. They noted a reduction in glomerular filtration
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 i.p. injection. Ten mice
received 2.5 mg/kg and 10 others received 5.0 mg/kg. Urine samples were col-
lected 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 micro-
scopically. The proximal convoluted tubules were swollen in all animals and
necrotic in one.
Cornish et al. (1973) administered from 0.3 to 2.0 ml PCE/kg i.p. to
rats. SGOT levels were elevated at all dose levels; phenobarbital pretreatment
did not alter the response.
Fujii (1975) observed an increase in serum enzyme activities (i.e.,
alkaline phosphatase, SGOT, and SGPT) within 24 hours after a single dose of
13 mmol PCE/kg given orally to rabbits. These changes in serum enzyme activi-
ties, indicative of liver damage, were mild and transient.
6.2.4 Effects on the Heart
The effects of PCE upon the heart have not been systematically investi-
gated. Kobayashi et al. (1982) demonstrated in rabbits under urethane anes-
thesia and in cats and dogs under pentobarbital anesthesia that PCE can sensi-
tize the myocardium to epinephrine, leading to ventricular arrhythmias and
6-23
-------
premature contractions, bigeminal rhythm, and tachycardia. Rabbits were the
most sensitive; doses of 0.6 |jg epinephrine per kg and 5 mg PCE per kg led to
tachycardia. In dogs, 20 to 40 mg PCE per kg, i.v., produced significant
depressions in the rate of rise of left intraventricular pressure with the
absence of significant effect upon arterial pressure. The mean lethal dose in
cats was 81.4 ± 14.4 mg PCE per kg, i.v. As the authors noted, and as described
in the preceding chapter, the absence of cardiac effects in humans in the
literature may be largely due to the limitations on absorption of PCE by the
lungs. Additionally, the levels to which individuals are ordinarily exposed
both environmentally and occupationally are obviously below the threshold for
arrhythmia production via endogenously produced epinephrine. Experiments on
unanesthetized laboratory animals involving tasks designed to elicit endogenous
epinephrine in combination with vapor levels of PCE should produce more useful
information on the cardiotoxicity of PCE.
A lack of cardiotoxicity of PCE was reported by Reinhardt et al. (1973).
Unanesthetized dogs (17) were exposed to PCE in air at levels of 5,000 or
3
10,000 ppm (33,915 or 67,830 mg/m ). No positive responses were observed. In
this study, a response considered indicative of cardiac sensitization was the
development of a life-threatening arrhythmia or cardiac arrest following a
challenge of epinephrine.
Christensen and Lynch (1933) observed depression of the heart in 5 dogs
at near-LDgQ oral doses.
6.2.5 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 acute or 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.
6-24
-------
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
available generally provide subjective information on effects resulting from
short-term exposures at levels near 100 ppm (678 mg/m3). The observations in
these studies follow the expected pattern for a non-specific anesthetic effect.
Effects associated with chronic exposures, on the other hand, are available
from animal experiments.
In humans, short-term effects of PCE exposure seem to occur beginning at
100 to 200 ppm (678 to 1356 mg/m3). Such effects include dizziness, confusion,
nausea, headache, and irritation of the eyes and mucuous tissue. Higher
levels of exposure increase the symptoms and eventually produce faintness and
unconsciousness. Long-term exposures are reported to make the above symptoms
more pervasive and also to affect short-term memory and produce disorientation,
irritability, ataxia, and sleep disturbance. These conclusions are based on
largely anecdotal data in field studies or limited laboratory experiments.
Due to the paucity of experiments, the poor quality of the extant data base
and the age of the data base, these conclusions should be viewed with consid-
erable skepticism. Newer, more sensitive methods, better experimental control,
more quantitative analysis, and a wider range of dependent variables could
demonstrate effects at lower levels.
While suffering from the same deficiencies as those outlined for the data
on humans and PCE exposure, data from laboratory animals are remarkably conso-
nant with human effects. Effects on open field activity have been reported at
200 ppm (1356 mg/m3) exposure levels. Effects of higher level exposure seem
to relate well to the findings in humans. It appears that the rat is a good
behavioral model for PCE effects in man.
Investigations in rats have likened the effects of PCE to those of narcotic
drugs and anesthetics. The lowest effects levels for PCE seem to be in CNS or
behavioral variables. This seems to be true for either single or repeated
long-term exposures. Similar conclusions are reached in a review by Annau
(1981).
Effects upon the human kidney and/or liver are not well documented and
appear to be confined to those exposure scenarios in which PCE levels are
6-25
-------
exceedingly high and exposures are short in duration. Studies of experimental
animals suggest that reversible changes (congestion, fatty infiltration) occur
at exposure levels of about 200 ppm (Carpenter, 1937; Rowe et al. , 1952; Kylin
et al., 1963). The relevance of these data to the likelihood of similar
effects in humans at similar levels is unknown.
Limited data involving anesthetized laboratory animals indicate that low
doses of PCE, administered i.v., can sensitize the heart to i.v. challenges of
epinephrine. However, it is not considered likely that PCE will exhibit this
action in humans under exposure conditions normally experienced in the environ-
ment and the workplace.
6.4.2 Oral Exposure
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.
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.
6-26
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6.5 REFERENCES FOR CHAPTER 6
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Coler, H. R., and H. R. Rossmiller. 1953. Tetrachloroethylene exposure in a
small industry. Ind. Hyg. Occup. Med. 8:227.
6-27
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Cornish, H. H., B. P. Ling, and M. L. Barth. 1973. Phenobarbital and organic
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of muscles exposed to chlorinated hydrocarbons. Farmakologiya i
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cortex of rats with the narcotic effect of substances with different
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Dmitrieva, N. V., and E. V. Kuleshov. 1971. Changes in the bioelectric activity
and electric conductivity of the brain in rats chronically poisoned with
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translation)
Dmitrieva, N. V., E. V. Kuleshov, and E. K. Orjonikidze. 1968. Changes in the
impedance and bioelectrical activity of the cerebral cortex of rats under
the action of anesthetic drugs. Zhur. vysshei Nervnoi Deyatel 'nosti
18(3):463-468. (English translation)
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resistance of the organism to tetrachloroethylene. Revista Igiena Bact.
Virus Parazit. Epidem. Pneum. Igiena 24(3):147-151. (translation).
Dumortier, L., G. Nicolas, and F. Nicolas. 1964. A case of hepato-nephritis
syndrome due to perchloroethylene. Arch. Mai. Prof. 25:519-522 (English
translation)
Duprat, P., L. Delsaut, and D. Gradiski. 1976. Irritant potency of the princi-
pal aliphatic chloride solvents on the skin and ocular mucous membranes
of rabbits. Europ. J. Toxicol. 3:171-177.
Dybing, F., and 0. Dybing. 1946. The toxic effect of tetrachloroethane and
tetrachloroethylene in oily solution. Acta Pharmacol. 2:223-226, 1946.
Eberhardt, H., and K. J. Freundt. 1966. Tetrachloroethylene poisoning. Arch.
Toxikol. (Berlin) 21:338-351. (English translation)
Fishbein, L. 1976. Industrial mutagens and potential mutagens. I. Halogenated
aliphatic derivatives. Mutat. Res. 32:267-308.
Franke, W., and F. Eggeling. 1969. Clinical and statistical studies on em-
ployees of chemical cleaning plants exposed to perchloroethylene. Med.
Welt 9:453-460. (English translation)
Friborska, A. 1969. The phosphatases of peripheral white blood cells in
workers exposed to trichloroethylene and perchloroethylene. Br. Med. J.
Ind. Med. 26:159-161.
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Fujii, T. 1975. The variation in the liver function of rabbits after admini-
stration of chlorinated hydrocarbons. Jpn. J. Ind. Health 17:81-88.
(English translation)
Fuller, B. B. 1976. Air pollution "assessment of tetrachloroethylene." Mitre
Technical Report - 7143. February.
Gehring, P. 1968. Hepatotoxicity of various chlorinated hydrocarbon vapors
relative to their narcotic and lethal properties in mice. Toxicol. Appl.
Pharmacol. 13:287-298.
Gold, J. H. 1969. Chronic perch!oroethylene poisoning. Can. Psychiatric
Assoc. J. 14:627-630.
Goldberg, M. E. , H. E. Johnson, U. C. Pozzani, and H. F. Smyth, Jr. 1964.
Effect of repeated inhalation of vapors of industrial solvents on animal
behavior. I. Evaluation of nine solvent vapors on pole-climb performance
in rats. Am. Ind. Hyg. Assoc. J. 25:369-375.
Hake, C. L., and R. D. Stewart. 1977. Human exposure to tetrachloroethylene:
inhalation and skin contact. Environ. Health Persp. 21:231-238.
Handbook of Toxicology, Volumes II-V, Philadelphia: 1959. W. B. Saunders Co.
Volume V., p. 76.
Hughes, J. P. 1954. Hazardous exposure to a so-called safe solvent. J. Am.
Med. Assoc. 156:234-237.
Klaassen, C. D., and G. L. Plaa. 1966. Relative effects of various chlorinated
hydrocarbons on liver and kidney function in mice. Toxicol. Appl. Phar-
macol. 9:139-151.
Klaassen, C. D., and G. L. Plaa. 1967. Relative effects of various chlorinated
hydrocarbons on liver and kidney function in dogs. Toxicol. Appl.
Pharmacol. 10:119-131.
Kylin, B., I. Sumegi, and S. Yllner. 1965. Hepatotoxicity of inhaled tri-
chloroethylene and tetrachloroethylene - long-term exposure. Acta
Pharmacol. Toxicol. (Kbh). 22:379-385.
Kylin, B., H. Reichard, I. Sumegi, and S. Yllner. 1963. Hepatotoxicity of
inhaled trichloroethylene, tetrachloroethylene, and chloroform—single
exposure. Acta Pharmacol. Toxicol. 20:16-26.
Lamson, P. D. , B. H. Robbins, and C. B. Ward. 1929. The pharmacology and
toxicology of tetrachloroethylene. Am. J. Hyg. 9:430-444.
Larson, N. A., B. Nielsen, and A. Ravin-Nielsen. 1977. Perch!oroethylene
intoxication—a hazard in the use of coin laundries. Ugeskr. Laeg.
39(5):270-275. (English translation)
Lehmann, K. B. 1911. Experimental studies on the influence of technically and
hygienically important gases and vapors on the organism. Arch. Hyg.
74:1-60. (German)
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Lehmann, K. B., and L. Schmidt-Kehl. 1936. The 13 most Important chlorinated
hydrocarbons of the aliphatic series from the standpoint of occupational
hygiene. Arch. Hyg. 116:131-268. (German)
Levine BMP Fierro, S. W. Goza, and J. C. Valentour. 1981. A tetrachloro-
ethylene fatality. J. Forensic Sci. 26(1):206-209.
Ling, S. , and W. A. Lindsay. 1971. Perchloroethylene burns. Brit. Med. J.
3(5766):115.
Lob, M. 1957. Dangers of perchloroethylene. Arch. Gewerbepath. Gewerbehyg.
16:45-52. (English translation)
Mazza, V. 1972. Enzyme changes in experimental tetrachloroethylene intoxica-
tion. Folia Med. 55(9-10):373-381. (English translation)
Mazza, V., and A. Brancaccio. 1971. Adrenal cortical and medullar hormones in
experimental tetrachloroethylene poisoning. Folia Med. 54:204-207.
(English translation)
Meckler, L. C., and D. K. Phelps. 1966. Liver disease secondary to tetra-
chloroethylene exposure. J. Am. Med. Assoc. 197(8):144-145.
Method, H. C. 1946 Toxicity of tetrachloroethylene. J. Am. Med. Assoc.
131:1468.
Miller, T. A. 1966. Anthelmintic activity of tetrachloroethylene against
various stages of Ancylostoma cam'urn in young dogs. Am. J. Vet. Res.
27(119): 1037-1040.
Moeschlin, S. 1965. Poisons, diagnosis and treatment. New York, N.Y.:
Gruner and Stratton.
Morgan, B. 1969. Dangers of perchloroethylene. Brit. Med. J. 2:513.
National Institute for Occupational Safety and Health. 1976. Criteria for a
recommended standard...occupational exposure to tetrachloroethylene
(perchloroethylene). HEW publication no. (NIOSH) 76-185. U.S. Department
of HEW, PHS, CDC, NIOSH. July.
Navrotskii, V. K., Kashin, I. L. Kulinskaya, L. F. Mikhaylovskaya, L. M.
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assessment of the toxicity of a number of industrial poisons when inhaled
in low concentrations for prolonged periods. Trudy S'ezda Gigen Vkramkoi:
SSR 8:224-226. (translation)
Parker, J. C., L. J. Bahlman, N. A. Feidel, H. P. Stein, A. W. Thomas, B. S.
Woolf, and E. J. Baier. 1978. Tetrachloroethylene (perchloroethylene).
Current NIOSH Intelligence Bulletin #20. Am. Ind. Hyg. Assoc. J. 39:3.
Patel, R., N. Janakiraman, and W. D. Towne. 1977. Pulmonary edema due to
tetrachloroethylene. Environ. Health Persp. 21:247-249.
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Plaa, G. L., and R. E. Larson. 1965. Relative nephrotoxic properties of
chlorinated methane, ethane, and ethylene derivatives in mice. Toxicol.
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"Plaa, G. L., E. A. Evans, and C. H. Mine. 1958. Relative hepatotoxicity of
seven halogenated hydrocarbons. J. Pharmacol. Expt. Ther. 123:224-229.
Reinhardt, C. F., L. S. Mullin, and M. B. Maxfield. 1973. Epinephrine-induced
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Occup. Med. 15:953-955.
Rowe, V. K., D. D. McCollister, H. C. Spencer, E. M. Adams, and D. D. Irish.
1952. Vapor toxicity of tetrachloroethylene for laboratory animals and
human subjects. Arch. Ind. Hyg. Occup. Med. 5:566-579.
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Med. 67:2359-2361.
Savolainen, H., P. Pfaffli, M. Tegen, and H. Vainio. 1977. Biochemical and
behavioral effects of inhalation exposure to tetrachloroethylene and
dichloromethane. J. Neuropathol. Exp. Neurol. 36:941-949.
Schlingman, A. S. 1925. Critical tests of tetrachloroethylene: a new anthel-
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Schwetz, B. A., B. K. Leong, and P. J. Gehring. 1975. The effect of maternally
inhaled trichloroethylene, perchloroethylene, methyl chloroform, and
methylene chloride on embryonal and fetal development in mice and rats.
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Assoc. 208(8):1490-1492.
Stewart, R. D., and H. C. Dodd. 1964. Absorption of carbon tetrachloride,
trichloroethylene, tetrachloroethylene, methylene chloride, and 1,1,1-
trichloroethane through the human skin. Am. Ind. Hyg. Assoc. J. 25:
439-446.
Stewart, R. D., D. S. Erley, A. W. Schaffer, and H. H. Gay. 1961a. Accidental
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Stewart, R. D., E. D. Baretta, H. C. Dodd, and T. R. Torkelson. 1970. Experi-
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(English translation)
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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 "teratogenic11 is defined as the tendency to
produce physical and/or functional defects in offspring HI 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 carcinogenidty) 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.
7-1
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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
® 3
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. Seventeen 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 sporadic 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 tera-
togenicity 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
7-2
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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. Half
of the fetuses in each litter were examined for soft tissue malformations and
the remaining half of the fetuses were examined for skeletal malformations.
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.
Bellies et al. (1980) exposed Sprague-Dawley rats to PCE at 300 ppm
(2034 mg/m3) for 7 hours daily, 5 days per week. Controls were administered
filtered air. Nineteen to 24 rats were examined in this study. One-half of
the rats were exposed for 3 weeks prior to mating. All rats were exposed
during gestation either between days 0 through 18 or between days 6 through
18, with day 0 designated as the day when spermatozoa were observed in smears
of vaginal contents. Three rats (days 6-18) died on the second day of preges-
tational 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
end of the exposure showed 568 ppm (4061 mg/m3) 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 malforma-
tion as such. No teratogenic effects were observed.
7-3
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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
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
7-4
-------
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
3
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 thus 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 1n 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 was 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 embryotoxlcity. Elevation in ethoxycoumarein dealkylase activity (an
indicator of P.™ activity) was observed in maternal livers but not fetal
livers with pregnancy exposure. Ethoxyresorufin dealkylase activity (an
indicator of P,4g activity) was not elevated in maternal livers with PCE
exposure and was not detectable in fetal livers. No other effects were ob-
served in pregnant rats exposed for 6 hr/day on days 0-20 of gestation.
Manson et al. (unpublished, 1982) conducted a study on postnatal evaluation
of offspring of the female rats exposed to 1000 ± 125 ppm (6780 ± 847 mg/m )
7-5
-------
PCE before and/or during pregnancy. The four treatment groups are those
described above (Tepe et al., 1982). The purpose of the postnatal evaluation
was to determine 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 activity 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. Fifteen to 22 rabbits were used in this study. The rabbits were
divided into six groups as follows:
Days of Exposure
Group Cone (ppm) PregestationalGestational
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, 4, 5, and 6) to 500 ppm (3390 mg/m ) PCE, 7 hours per day.
Mean body weights of rabbits during pregestational and gestational expo-
sure 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
7-6
-------
TABLE 7-1. SUMMARY OF REPRODUCTIVE/TERATOGENIC EFFECTS OF PCE IN LABORATORY ANIMALS
Study
TCI Purity
Species
Experimental
Conditions
Study
Results
Beliles et al. ,
1980 *
Hanson et al. ,
1982
Technical grade
91.43%
CD rats, New
Zealand rabbits
Elovaara et al., 99% pure
1979 dissolved in
olive oil
Technical grade
(Dow Chemical
Company)
Nelson et al., Technical grade
1980 98.5% pure
White Leghorn
chick embryos
Long-Evans
hooded rats
Sprague-Dawley
rats
Schwetz et al., Dow-Per
1975 99.99% pure
Sprague-Dawley
rats, Swiss-
Webster mice
Inhalation, 500 ppm
before, during
gestation, and both
before and after
gestation.
Injection into air
space 5 to 10 umole/
egg on days 2,3,6 of
incubation, examined 14
days after incubation.
Inhalation, 1000 ppm
before, during
gestation, and both
before and during
gestation.
Inhalation, 100 or
900 ppm on various
days of gestation.
Inhalation, 300 ppm
on days 6-15 of
gestation.
Teratology
study
Chick egg
teratology
Postnatal
function
evaluations
Postnatal
behavioral
testing
Teratology
study
Rats maternal
toxicity, changes in
liver weight, embryo
toxicity, reduced
ossification; rabbits:
variations in size
and color of rabbit
placenta.
Malformed embryos at
highest dose, de-
creased embryo length.
No detrimental
effects in off-
spring observed up
until 8 months
after birth.
Maternal weight
loss at highest
dose, slight,
subtle behavioral
changes in off-
spring.
Maternal toxicity,
embryo toxicity
(lowered weight of
mice, increased re-
sorption in rats,
subcutaneous edema
in mice offspring).
-------
I
oo
TABLE 7-1. (Continued)
Study
Tepe et al . ,
1982
TCI Purity
Technical grade
(Dow Chemical
Company)
Species
Long-Evans
hooded rats
Experimental
Conditions
Inhalation, 1000 ppm
before, during
gestation, and both
before and during
gestation.
Study
Teratology
study
Results
Maternal toxicity,
embryo toxicity,
depressed fetal
weight, skeletal
and soft tissue
variations.
-------
exposure. Placenta! abnormalities were reported at all exposure levels (varia-
tion in size and color); however, these were not statistically significant as
tested by rank sum analysis and were thought to reflect changes in a few
litters. In addition, histopathologic evaluation failed to reveal any signifi-
cant change in the placenta.
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 primarily reflect delayed development and generally
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. PCE also has been implicated in producing
adverse germ cell effects with alteration in sperm morphology (see Section
8.1). The potential of PCE to cause adverse effects on reproduction or the
developing conceptus is based on a limited number of studies. The teratogenic
potential of PCE for humans is unknown.
7-9
-------
7.2 REFERENCES
Bellies, R. P.; Brusick, D. J.; Mecler, F. J. 1980. Teratogenic-mutagenic
risk of workplace contaminants: trichloroethylene, perchloroethylene, 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. 1970. Studies on placental transfer: trichloroethylene. Ind. Med.
39:46-49.
Manson, J. M. ; Tepe, S. J. ; Lowrey, B.; L. Hastings. 1982. Postnatal evalua-
tion 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.; Dorfmueller, 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. FR (1978 August 22) 43:37382-37388.
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. FR (1979 July 26) 44:44089-44092.
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. FR (1980 December
12) 45:81869-81876.
7-10
-------
8. MUTAGENICITY
The objective of this mutagenicity evaluation is to determine whether PCE
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 PCE to cause gene mutations in bacteria has been studied
by several investigators. Many of these investigators used the Ames Salmonella/
microsome test or modifications of that test. Different purities of PCE
(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. 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.
"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
Ames/Salmonella assay
Results*
+** _
Escherichia coli K12/343/113 ( )
Multi-purpose test
Saccharomyces cerevisiae D7 reverse
mutation test (ivl-1 locus)
Drosophila 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
Drosophila sex chromosome loss
assay
Rat dominant lethal assay
References
Williams and Shimada 1983,
Margard 1978, SRI Inter-
national 1983, NTP 1983,
Bartsch et al. 1979, Cerna
and Kypenova 1977 (abstract)
Henschler 1977, Greim et al.
1975
Call en et al. 1980,
Bronzetti et al. 1983
Beliles et al. 1980
Beliles et al. 1980,
Cerna and Kypenova 1977
(abstract)
Bronzetti et al. 1983
Rampy et al. 1978,
Beliles et al. 1980
Cerna and Kypenova
1977 (abstract)
Ikeda et al. 1980
Beliles et al. 1980
Beliles et al. 1980
C. Other Tests Indicative of DNA Damaging Activity
Unscheduled DNA synthesis in WI-38 -"*"
Hepatocyte primary culture/DMA repair +,•
test
Mitotic recombination tests in
Saccharomyces cerevisiae D7
Beliles et al. 1980
Williams and Shimada
1983, Williams 1983
Call en et al. 1980,
Bronzetti et al. 1983
D. DNA Binding Studies
Whole mice
E. Germ Cell Tests
Altered sperm morphology
mouse
+
rat
Schumann et al. 1980
Beliles et al. 1980
* + designates positive; negative; wk weak response. Dose-response
relationships were not established for the reported + results or wk results.
**Although 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).
TQuestionable evidence for weak or borderline activity in specific data sets.
ttPositive 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
Test system/strain Purity/source
Ames/Salmonella Perchlor 200-
TA98, TA100, TA1535 low-stabil ized
99.93% purity
PPG Industries,
Inc.
m Ames/Salmonella Perchlor 230-
' TA98, TA100, TA1535 stabilized
TA1537, TA1538 99.80% purity
PPG Industries,
Inc.
Ames/Salmonella High purity
TA100, TA1535 Perchlor
low-stabil ized,
99.98+% purity,
PPG Industries,
Inc.
Concentrations
tested
1% v/v for TA98,
TA1538, and
TA1537; 0.1
1.0, 2.5, 5.0,
7.5, and 10%
for TA100 and
TA1535
1% v/v for TA98,
TA1538, and
TA1537; 0.1
1.0, 2.5, 5.0,
7.5, and 10%
for TA100 and
TA1535
0.1, 1.0, and
2.5% v/v
Metabol ic
activation
Aroclor-
induced rat
S9 mix
Aroclor-
induced rat
S9 mix
Aroclor-
induced rat
S9 mix
Protocol
Gas-phase
exposure in
airtight
chambers
Gas-phase
exposure in
airtight
chambers
Gas-phase
exposure
in airtight
chambers
Reported
result Reference
Positive Williams
at 2.5% (>97% and
toxicity) in Shimada
base-pair 1983
substitution-
sensitive strain
(two to tenfold
i ncreases
+/- M.A.*)
dose-response
not established
Positive Williams
at 2.5% (>97% and
toxicity) Shimada
in base pair 1983
substitution
sensitive
strains (three to
ten-fold increase
+/- M.A.)
dose-response
not established
Negative Williams
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/Salmonel la
TA98, TA100, TA1535
TA1537, TA1538
Ames/Salmonel la
TA98, TA100, TA1535
TA1537, TA1538
Purity/source
Nonstabi 1 ized
high purity
Detrex
Industries, Inc.
Stabi 1 ized
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
Metabol ic
activation
Arocl or-
induced rat
S9 mix
Aroclor-
induced rat
S9 mix
Protocol
Standard
plate test
in airtight
chambers
Standard
plate test
in airtight
chambers
Reported
result
Negati ve
Positive
(twofol d
increases
in frameshi
Reference
Margard
1978
Margard
1978
ft-
oo
sensitive
strains and
TA100) at 0.1
ml/plate (160
mg/plate). >90%
toxicity. S9 mix
increased response
(10-to 17-fold
increases)
Ames/Salmonella
TA98, TA100, TA1535
TA1537
99+% purity
Aldrich
0.025, 0.05,
0.1, 0.5, 1.0
and 1.5 added
to petri plate
at bottom of
desiccator
Aroclor-induced
female and male
rat 1iver S9 mix
and Aroclor-in-
duced female and
male mouse liver
S9 mix
Gas-phase
exposure in
airtight
chambers
Negative
SRI
Inter-
national
1983
Ames/Salmonella
TA98, TA100, TA1535
TA1537
Technical grade 3, 10, 33, 100,
Fisher 333 ug/plate
Aroclor-induced
rat liver S9 mix
Aroclor-induced
hamster liver
S9 mix
Preincu-
bation assay
10 min. 37°C
Negative
NTP 1983
(continued on the following page)
-------
TABLE 8-2. (continued)
Test system/strain
Ames/Sal monel la
TA100
Ames/Sal monel la
tester strains
not rep'orted
E. coli K12/343/113 (>)
oo
en
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
(4x10 M)
Not
reported
0.9 mM
Not
reported
Inhalation at
100 ppm and
500 ppm for
5 consecutive
days
Metabolic
activation
Phenobarbital-
induced mouse
1 iver microsomes
with and without
cofactors
Phenobarbital-
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-
peritoneal ly
after last
exposure and
removed 3
hours
Reported
resul t
Negative
Positive
Negative
Positive
(mortal ity
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 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 percent purity), high purity Perchlor (low-stabilized, 99.98+ percent
purity), and Perchlor 230 (stabilized, 99.80 percent purity).* Perchlor 200
and Perchlor 230 were evaluated in tester strains TA98, TA1537, and TA1538 at
1 percent (v/v) and in tester strains TA100 and TA1535 at 0.1, 1.0, 2.5, 5.0,
7.5, and 10 percent (v/v). High-purity Perchlor was evaluated in TA100 and
TA1535 at 0.1, 1.0, and 2.5 percent (v/v). These concentrations represent the
predicted concentration of the compound in the gas phase based upon calcula-
tions 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 PCE was added to the chamber for a 0.1,
1.0, 2.5, 5.0, 7.5, and 5 percent v/v vapor target concentration, respectively.)
Positive responses were obtained for Perchlor 200 and 230 at 2.5 percent v/v
(or 1.55 ml per desiccator) but not for high-purity Perchlor at the concentra-
tions 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 percent) 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 percent v/v) also caused
increases in revertant numbers in TA1535 (approximately tenfold increases) and
TA100 (approximately threefold increases). (See Figure 8-1 for an illustra-
tion of these responses). Negative results were found in frameshift-sensitive
strains (TA98 and TA1538) for both Perchlor 200 and 230 in the presence or
absence of S9 mix. At the next highest concentration (5 percent 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
percent killing) at the concentration (i.e., 2.5 percent v/v) that induced the
mutagenic responses as determined by a simultaneous cytotoxicity test. Although
*Perchlor 200 contained 0.012 (percent by weight) of hydroquinone mono-
methyl ether (HQMME) which provides minimal stabilization; Perchlor 230
contained 0.011 percent HQMME, 0.07 percent cyclohexene oxide, and 0.05 percent
B-ethoxy propiom'trile; high-purity Perchlor contained 0.01 percent HQMME
(written communication from Dr. A. Philip Leber of PPG Industries, Inc., 1983).
8-6
-------
TA 1535
TA1535
-CD
<
o.
cc
LU
CL
I
cc
LU
UJ
DC
m
300
200
100
0
300
200
100
0.1 1.0
2.5
5.0
7.5
10.0
0.1 1.0
TA100
600
500
400
300
200
100
500
400
300
200
100
TA 100
0.1 2.5 5.0 7.5 10.0
TETRACHLOROETHYLENE, Stabilized (%V/V)
0.1 1.0 2.5 ' 50 7.5 10.0
TETRACHLOROETHYLENE, Low Stabilized (%V/V)
Figure 8-1.
Dose-response curves for Perch!or 200 (low-stabilized tetrachloroethylene) and Perchlor 230
(stabiliz tetrachloroethylene) using Salmonella typhimurium tester strains TA100 and TA1535 in the
presence of (o o) and in the absence of (»——•) S9 mix~(150 ul of Aroclor-induced rat 1'ver).
Each data point represents the geometric mean of triplicate plates from one experiment. (Williams
and Shimada 1983)
-------
the test materials may be quite toxic at 2.5 percent 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 toxi-
city, 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 percent)
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 percent 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 PCE (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 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 PCE that contained no detectable epoxides
or other stabilizing components, and the stabilized material was identified as
an industrial degreasing grade of PCE that contained 0.07 percent (by weight)
epichlorohydrin, 0.007 percent N-methylmorpholine, 0.07 percent beta-hydroxy-
pronitrile, and 0.01 percent hydroquinone monoethyl ether (written communica-
tion from L. Schlossberg, Detrex Chemical Industries Inc., 1981). Nonstabil-
ized test material was not detected as positive in the presence or absence of
8-8
-------
*
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 percent
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 International
(1983 EPA-sponsored test) reported PCE (99+ percent purity, Aldrich) 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 PCE (technical grade, source Fischer) and obtained negative results. Four
standard tester strains were used: TA98, TA100, TA1535, and TA1537. A pre-
incubation protocol was followed in which the cells, S9 activation system, and
chemical are preincubated in test tubes for 20 minutes at 37°C
"Calculation based on the density of tetrachloroethylene as 1.586 g/ml.
8-9
-------
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 PCE (99.7 percent
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, PCE (purity and source not reported)
induced both base-pair substitution and frameshift 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, PCE induced significant increases in the number of revertants in tester
strains TA1950, TA1951, and TA1952 at dosage levels reported as representing
the LD^g and one-half of the LDcn- 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 PCE (91.43 percent pure from North Strong Division Chemicals) in the pres-
ence 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 intraperi-
toneally into the mice after the last exposure. The bacteria were removed
8-10
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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
i_n 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 PCE 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 revert-
ants) 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 Greim et al. (1975).
Greim et al. (1975) reported that negative results were obtained when PCE
(purity reported as analytical grade; Merck-Darmstadt) was assayed in the
multi-purpose test system of Escherichia coli K12/343/113 ( ). Tests were
conducted in the absence and presence of metabolic activation (phenobarbital-
induced mouse liver microsomal fraction plus NADPH 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 percent survival. The genetic markers
evaluated for mutation induction were the missense marker arg and the frame-
shift 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 PCE
itself is mutagenic. The positive responses found may be due to contaminants
and/or added stabilizers. The induced increases in revertants did not require
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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
percent v/v and at 160 mg/plate) were needed for the detection of mutagenicity.
Also, there was a very narrow range of effective concentrations because of
toxicity; thus dose-response relationships were not established. When tested,
highly purifed samples were not detected as mutagenic under the conditions in
which technical samples caused increases in 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 PCE are weakly mutagenic
in Salmonella and that the positive responses may be due to impurites and/or
added stabilizers.
8.1.2 Drosophila
Beliles et al. (1980) used the sex-linked recessive lethal assay in
Drosophila melanogaster to test PCE (91.43 percent 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 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 PCE in Drosophila.
The ability of PCE 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 PCE was not properly evaluated.
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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 PCE exposure. Animals were
exposed to 300 ppm (2.03 mg/1) or 600 ppm (4.07 mg/1) PCE* by inhalation 6
hr/day, 5 days/week, for one year. Three animals per dose were examined. The
authors reported "zero" chromosomal aberrations per 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 percent 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 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 PCE.
A rat bone marrow assay was also performed by Beliles et al. (1980) and
reported as negative. Ten males and ten females [CRLCOBS CD(SD)BR] were
exposed to an acute dose of 100 ppm and 500 ppm PCE (91.43 percent 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 percent cells
with aberrations versus 0.7 percent 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 percent cells with aberrations) was found in females. This response
^Formulation (liquid volume percent): trichloroethylene, 3 ppm; hexa-
chloroethane, <12 ppm; carbon tetrachloride, 2 ppm; 4-methyl morpholine, 44
ppm; nonvolatile residue, 2 ppm; and tetrachloroethylene balance.
8-13
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was not dose-related. The female subchronic control group had a very low
background (0.3 percent 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 LD5Q of PCE or dosed intraperi-
toneally for five applications in 24-hour intervals (dose of one injection
equalled 1/6 LD™) 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 in the mitotic index of peripheral lymphocytes
cultured from ten workers (seven males and three females) occupationally
exposed to technical grade PCE (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. 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.
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8.2.3 Drosophila
In addition to performing the sex-linked recessive lethal assay discussed
earlier, Beliles et al. (1980) also conducted a sex chromosome loss assay on
low-purity (91.4 percent, North Strong Division Chemicals) PCE for nondisjunc-
tion in Drosophila melanogaster. Males were exposed to 100 and 500 ppm of
tetrachloroethylene for 7 hours. The phenotypic classes used allow for detec-
tion 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 PCE 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 PCE to cause chromo-
somal aberrations, the data collectively indicates that PCE is not strongly
clastogenic. However, there have been no adequate studies on the ability of
PCE 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. Beliles et al. (1980) assessed the ability
of PCE (91.43 percent 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 source of
metabolic activation (i.e., 59 mix). The test material was examined at 0.01,
0.05, 0.1, and 0.5 percent (v/v) using conventional liquid phase exposure.
Scheduled DNA synthesis was blocked by treatment with hydroxyurea, and UDS was
measured by liquid scintillation counting of incorporated tritiated thymidine
([3H]-TdR) into DNA. A very slight increase was seen at 0.01 percent (v/v)
PCE 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
percent. A toxic response was reported at 0.5 percent (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 percent was found, suggesting that more cells
are entering the S phase. Cells accelerated into S phase would account for
the increases seen in [ H]-TdR incorporation at 0.01 percent. Therefore, it
8-15
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is uncertain whether PCE induced a weak UDS response in this system. Also,
other problems were found in this study. The positive controls gave weak
responses. N-methyl-N1-nitro-N-nitrosoguanidine elicited a weak response
(1.8-fold increase), but at a toxic concentration (5 ug/ml) as indicated by a
decrease in total DNA (6.27 ug of DMA versus 23.76 ug of DMA 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 PCE, Perchlor 200 (low-stabilized, 99.93 percent
purity) and Perchlor 230 (stabilized, purity 99.80 percent), for their ability
to cause UDS in the hepatocyte primary culture (HPC)/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 [ H]-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 PCE 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 percent v/v) and Perchlor 200
(at concentrations of 0.0001, 0.001, 0.1, and 1.0 percent) when tests were
conducted using conventional liquid-phase exposure in which PCE 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 percent v/v (desired air concentration) for 3
and 18 hours. Perchlor 230 at 0.1 percent caused an increase in UDS when the
cells were exposed for 3 hours (6.2 +4.9 net nuclear grain count, 50 percent
cells showing toxic effects) and for 18 hours (15.9 +1.6 net nuclear grain
*Nuclear grain counts were reported as the mean + standard deviation.
Cytoplasmic grain counts in three nuclear size areas adjacent to the nucleus
were determined. The highest cytoplasmic grain count was subtracted from the
nuclear count. This value is referred to as "net" grain count.
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count, 25 percent cells showing toxic effects). Perchlor 200 caused an increase
in UDS at 0.1 percent v/v for the 3-hour treatment (10.8 + 6.1 percent net
nuclear grain count, 75 percent cells showing toxic effects) but not for the
18-hour treatment. At 1.0 percent and 2.5 percent for both materials at both
treatment times, nearly 100 percent of cells showed toxic effects. It should
be pointed out that the background control values were taken from the conven-
tional 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+ percent) 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 BgC3F, mice and male Osborne-Mendel rats on PCE (99+ percent purity,
Aldrich Chemical Company). Conventional liquid-phase exposure conditions were
used. Negative results were reported for both rat and mice hepatocytes when
PCE was added to the culture medium at 0.00001, 0.0001, 0.001, 0.01, and 0.1
percent (v/v) for 18 hours. The material was reported as "toxic" at 0.01
percent 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.
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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8.3.2 MJtotic Recombination
Callen et al. (1980) evaluated the ability of PCE (purity not reported,
stabilized 0.01 percent thymol, Eastman Kodak) to cause mitotic gene conversion
(nonreciprocal recombination) at the trp-5 locus, mitotic crossing-over (reci-
procal 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
PCE. No exogenous source of metabolic activation was used in these studies.
At 4.9 mM (84 percent survival), no significant increases in the frequency of
gene conversion (1.9 convertants/10 survivors versus 1.4 convertants/10
survivors in background control) and mitotic crossing-over (5.3 mitotic recom-
binants/10 survivors versus 3.3 mitotic recombinants/10 survivors in back-
ground control) occurred. As shown in Figure 8-2, when the concentration of
PCE was increased to 6.6 mM (58 percent survival), increases in mitotic gene
conversion and mitotic crossing-over did occur (8.3 convertants/10 survivors
and 52.6 mitotic recombinants/10 survivors, respectively). Mitotic recombina-
tional activity was not determined at 8.2 mM because less than 0.1 percent
survival was found. No significant increases in gene mutations were observed
at 4.9 mM (3.8 revertants/10 survivors versus 2.9 revertants/10 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 the number of trp-5 conver-
tants. 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 recom-
binational events are inducible by a common factor(s) (see Fabre Fabre and
Roman 1977, Fabre 1978). The positive findings of Callen and coworkers for
mitotic recombination should be confirmed by repeating the assay using appro-
priate selection conditions. In addition, because the responses were observed
within a narrow "window," at least one additional concentration between the low
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50
CO
cc
O
en
D
to
o
to
CD
O
O
LU
CC
O
o
40
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 Callen et al. 1980)
8'19
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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 PCE (99.5 percent pure, stabilized with 0.01 percent 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 percent 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 PCE 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 intrasan-
guineous host-mediated assay using S. cerevisiae D7 as the indicator organism
o
and CD-I mice as the host. Stationary yeast cells (4 x 10 cells) were in-
jected into the retro-orbital sinus of mice. After injection of the yeast, an
acute oral dose of 11 g PCE/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 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 PCE (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 DNA repair synthesis in mammalian cells provide suggestive evidence that
certain technical samples of PCE may be active in damaging DNA. However, toxic
concentrations of material were needed to elicit these responses. The possibi-
lity that impurities and/or added stabilizers caused the increased effects
should be considered.
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8.4 DNA BINDING STUDIES
Chemical adduct formation is a critical step in certain types of muta-
genesis. Schumann et al. (1980) reported that there was no detectable bindinq
14
of C-labeled PCE (99 percent 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 C-label (26,593 dpm/umol in the inhala-
tion 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
-5
binding level of 10 alkylations per nucleotide (Stott and Watanabe 1982)
there would be 5-10 dpm per sample. This is at the limit of practical detec-
tion. Therefore, the possibility of binding at slightly less than 10 alkyla-
tions 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 burden. This assessment is almost
always based on animal experimentation. The ability of PCE 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 percent 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 hours per day for 5 days (Bellies
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 DNA.
Beliles et al. (1980) also examined PCE 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
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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 percent abnormal sperm were
observed (versus 6.0 percent 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 PCE alters
germ cell DNA 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 PCE (or impurities) reaches the germ tissue. However,
if PCE reached germ cell DNA, there may be no serious risk of mutation because
of the largely negative or marginal results found in mutagenicity tests dis_
cussed previously.
8.6 MUTAGENICITY OF METABOLITES
Trichloroacetic acid (TCA) is a known human metabolite of PCE. The
formation of TCA is thought to occur through the formation of an epoxide,
tetrachloroethylene oxide, and its subsequent rearrangement to trichloroacetyl
chloride or trichloroacetaldehyde, 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 PCE.
Tetrachloroethylene oxide, which is considered to be the biologically
active intermediate of the parent compound PCE, was assayed for mutagenicity
in the absence of exogenous metabolic activation in several bacterial tests
(Kline et al. 1982). Tetrachloroethylene oxide increased the number of re-
vertants 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 coll 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 DNA polymerase-deficient
strain in comparison with its polymerase-proficient parent.
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Waskell (1978) tested TCA at 0.45 tug/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 trichloro-
ethanol in 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 chromatid exchange (SCE) formation in primary cultures of human
lymphocytes.
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 PCE 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.
8.7 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 DNA repair assays. In general, the responses were
weak, and eliciting them required rather high toxic concentrations of tetra-
chloroethylene. 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 tetrachloroethylene
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 j_n vitro on a technical
sample of PCE; the preliminary results were negative. (These studies are in
the process of being peer-reviewed, and were not discussed in this chapter,
8-23
-------
except for the Ames test.) The inconsistencies of available results on dif-
ferent samples of PCE may be a function of the toxicity of the test material,
of exposure conditions used for testing this volatile chemical, or of differ-
ences in sample contaminants and/or added stabilizers. Information on chemical
composition of the PCE test samples was scarce.
Although PCE 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 PCE to cause chromosome nondis-
junction, which would result in aneuploidy, a significant genotoxic effect.
Because the epoxide of PCE was mutagenic in bacterial studies, the concern
should be raised that it may pose a mutagenic hazard. It should be 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 warrant a provisional
classification of PCE either as nonmutagenic or mutagenic. If PCE 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 PCE may contain mutagenic impurities and/or added
stabilizers. Although there may be mutagenic agents in certain preparations
of PCE, usually large amounts of material (at toxic levels) were required to
elicit weak responses.
8-24
-------
8.8 REFERENCES
Bartsch, H. , C. Malaveille, A. Barbin, and G. Planche. 1979. Mutagenic and
alkylating metabolites of halo-ethylenes, chlorobutadienes and dichloro-
butenes 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
i_n vitro and i_n 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 screening system test. Mutat. Res. 46:36 (Abst.).
Fabre, F. 1978. Induced intragenic recombination in yeast can occur during
the G-L mitotic phase. Nature 272:795-798.
Fabre, F., and H. Roman. 1977. Genetic evidence for inducibility of recombina-
tion competence in yeast. Proc. Natl. Acad. Sci. USA 74:1667-1671.
Greim, H. , G. Bonse, Z. Radwan, D. Reichert, and D. Henschler. 1975, Mutagen-
icity in vitro and potential carcinogenicity of chlorinated ethylenes as
a function of metabolic oxirane formation. Biochem. Pharmacol. 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. lexicological European Research
3:63-67.
Henschler, D. 1977. Metabolism and mutagenicity of halogenated olefins: a
comparison of structure and activity. Environ. Health Perspec. 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.
8-25
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Kline, S.A., E.G. McCoy, H.S. Rosenkranz, and B.L. Van Duuren. 1982. Mutageni
city of chloralkene epoxides in bacterial systems. Mutat. Res. 101:
115-125.
Margard, W. 1978. In vitro bioassay of chlorinated hydrocarbon solvents.
Battene 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 carci-
nogens as mutagens in the Salmonella/microsome test: Assay of 300 chem-
icals. Proc. Natl. Acad. Sci. USA 72:5125-5139.
NTP. 1983. Unpublished data on tetrachloroethylene: on Salmonella/microsome
preincubation test provided by Dr. 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.
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 Milman, project officer. Unpublished.
8-26
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Williams, G.M., and T. Shimada. January 1983. Evaluation of several halo-
genated ethane and ethylene compounds for genotoxicity. Final report for
PPG Industries, Inc., Pittsburgh, Pennsylvania. Unpublished.
8-27
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9. CARCINOGENICITY
The purpose of this section is to provide an evaluation of the likelihood
that tetrachloroethylene (perchloroethylene, PCE) is a human carcinogen and,
on the assumption that it is a human carcinogen, to provide a basis for esti-
mating its public health impact, including a potency evaluation, in relation
to other carcinogens. The evaluation of carcinogenicity depends heavily on
animal bioassays and epidemiologic evidence. However, information on mutage-
nicity 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 epide-
miologic evidence, the quantitative aspects of assessment, and finally, a
summary and conclusions dealing with all of the relevant aspects of the carci-
nogenicity of PCE.
9.1 ANIMAL STUDIES
Two long-term animal bioassays have been performed to assess the carcino-
genic potential of PCE. In one study involving exposure of rats and mice to
PCE by gavage, the National Cancer Institute (NCI) (1977a) reported the induc-
tion 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 PCE by inhalation,
the Dow Chemical Company (Rampy et al., 1978) reported no evidence for the
carcinogenicity of the chemical. However, limitations in this study make it
difficult to assess the carcinogenic potential of PCE.
9-1
-------
9.1.1 National Cancer Institute Bioassay (1977a)
The PCE sample used in this bioassay was purchased from the Aldrich Chem-
i-cal 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).
The carcinogenicity of PCE 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 ani-
mals 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. PCE 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 with-
out treatment followed by 4 weeks with treatment. Body weights and food con-
sumption 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, embedded in paraplast, and sectioned for slides.
9-2
-------
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.
PCE was found to be carcinogenic in mice in this study. Results summa-
rized in Table 9-1 indicate that PCE induced highly statistically significant
(p < 0.001) increases in the incidence of hepatocel1ular 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 we!1-differentiated hepatocytes arranged in rather
uniform hepatic cordSj 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 hepatocel1ular carcinomas were detected
at 27 weeks in the low-dose group, 40 weeks in the high-dose 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 un-
treated 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 fre-
quently observed finding. A low incidence of bloating or abdominal distension
was noted in 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
9-3
-------
TABLE 9-1. INCIDENCE OF HEPATOCELLULAR CARCINOMAS
IN B6C3F1 MICE FED PCE
Dose (mg/kg/day)a
Hepatocellular carcinomas P values
Males
untreated 2/17 (12%)
vehicle-control 2/20 (10%)
536 32/49 (65%) P < 0-001
1072 27/48 (56%) P < 0.001
Females
untreated 2/20 (10%)
vehicle-control 0/20 (0%)
386 19/48 (40%) P < 0.001
772 19/48 (40%) P < 0.001
aTime-weighted average doses.
bProbability level (p-values) for the Fisher Exact Test comparison of dose
groups with vehicle-control group.
SOURCE: National Cancer Institute, 1977a.
9-4
-------
50 •
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ou —
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3 15 30 45 60 75 90 105 12
— bU
— 40
~
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- 20
- 10
0
TIME ON TEST (WEEKS)
Figure 9-1. Growth curves for male and female mice in the PCE chronic study
SOURCE: National Cancer Institute, 1977a.
9-5
-------
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 carcinogenicity of PCE in rats.
In an attempt to characterize impurities present in the PCE product used
in this study, documented chemical analyses of the test samples performed at
the carcinogenicity testing laboratory of the NCI bioassay program were exam-
ined by the analytical department of the Diamond Shamrock Corporation (inter-
office memorandum from E. A. Rowe to G. K. Hatfield, Diamond Shamrock Cor-
poration, October 8, 1979, obtained with the documented chemical analyses
from G. K. Hatfield, January 28, 1981). As indicated in the interoffice
memorandum, PCE 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
9-6
-------
: SURVIVAL
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Figure 9-2. Survival comparisons of male and female mice in the PCE
chronic study.
SOURCE: National Cancer Institute, 1977a.
9-7
-------
750
in
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30
45
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-600
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105
120
TIME ON TEST (WEEKS)
Figure 9-3. Growth curves for male and female rats in the PCE chronic study.
SOURCE: National Cancer Institute, 1977a.
9-8
-------
~
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— -0.6
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- 0.2
0 15 30 45 60 75 90 105 120
TIME ON TEST (WEEKS)
Figure 9-4. Survival comparisons of male and female rats in the PCE
chronic study.
SOURCE: National Cancer Institute, 1977a.
9-9
-------
carcinogenicity testing laboratory, was reproduced. Evaluation of the ana-
lytical method led to the conclusions presented in the previously mentioned
memorandum that the method could not distinguish epichlorohydrin from tri-
chloroethylene, and that the PCE product used in the NCI (1977a) bioassay
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 PCE samples analyzed at the beginning of the bioassay, at 1 year
into the bioassay, and at 2 years into the bioassay, respectively. The conclu-
sion stated in the memorandum is that the contaminant was probably epichloro-
hydrin, since: 1) epichlorohydrin was a commonly used stabilizer for PCE at
that time; 2) a reported analysis by a competitor of Diamond Shamrock showed a
maximum of 0.015% trichloroethylene in any PCE product manufactured in the
United States, and 3) the decreased amount of impurity found in the PCE samples
as the bioassay progressed suggests some decomposition in that, according to
the experience of Diamond Shamrock, epichlorohydrin but not trichloroethylene
levels would decrease during storage. Nonetheless, although the evidence
provided by Diamond Shamrock indicates that epichlorohydrin was present in the
PCE product used in the NCI (1977a) bioassay, the quantity of epichlorohydrin
in the test material used in this study remains uncertain. Furthermore, the
different levels of unknown material shown in 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 epichlorohydrin impurity in the PCE
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 epichlorohydrin via inhalation. They found that
1/100 rats exposed to 30 ppm (0.115 mg/L air, equivalent to 13.2 mg/kg/day) of
9-10
-------
epichlorohydrin for 6 hours/day, 5 days/week, for 730 days developed squamous
cell nasal carcinomas, as compared wit.h 0/50 in control animals.
The epichlorohydrin impurity in the NCI PCE high-dose male mice experiment
was approximately 1,032 mg/kg/day x 0.041% = 0.42 mg/kg/day, from the discus-
sion 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. (1980) rat experiment. Therefore,
it is unlikely that epichlorohydrin impurities at the level estimated to be
present in the NCI experiment could have contributed appreciably to the positive
response.
In a second carcinogenicity bioassay sponsored by the National Toxicology
Program (NTP, 1983, draft), PCE was given orally to 86C3F1 mice and to four
strains of rats (Sherman, Fischer 344, Long-Evans, and Wistar). The PCE sam-
ple 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 PCE 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 toxi-
city 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 con-
duct 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
9-11
-------
study. The study remains under audit as of June 1985.
9.1.2 Dow Chemical Company Inhalation Study (Rampy et al., 1978)
Two groups of weanling Sprague-Dawley (Spartan substrain) rats, each com-
posed of 96 males and 96 females, were exposed to 600 ppm (4068 mg/m3 air) or
300 ppm (2034 mg/m3 air) of PCE 6 hours/day, 5 days/week, for 52 weeks. An un-
treated 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 sur-
vive 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; PCE, balance. Expo-
sure was done in 3.7 m3 inhalation chambers with a dynamic airflow system.
Analyses of PCE 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 (1,245 analyses). Analytical concentrations
by infrared analysis within +_ 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 evalua-
tion.
Clinical signs of toxicity were not observed with the nominal concentra-
tions of PCE used in this study. Mean body weight gains were similar among
groups. Hematology and urine analyses showed no treatment-related effects
9-12
-------
of PCE. The mortality patterns exhibited in the study are described in
Table 9-2. Mortality in high-dose males was slightly greater than in con-
trols 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.
TABLE 9-2. CUMULATIVE SURVIVAL OF SPRAGUE-DAWLEY RATS
EXPOSED TO PCE FOR 12 MONTHS
Month
of study
Initial
6
12
18
24
31
0
Males
189
187
183
155
44
1
ppm
Females
189
188
185
151
70
12
300
Males
94
94
91
74
26
1
ppm
Females
91
91
91
77
37
6
600
Males
94
88a
84 a
55a
13a
1
ppm
Females
94
94
91
86a
493
5
ap < 0.005 by the Fisher Exact Test.
SOURCE: Adapted from Rampy et al., 1978.
No carcinogenic effects of PCE were observed from pathologic examination
of the animals. Statistical analysis of the data showed numerous nonneoplastic
abnormalities that occurred spontaneously and were within the normal variation
encountered in lifetime studies with this strain of rat. With respect to tumor
findings, analysis of the data did not reveal a definite increased tumor inci-
dence in animals exposed to PCE. 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
9-13
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gross necropsy; however, microscopic 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 sta-
tistically significant (p > 0.05) increase in tumor incidence.
The results of this study do not indicate a definite carcinogenic effect
of PCE in Sprague-Dawley rats; i.e., the tumor incidence between control and
treated rats was similar. However, this study has the following drawbacks:
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.
9.1.3 Intraperitoneal Administration Study (Theiss et al., 1977)
Theiss et al. (1977) tested PCE for carcinogenicity 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 tolera-
ted dosage, defined as the dosage which five mice tolerated after six intra-
peritoneal 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 PCE weekly until total dosages of 1,120, 4,800, and 9,600 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. PCE did not statistically increase (p > 0.05) the incidence of
pulmonary tumors in this study (Table 9-3). This strain was sensitive to the
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positive control chemical urethane, as shown in Table 9-3.
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.
TABLE 9-3. PULMONARY TUMOR RESPONSE TO PCE
Compound
Tri capryl i n
PCE
Urethan
Dosage (mg/kg)
—
80
200
400
1,000 mg/kg
(1 injection)
No. survivors/
No. animals
46/50
15/20
17/20
18/20
20/20
No. lung
tumors/mouse
0.39 +; 0.06a
0.27 + 0.07
0.41 + 0.10
0.50 _+ 0.12
19.6 _+ 2.4
aMean +_ S.E.
SOURCE: Adapted from Theiss et al., 1977.
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, 1977b; 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.
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9.1.4 Skin Painting Study (Van Duuren et al., 1979)
A carcinogenicity study of purified PCE 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 PCE followed
14 days later by the applications of 5.0 yg 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 PCE 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.
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
PCE 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 histologic-
ally.
PCE did not show initiating activity in the initiation-promotion experi-
ment; the number of mice with skin papillomas (squamous cell carcinomas) was:
4 (0) initiated with PCE, 15 (3) treated with PMA alone, and 0 (0) in the
control groups. The study involving repeated application to the skin produced
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lung and stomach tumors in 16 and 0 PCE-treated mice, respectively; 11 and 2
vehicle-treated controls, respectively; and 30 and 5 untreated controls,
respectively.
The negative results for PCE in mouse skin as an initiator and as a com-
plete 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 PCE 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-1,3-dichloropropene, and 2-chloroproponal.
They found that none of the three compounds induced a response as initiators 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 Ha:ICR 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-
dichloropropene (p < 0.0005) by subcutaneous injection.
9.2 EPIDEMIOLOGIC STUDIES
There are seven epidemiologic studies either completed or currently in
progress that relate to PCE exposure. Only two of these studies, however, have
actually identified persons exposed to PCE. Because PCE has been used in the
dry-cleaning industry, however, the present discussion also includes three pro-
portionate 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-
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cleaning industry.
9.2.1 Kaplan (1980)
Kaplan (1980) did a retrospective cohort mortality study of dry-cleaning
workers exposed to PCE 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 PCE exposure were "much higher" for cleaners (machine operators) than for
other employees of dry-cleaning establishments. A geometric mean time-weighted
average for PCE 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 provid-
ed 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 dry-cleaners exposed for more than 1 year prior to
1960. The primary solvent used was PCE. 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 deaths for
whites were derived by multiplying the person-years accumulated for the entire
9-18
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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 neo-
plasms of the colon in whites as an example, this correction was made in the
following manner:
(11/247 x 38) + 11 x 11 x 100 = 182
11 6.98
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
certi fi cates;
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 tests of significance of any of the SMRs were done by the author. However, the
9-19
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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 significant (p < 0.05).
The author points out that because the expected numbers of deaths were
calculated using U.S. rates, which include a higher socioeconomic class than
the 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 PCE found in this study is probably underestimated. Furthermore, the expec-
ted 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 neo-
plasms 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 PCE.
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.
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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 PCE 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-clean-
ing establishments in the United States used petroleum distillates as the
primary cleaning agent prior to 1960, it is quite possible that most of the
shops in this study used petroleum distillates as the cleaning solvent prior to
changing to PCE.
Other important confounding variables were also not controlled. For
example, smoking is a major confounding variable to be considered when evalu-
ating 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 confounding variable for colon cancer.
Another weakness of this study is that 16% 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 PCE for at least one
year prior to 1960, so that the size of the cohort has also been increased.
The results should be available by 1986.
In summary, this study appears suggestive that dry-cleaning workers ex-
posed to PCE are at an elevated risk of colon cancer mortality. Potential
9-21
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exposure to petroleum distillates for approximately half of the cohort, how-
ever, limits any conclusions with regard to the carcinogenicity of PCE 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 establish-
ments; 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 that could
account for the excess proportion of cancers observed.
Only limited conclusions can be drawn from this study as to the potential
carcinogenicity of PCE. 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
9-22
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confounding variables were not considered in the analysis of the results.
These variables include smoking (with.regard to the lung cancer excess),
socioeconomic 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 PCE 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 PCE 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 PCE. 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 de-
ceased 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 employ-
ment in the dry-cleaning industry with excess risks of certain types of cancer,
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cannot be said to demonstrate that PCE exposure presents a carcinogenic risk
because of the possible confounding influence of other dry-cleaning agents, and
because of the fact that no distinction was made between laundry and dry-clean-
i ng workers.
9.2.4 Lin 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 had been selected at random from among contempora-
neous admissions to the same hospital. Cases and controls were asked about
demographic characteristics, residential history, toxic exposures, animal con-
tacts, 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. Further-
more, the dry-cleaners may have used a variety of dry-cleaning agents other
than PCE. 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 PCE exposure.
9.2.5 Duh and Asal (1984)
Duh and Asal (1984) analyzed the death certificates of 440 laundry and
9-24
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dry-cleaning workers who had died in Oklahoma during the period 1975-1981.
Occupations were identified from the death certificates. Laundry workers could
not be separated from dry-cleaning workers because they shared similar occupa-
tional codes. Age, sex, race, and cause distribution of deaths in the United
States in 1978 were used as the standards of comparison. An analysis was done
using the Standardized Mortality Odds Ratio (SMOR) (Miettinen and Wang, 1981).
The SMOR differs from the more traditional Proportionate Mortality Ratio (PMR)
in that it compares the number of deaths from the cause of interest to the
number of deaths from auxiliary causes in the exposed population (the odds) to
the expected odds derived from a comparison population, taking into account
covariates such as age, sex, and race. The authors used "other diseases" and
"all circulatory diseases" as auxiliary causes for the analyses, and indicated
that the results were almost the same as that of the PMR method when "all other
diseases" were used as auxiliary causes, but they differed on two points when
"all circulatory diseases" were used as auxiliary causes. (The authors did not
indicate what these two points were.) The SMOR for homicide was the only SMOR
in the study group that was significantly elevated for non-cancer causes of
death (SMOR = 3.8, 95% confidence interval (CI) = 1.4-10.6; observed deaths =
7, expected deaths = 2.9). Deaths from ischemic heart disease were signifi-
cantly decreased (SMOR = 0.8, 95% CI = 0.7-1.0; observed = 134, expected =
153.7). Among deaths from cancer, the SMORs for cancer of the respiratory
system (SMOR = 1.8, 95% CI = 1.3-2.5; observed - 39, expected = 23.8); cancer
of the lung (SMOR = 1.7, CI = 1.2-2.5; observed = 37, expected = 22.6); and
cancer of the kidney (SMOR = 3.8, 95% CI = 1.9-7.6; observed = 7, expected =
1.9) were significantly elevated. Deaths from cancer of the breast were
significantly decreased (SMOR = 0.1, 95% CI = 0.0-0.4; observed = 1, expected
= 10.5). Since, as the authors noted, wage levels in the laundry and dry-
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cleaning industry are not high, the excess of homicide and the deficit of
breast cancer and ischemic heart disease may have resulted from the inverse
correlation with socioeconomic status.
Smoking histories were not available for any of the death certificates
analyzed. Because both lung and kidney cancer have been associated with
smoking, it is difficult to evaluate the study's elevated SMORs for those
diseases with respect to dry-cleaning exposures. The authors did note, how-
ever, that mortality from other smoking-related diseases such as emphysema,
ischemic heart disease, cancer of the buccal cavity, and pharyngeal cancer was
slightly less than expected.
As the authors noted, there may have been a bias in the analysis because
the occupational code that identified the study group included both laundry and
dry-cleaning workers. Such a bias would tend to weaken any association that
might have been present by adding to the group at potential risk a number of
individuals that would not otherwise have been considered exposed.
Of interest is the authors' report that petroleum solvents accounted for
greater than 50% of the dry-cleaning solvents used in Oklahoma in 1983.
As with several of the other epidemiologic studies reviewed here, this study
suggests that dry-cleaners and/or laundry workers may be at an excess risk of
cancer. However, the possible confounding effects of both petroleum solvents
and smoking, and the fact that dry-cleaners could not be separated from laundry
workers in the analysis, preclude the conclusion that PCE exposure is associ-
ated with an excess risk of cancer.
9.2.6 Asal (personal communication, 1985)
Dr. Nabih Asal of the University of Oklahoma School of Public Health is
currently analyzing the data from a kidney cancer case-control study using both
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hospital and population controls. A variety of potential risk factors are
being evaluated including occupation. While information will not be available
from this study on cases and controls who worked in the dry-cleaning industry
with regard to exposure to specific dry-cleaning solvents, Dr. Asal is also
conducting a cohort study of dry-cleaning establishment owner-operators. Data
on solvent use is extensive for the different dry-cleaning establishments, and
thus it is hoped that this study will provide useful data with which to evalu-
ate PCE with regard to carcinogenicity.
9.3 RISK ESTIMATES FROM ANIMAL DATA
The evidence for the carcinogenicity of PCE serving as the basis for the
quantitation of risks consists of a single positive mouse study (NCI, 1977a).
In this study, there was a statistically significant increase in the incidence
of malignant tumors of the liver in both sexes of B6C3F1 mice.
Because of a question regarding the possible presence of contaminant
epoxides in PCE (< 0.05% the maximum possible), the assay was repeated (NTP,
1983, draft) with PCE without detectable amounts of epoxides, although at lower
dose levels. Doses of 50 mg/kg to 200 mg/kg produced time- and dose-related
increased incidences of hepatocarcinomas. At the present time, the NTP final
report is not available.
The cancer data from epidemiologic studies are inconclusive. The evidence
for mutagenicity of PCE is also inconclusive. However, the more conservative
public health view would regard PCE as a possible human carcinogen. The calcu-
lations in this document for estimating, in quantitative terms, the impact of
PCE as a carcinogen are made independently of the overall weight of evidence
for the carcinogenicity of PCE. The calculations are made as if PCE were a
human carcinogen.
9-27
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9.3.1 Selection of Animal 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 these data sets to use in
the mathematical extrapolation model. It may also be appropriate to correct
for metabolic differences between species and for different routes of admini-
stration.
For PCE, only the data from the NCI (1977a) oral gavage lifetime study
in male and female mice, showing a significant excess of tumors (hepatocellu-
lar carcinomas) have been used. Table 9-4 gives the time-weighted average
dosage and tumor incidence (at "low" and "high" doses) in male and female
B6C3F1 mice for the NCI study. PCE in corn oil was given, once a day, by
gastric intubation, 5 days/week, starting at 25 days of mouse age and con-
tinuing for 78 weeks. During the final 26 weeks, doses were given in a cyclic
pattern of one week without treatment followed by four weeks with treatment.
After the 78-week treatment period, the mice were given an additional 12 weeks
before sacrifice.
Taking into account differing dosage levels, male and female mice appear
to be equally susceptible to a carcinogenic response to PCE. Furthermore,
since the "low" dose and "high" dose response did not differ statistically,
a plateauing of the response is indicated. Hepatic tumors (0-10%) were not
detected in untreated or vehicle-treated groups until 90-91 weeks. For treated
male mice, tumors were detected at 27 weeks in the "low" dose group and 40
weeks in the "high" dose group, and for treated female mice at 41 weeks in the
"low" dose group and 50 weeks in the "high" dose group.
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TABLE 9-4. INCIDENCE OF HEPATOCELLULAR CARCINOMAS IN MALE AND FEMALE B6C3F1 MICE FED PCE BY GAVAGE
I
ro
Sex and no.
mice
Males
20
20
50
50
Femal es
20
20
50
50
aSum (dose in
Median Average
survival terminal
(weeks) weight (g)
>90
>90
30
78
43
>90
>90
25
62
50
mg/kg x no. of days at that dose)
Time-weighted average
gavage dose9
mg/kg/day
untreated
0; vehicleb
536
1,072
untreated
0; vehicle
386
772
mg/animal /day
0
0
16.1
32.2
0
0
9.7
19.3
Incidence
2/17 (12%)
2/20 (10%)
32/49 (65%)
27/48 (56%)
2/20 (10%)
0/20 ( 0%)
19/48 (40%)
19/48 (40%)
Sum (no. of days receiving any dose)
bCorn oil.
SOURCE: NCI, 1977a.
-------
9.3.2 Interspecies Dose Conversion
9.3.2.1 General Consi derations--Sea ling of toxicologic effects, including car-
cinogenicity, among species has been elaborated in several published papers
(Freireich et a!., 1966; Gillette, 1976; Krasovski, 1976; Dedrick, 1973; Rail,
1977; Dedrick and Bischoff, 1980; and Gehring et al., 1980). The major compo-
nents requiring consideration in determining an appropriate extrapolation base
for scaling carcinogenicity data from laboratory animals to man are: 1) toxico-
logic data, 2) metabolism and kinetics, and 3) covalent binding. The published
literature concerning the biological basis for extrapolation of the dose-carci-
nogenic response relationship of laboratory animals to man has been reviewed
in a paper prepared by Davidson (1984) for the Carcinogen Assessment Group
and presented by Parker and Davidson (1984).
9.3.2.1.1 Toxicologic Data. Across species (mice, rabbit, cat, dog, and man)
the acute 1059 of PCE is similar (Tables 6-1 and 6-2; Chapter 6). This simi-
larity is most probably due to an equally effective concentration, leading to
central nervous system depression and resulting in respiratory depression and
death. While other toxicologic end points (modalities of liver and kidney
damage) have been measured (Klassen and Plaa, 1967), these measures are dif-
ficult to evaluate in the numerous and separate studies, and comprehensive com-
parative studies have not been conducted across species using the same toxico-
logic end point.
9.3.2.1.2 Metabolism and Kinetics. It is generally accepted that the liver
and kidney toxicity. and carcinogenicity potential of the halogenated ethy-
lenes relate to the metabolic conversion of these compounds to biologically
reactive intermediates (reviewed by Bolt et al., 1982). All of the chlorinated
ethylenes are initially biotransformed by microsomal monooxygenases to epoxide
intermediates which bind to cellular macromolecules, are molecularly rearranged
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to less reactive species, or are modified by conjugation through the "detoxi-
fication mechanisms." Bolt et al. (1982) postulated that the reactivities and
hence toxicities of the individual epoxides formed from chlorinated ethenes
depend on the type of chlorine substitution. Symmetric substitution such as
is seen with PCE renders the epoxide relatively more stable and therefore less
reactive, while asymmetric substitution results in highly unstable reactive
epoxides (e.g., vinyl chloride). However, tumors may arise in animals by
genetic or epigenetic mechanisms in which recurrent cytotoxicity may lead
to the development of cancer. For example, the accumulative effect of even
minimally reactive metabolites may produce cancer. Therefore, the extent of
metabolic activation- of these metabolites, across species, provides a compara-
tive index of cellular toxicity and a basis for determining a "mouse to man"
extrapolation scaling factor. Moreover, such a comparison provides one index
in the evaluation of the "susceptibility" of a species (e.g., mouse versus
rat).
The metabolism and kinetics of PCE have been studied in three species:
mouse, rat, and man. The principal end metabolites of PCE that have been
reported as common to these species are C02 and the urinary metabolites
trichloroacetic acid (TCA) and oxalic acid. Pegg et al. (1979) specifically
identified the major urinary metabolite in rats as oxalic acid. There is
strong evidence that PCE oxide is formed by mixed-function oxidase (MFO) systems
as a precursor of these metabolites, and the additional reactive intermediate
trichloroacetyl chloride as a precursor of TCA (Bolt et al., 1982). In
general, all of the end metabolites have been poorly identified and quantified
in mouse, rat, and man, and there is no comparative experimental evidence that
the metabolic pathway(s) for PCE qualitatively differs for these species.
9-31
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9.3.2.1.2.1 Oral Studies. Comparative balance studies for the disposition of
PCE (extent of metabolism) have been carried out after both oral and inhalation
exposure in both rats and mice. The oral exposure studies are the most perti-
nent to the conditions of the NCI-PCE bioassay.
Pegg et al. (1979) and Schumann et al. (1980) administered 14C-PCE in corn
oil vehicle as single intragastric doses of 1 and 500 mg/kg to Sprague-Dawley
rats and B6C3F1 mice. 14C-radioactivity was determined for exhaled breath,
urine, feces, and carcass for 72 hours following dosing. In addition, pulmonary
excretion of unchanged PCE was measured. The data from these two studies are
collated and re-expressed in milligram equivalent units in Table 9-5, thus
allowing comparison of the metabolism of PCE in the rat and mouse.
As indicated in Table 9-5, the recovery (95-97%) of ^C-radioactivity in
these experiments indicates that virtually complete absorption of PCE from the
gastrointestinal tract occurs in the rat and mouse. In the rat, peak blood
concentration of PCE occurred 1 hour after oral dosing; disappearance followed
apparent first-order kinetics with a half-life of 6 hours (a reflection of
first-order kinetics for pulmonary excretion and not for metabolism). This
relatively long half-life suggests that PCE doses are not eliminated completely
from the body in less than 24 to 30 hours.
For rats, oral doses of 1 and 500 mg/kg were metabolized 29% (0.07 mg)
and 10% (12.52 mg) respectively, while 71% and 90% of the doses were excreted
unchanged through the lungs. Thus, even at low doses (1 mg/kg) the metabo-
lism of PCE is limited and saturable. For mice, the metabolism of PCE can be
compared to rats at the 500 mg/kg dose. Mice metabolized 17% (1.75 mg) of this
dose and rats 10% (12.52 mg). In the rats, the greater portion of the dose
(83%) was eliminated unchanged through the lungs, indicating limited metabolism
in this species also. These observations of very limited metabolism by rats
9-32
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TABLE 9-5. DISPOSITION OF 14C_PCE RADIOACTIVITY FOR 72 HOURS AFTER SINGLE ORAL DOSES TO
SPRAGUE-DAWLEY RATS AND B6C3F1 MICE
UD
I
GJ
CO
Expi red
unchanged
Metabolized
14co2
Urine
Feces
Carcass
Total
Rats
1 mg/kg
(0.25 mg/animal )
Mg-eq
per animal
0.174 (71%)
0.007
0.040
0.015
0.008
0.070 (29%)
0.244
(av. of 3)a
500 mg/kg
(125 mg/animal )
Mg-eq
per animal
110.67 (90%)
0.57
5.72
4.82
1.41
12.52 (10%)
123.19
Mice (av. of 3)
500 mg/kg
(12.25 mg/animal )
Mg-eq
per animal
8.90 (83%)
0.14
1.53
0.13
0.05
1.85 (17%)
10.75
aBased on average experimental animal weight (grams): 250, rat; 24.5, mouse,
SOURCES: Adapted from Pegg et al., 1979 and Schumann et al., 1980.
-------
and mice are in accord with those of Filser and Bolt (1979) who, for the rat,
estimated Vmax at less than 1.2 mg/kg/hour. Using this estimate for a 250-g
rat, 7 mg of an absolute dose of 125 mg (500 mg/kg) would be expected to be
metabolized in 24 hours. This compares favorably with the 12.52 mg experimen-
tal value given in Table 9-5.
The disposition of the oral dose of 500 mg/kg to rats and mice is compa-
rable to the oral doses given to mice in the NCI bioassay of PCE (386 to 1,072
mg/kg; Table 9-4). The ratio of the dose metabolized per animal for the rat
versus mouse from Table 9-5 is as follows:
Amount metabolized
Dose ratio: rat/mouse
500 mg/kg 12.52/1.85 = 6.77
The body weight ratios for the allometric extrapolation bases wO«67 (surface
area) and t-jl-O (mg/kg) are:
W0.67 = (250/24.5)°-67 = 4.74
Wl-0 = (250/24.5) = 10.20
Hence, the comparative metabolism of PCE for the rat and mouse is more con-
sistent with metabolism proportional to surface area (wO-67) than to body
wei ght.
The long half-life of PCE (6-7 hours) in the rat or man (55-65 hours)
suggests that for daily dosing (24-hour interval) metabolism continues through-
out the entire interval between successive exposures. Therefore, 1) the daily
exposure to PCE represents chronic administration rather than repetitive dosing
in which the dose is completely cleared from the body before the next dose, as
is the case for other chloroethylenes with short half-lives (e.g., trichloro-
9-34
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ethylene, vinylidene chloride); and 2) the "repair" mechanisms are constantly
overshadowed by metabolism.
Buben and O'Flaherty (1985) examined the extent of PCE metabolism in male
mice during subchronic administration by gavage in corn oil vehicle for 6 weeks
(5 days/week; conditions similar to the NCI bioassay of PCE). While the strain
of mice was Swiss-Cox and not the B6C3F1 used in the NCI bioassay, the results
in this study are probably relevant and generally applicable. The investiga-
tors found that PCE metabolism in mice was capacity-limited and dose-dependent.
Metabolism was measured by urinary metabolite(s) excreted per day. TCA was
found to be the only urinary metabolite. Oxalic acid was not considered to
be a significant PCE metabolite, since the oxalic acid excreted in treated and
control animals was comparable. However, labeled PCE was not used, and hence
there was no assurance that a portion of the oxalic acid excreted was not de-
rived from PCE, as was found for the rat by Schumann et al. (1980) and for mice
by Yllner (1961). Because urinary TCA was the sole measure of metabolism in
the Buben and O'Flaherty (1985) study, it is judged that this study represents
somewhat of an underestimate of total metabolism, since other metabolites,
including CO;? in exhaled air, tissue-bound metabolites (Schumann et al., 1980;
Pegg et al., 1979), and metabolite(s) in feces (< 5%; oxalic acid, etc.), were
not taken into account. Nonetheless, urinary TCA can be expected to represent
70% to 80% of the amount of PCE metabolized. Daily urinary TCA tended to in-
crease in amount during each week of treatment, with leveling (plateauing)
towards the end of the week, presumably because of the relatively long half-
life of PCE. Urine collections (24-hour) were taken at the plateau or steady-
state metabolism.
Figure 9-5 shows the dose-dependent relationship between administered dose
(mg/kg/day) and excreted urinary metabolite (mg TCA/kg/day). This hyperbolic
9-35
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U>
cn
O>
D)
O
CD
i<
Ld
I I I I 11 I I I I[I 11 I I I I I I[I I I I I I I I I [ I I I I I I I I I [ I I 11 1111 I I I II I I 11 I I I
200
400 600
800 1000 1200
DOSE (mg/kg)
1400 1600
11 I I I I I I I I I I I I I I 11 I I I I I
1800 2000
Figure 9-5. Relationship between the PCE dose and the amount of total urinary metabolite
excreted per day by mice in each group.
SOURCE: Buben and O'Flaherty, 1985.
-------
curve fitted well to the Michaelis-Menten equation with a Vmax of 136 mg
PCE/kg/day and a Km of 660 mg/kg. The curve deviates from linearity above 200
mg/kg doses, indicating that even at "low" doses the metabolism of PCE is
capacity-limited in mice. Schumann et al. (1980) and Pegg et al. (1979, Table
2) found that for a single 500 mg/kg dose to B6C3F1 mice, 75.5 mg PCE/kg was
metabolized. From Figure 9-5, for a comparable dose 60 mg TCA/kg was observed.
On a molar equivalent basis, these two metabolized doses agree surprisingly
well (455 ymoles PCE versus 367 ymoles TCA, where 1 ymole PCE is metabolized
to 1 pinole TCA). This indicates that about 80% of the metabolism is repre-
sented by urinary TCA, i.e., measuring only the TCA excretion of the mice
probably underestimates total metabolism by about 20%. (Milligrams of urinary
TCA is essentially equivalent to milligrams of PCE metabolized, since the
molecular weight of TCA is 163.40 and that of PCE is 165.85.) Finally, an
important observation can be made from the data of Buben and O'Flaherty (Figure
9-5) relevant to the NCI bioassay. The NCI gavage averaging doses were for
male mice, 536 and 1,072 mg PCE/kg/day, and for female mice, 386 and 772 mg
PCE/kg/day. The amount of metabolites contributing to a carcinogenic response
can be estimated as follows from Figure 9-5 (assuming that Swiss-Cox mice and
B6C3F1 mice metabolize PCE similarly):
Uri nary
NCI metabolite % increase
gavage dose from Fig. 9-5 with dose
mg/kg/day mg TCA/kg/day
Males 536 60.95
38
1,072 84.18
Females 386 50.19
46
772 73.32
9-37
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Thus, for a 100% increase in PCE administered doses, metabolism increases only
38% to 46%. This observation may account in part for the failure to observe a
dose-related tumor increase in mice in the NCI study (Table 9-4). The NTP
bioassay used lower doses of 50 to 200 mg/kg, and dose-related increases of
hepatocarcinomas in mice were found. This is consistent with the dose-depen-
dent findings given in Figure 9-55 since metabolism contributing to tumorigenic
response increases proportionally greater for low gavage doses than for higher
doses. In fact, Figure 9-5 shows nearly a direct linear relationship between
amount of dose metabolized and dose administered below 200 mg/kg.
Finally, it is important to note that Buben and O'Flaherty (1985) demon-
strated that indices of PCE hepatotoxicity (increased liver weight, liver
triglyceride accumulation, glucose-6-phosphatase activity, and serum SGPT
activity) were highly correlative with amount of PCE metabolized by mice. The
degree of liver response, as measured by each toxicity parameter when plotted
against total urinary metabolites, was linear in all cases, suggesting that
the hepatotoxicity of PCE in mice is directly related to the metabolism of PCE.
9.3.2.1.2.2 Inhalation Studies. Pegg et al. (1979) and Schumann et al.
(1980) have also determined body burdens and metabolism of Sprague-Dawley rats
and B6C3F1 mice after inhalation exposure to 10 and 600 ppm PCE for 6 hours.
The animals were exposed to ^C-PCE and the radioactivity determined in urine,
feces, expired air, etc., and unchanged l^C-PCE in expired air for 72 hours
postexposure. The data from the Pegg et al. and Schumann et al. studies are
collated and re-expressed for comparison of rat and mouse metabolism in Table
9-6.
For rats exposed to 10 or 600 ppm, only 32% (0.47 mg) and 12% (9.11 mg)
respectively, of the estimated PCE uptake was metabolized. The remaining PCE
9-38
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TABLE 9-6. DISPOSITION OF l^C-PCE RADIOACTIVITY FOR 72 HOURS AFTER INHALATION EXPOSURE FOR 6 HOURS
TO SPRAGUE-DAWLEY RATS AND B6C3F1 MICE
CO
UD
Expired
unchanged
Metabol i zed
14co2
Urine
Feces
Carcass
Total
Rats (av. of 3)a
10 ppm
1.008 (68%)
0.053
0.275
0.076
0.063
0.467 (32%)
1.475
600 ppm
68.39 (88%)
0.54
4.54
2.36
1.67
9.11 (12%)
77.50
Mice (av.
of 3)
10 ppm
0.048
0.032
0.285
0.027
0.012
0.356
0.404
(12%)
(88%)
aAverage experimental animal weights (grams): 250, rat; 24.5, mouse.
SOURCES: Adapted from Pegg et al., 1979 and Schumann et al., 1980.
-------
uptake, 68% and 88%, was excreted via the lungs unchanged. Thus, limited and
saturable metabolism of PCE is evident. For mice, the metabolism of PCE can
be compared to rats only at the low exposure of 10 ppm exposure concentration.
The percentage of uptake metabolized was considerably more extensive (88% of
the dose), and only 12% was excreted unchanged. However, the absolute amount
of dose metabolized is comparable at 10 ppm between mice and rats. These
results suggest that PCE metabolism is limited and saturable in these species
and confirm similar observations after oral exposure.
The data of Table 9-6 from inhalation studies, when compared with that of
Table 9-5 from oral studies, indicated that, for the rat, inhalation of PCE at
600 ppm for 6 hours resulted in 9.1 mg of compound metabolized. This amount is
comparable to that of an oral 125 mg/kg dose (12.5 mg). Thus, inhalation expo-
sure to 600 ppm for 6 hours is approximately equivalent to an oral dose of 125
mg/kg.
For mice, the metabolism of PCE was determined only at 10 ppm exposure for
6 hours. The amount metabolized, 0.36 mg, is comparable to the amount metabo-
lized in rats, 0.47 mg.
9.3.2.1.2.3 Inhalation Exposure in Humans. PCE is readily absorbed
through the lungs into the blood by first-order passive diffusion processes.
Two major processes account for the known elimination of PCE from the body:
1) pulmonary excretion of unchanged PCE and 2) metabolism of PCE to urinary
metabolites. The metabolism of PCE appears to be very limited in humans,
as it is in experimental animals. Only a small percentage of the estimated
amounts absorbed by inhalation are metabolized to TCA and other chlorinated
metabolites found in the urine (see Chapter 4). Saturation of metabolism has
been estimated to occur in humans at relatively low exposure concentrations
(between 100 and 400 ppm PCE in inspired air--see Chapter 4). Estimates of the
9-40
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extent of metabolism in humans have been made from balance studies by measuring
urinary metabolites after accounting for a retained inhalation exposure dose
(see Chapter 4). However, there are several difficulties associated with
balance studies in humans. For example, problems are encountered in obtaining
an accurate measurement of the retained dose of PCE from inhalation exposure.
Balance studies in humans are obviously incomplete. One shortcoming in some
of the older PCE studies is the imprecision of the analytical methodologies;
for example, the Fujiwara reaction for measuring metabolites. Also, urinary
sampling may not have been sufficient considering the long half-life of PCE
metabolites. In addition, individual variations among subjects can be rather
high. Only urinary TCA was used in these studies as a measure of metabolism,
and the possibility exists that significant amounts of metabolites other than
TCA are excreted in the urine or by other routes. It has been proposed that a
much higher percentage of the dose is actually metabolized, and that a so-far
unrecognized pathway may exist for PCE and for other "metabolites which have not
yet been identified. However, measuring the amount of urinary TCA is one
approach that may be used to compare the metabolized dose of PCE among species.
The metabolized dose may be considered to be the effective dose in organ toxi-
city and carcinogenicity.
9.3.2.1.3 Covalent Binding. Reactive metabolites of PCE have been shown to
bind irreversibly to cellular macromolecules in vitro (Costa and Ivanetich,
1980), and in vivo (Pegg et al., 1979; Schumann et al., 1980). Pegg et al.
(1979) exposed rats by inhalation to 10 and 600 ppm l^C-PCE for 5 hours and to
oral doses of 1 and 500 mg/kg. For rats exposed by inhalation, hepatic irre-
versible binding of l^C-PCE occurred in a ratio of bound/total metabolized of
0.72 - 0.87; the turnover of bound radioactivity was slow as measured over 72
9-41
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hours postexposure. There was no significant difference in the ratio between
inhalation and oral exposure, or between low and high doses, showing that
binding occurs proportionally to the amount metabolized (Tables 9-5 and 9-6).
Schumann et al. (1980) compared the irreversible binding in vivo for rats
and mice exposed to 14C-PCE by gavage (500 mg/kg) as in the NCI bioassay.
Their results are shown in Table 9-7. Peak binding to liver protein occurred
at 6 hours for mice and 6 to 12 hours for rats, with binding in mice four- to
eightfold greater than in rats. These results are in accord with PCE metabo-
lism in these species (Tables 9-5 and 9-6). Schumann et al. found no detect-
able binding of high specific activity 14C-PCE to liver DNA when mice were
exposed by inhalation (600 ppm) or gavage (500 mg/kg). However, the limit of
detection was calculated to be 10 to 14.5 alkylations per 106 nucleotides, and
therefore very low levels of DNA binding may have occurred. These investiga-
tors also chronically administered PCE (11 daily doses) by gavage at increas-
ing doses to 1,000 mg/kg. Mice, but not rats, showed a 20% to 25% increase
of liver/body weight ratio. However, the DNA/g of liver protein decreased,
indicating hypertrophy rather than hyperplasia. Nonetheless, DNA synthesis
(l^C-thymidine incorporation) increased in both mice and rats in a dose-related
manner, indicating activation of "repair" mechanisms. These investigators
interpreted their data as demonstrating cytotoxicity (mouse more susceptible
than rat) enhancing spontaneous incidence of liver tumors, rather than to
genotoxicity. However, the present evidence is insufficient to support one
mechanism over another.
9.3.2.2 Calculation of Dose Metabolized from Animal Data
9.3.2.2.1 Method 1. In the NCI study, the time-weighted average gavage dose
in corn oil was 536 and 1,072 mg/kg/day for male mice, and 386 and 772 mg/kg/
day for female mice; or, based on the average terminal weights, 16.1 and 32.2
9-42
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TABLE 9-7. TOTAL HEPATIC MACROMOLECULAR BINDING FOLLOWING ORAL
EXPOSURE OF RATS AND MICE TO 500 MG/KG
Time
(hours)
1
6
12
24
48
72
A
0.9643
2.7168
1.6868
1.3862
1.0812
0.4723
ymole of
Mean
- Mi ce
+_ 0.1154
+ 0.5333
_+ 0.1119
_+ 0.2622
+_ 0.2895
+ 0.1310
PCE-bound/g
+ S.D.
B
0.1699
0.3310
0.4017
0.6120
0.3474
0.3210
hepatic protein
- Rats
_+ 0.0251
+_ 0.0271
+ 0.0860
+_ 0.0269
+ 0.0595
+_ 0.0601
A/B
5.7
8.2
4.2
2.3
3.1
1.5
SOURCE: Schumann et al., 1980.
9-43
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mg/male mouse/day, and 9.7 and 19.3 mg/female mouse/day (Table 9-4). For a
comparable dose of 500 mg/kg (12.25 mg/animal) in corn oil, virtually complete
absorption was observed (Schumann et al., 1980; Table 9-5). Therefore, it can
be assumed that the PCE doses given by gavage in the NCI study were completely
absorbed. However, a dose-dependent portion of the oral dose is excreted via
the lungs unchanged (83% for a 500 mg/kg dose) and only a fraction is metabo-
lized (17%). Since the percent of the dose exhaled unchanged increases with
increasing dose (Table 9-5), and percent metabolized decreases, these values
represent a conservative estimate (overestimate) for any higher dose of the NCI
bioassay. Hence, the daily body burdens (amount of dose metabolized contribu-
ting to carcinogenic response) for the assay mice are given in Table 9-8.
9.3.2.2.2 Method 2. The estimations of PCE body burden metabolized and con-
tributing to a tumorigenic response may be further refined using the data of
Buben and O'Flaherty (Figure 9-5) describing the relationship between gavage
dose and amount metabolized. Two reservations should be noted: 1) an assump-
tion is made that the relationship determined with Swiss-Cox mice is also
appropriate for B6C3F1 mice of the NCI bioassay, and 2) the relationship under-
estimates metabolism (by about 20%) since only urinary TCA excretion was used
as an index of total metabolism (see discussion in Section 9.3.2.1.2.1).
However, at the doses of the NCI bioassay (male mice, 536 and 1,072 mg/kg and
female mice, 386 and 772 mg/kg) urinary excretion is by far the major route of
metabolized compound, and therefore total urinary metabolites should be a
reasonable approximation of the amount of metabolism. Furthermore, the exper-
imental conditions of chronic daily administration of PCE by gavage closely
mimic the conditions of the NCI bioassay. The daily body metabolic load (amount
of dose metabolized contributed to carcinogenic response, as calculated from
the relationships of Figure 9-5) for the NCI bioassay are given in Table 9-9.
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TABLE 9-8. LIFETIME AVERAGE DAILY EXPOSURE (LAE) FOR B6C3F1 MICE IN THE NCI
BIOASSAY, ASSUMING 17% OF DOSE IS METABOLIZED
Gavage
dose
(mg/kg/day)
Males 536
1072
Females 386
772
Dose
metabol i zed
(mg/kg/day)
91.12
182.24
65.62
131.24
LAEa
(mg/kg/day)
56.41
112.82
40.62
81.24
aThe LAE of the mice is given by:
78 weeks x 5 days x dose metabolized (mg/kg/day)
90 weeks 7 days
SOURCE: NCI, 1977a.
TABLE 9-9. LIFETIME AVERAGE DAILY EXPOSURE (LAE) FOR B6C3F1 MICE IN THE NCI
BIOASSAY, USING A DOSE-METABOLISM (IN URINARY EXCRETION) RELATIONSHIP
Gavage
dose
(mg/kg/day)
Male mice 536
1072
Female mice 386
772
Dose
metabol ized
(mg/kg/day)
60.95
84.18
50.19
73.32
LAEa
(mg/kg/day)
37.73
52.11
31.07
45.39
aThe LAE of the mice is given by:
78 weeks x 5 days x dose metabolized (mg/kg/day)
90 weeks 7 days
SOURCE: Buben and O'Flaherty, 1985.
9-45
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9.3.3 Choice of Risk Model
9.3.3.1 General Considerations—The unit risk estimate for PCE 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 studies.
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 DNA. 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 quanta! response that is characteristic of
mutagenesis is associated with a linear (at low doses) nonthreshold 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 nonthreshold 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 evalu-
ation 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
9-46
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in rat liver and mouse skin).
Because its scientific basis, although limited, is the best of any of the
current mathematical extrapolation models, the nonthreshold model, which is
linear at low doses, has been adopted as the primary basis for risk extrapola-
tion to low levels of the dose-response relationship. The risk estimates made
with such a model should be regarded as conservative, representing a 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 dif-
ferences 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 equi-
valent 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 po-
tency 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 recog-
9-47
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nized. 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.
9.3.3.2 Mathematical Description of the 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 mono-
tonically 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 + qjd + qxd2 + .... + qkdk)J
where
qi >_ 0, i = 0, 1, 2, ..., k
Equi valently,
Pt(d) = 1 - exp [-(qLd + q2d2 + ... + qkdk)J
9-48
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where
pt(d) - fW - R(o)
is the extra risk over background rate at dose d or the effect of treatment.
The point estimate of the coefficients qi , i = 0, 1, 2, . . . , k, and con-
sequently, 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,
Pt(d), are calculated by using the computer program, GLOBAL83, developed by
Howe (1983). 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 q-,. Whenever q-, > 0, at
low doses the extra risk P-t(d) has approximately the form P-t(d) = q^ x d.
Therefore, q x d is a 95% upper confidence limit on the extra risk and
is a 95% lower confidence limit on the dose, producing an extra risk of R. Let
Lg be the maximum value of the log-likelihood function. The upper-limit, q-^,
is calculated by increasing q-, to a value q-i such that when the log-likelihood
is remaximized subject to this fixed value, q-^, for the linear coefficient,
the resulting maximum value of the log-likelihood Lj satisfies the equation
2 (L0 - LI) = 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 always linear. This is conceptu-
ally consistent with the linear nonthreshold concept discussed earlier. The
9-49
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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 experi-
ment, 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 =
-jP-j (1-Pi)
is calculated where N^ is the number of animals in the ith dose group, X^ is
the number of animals in the i*-n dose group with a tumor response, PJ is the
probability of a response in the i^n dose group estimated by fitting the multi-
stage model to the data, and h is the number of remaining groups. The fit is
determined to be unacceptable whenever X^ 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.
9.3.3.3 Adjustments for. Less Than Lifespan Duration of Experiment—If the
duration of experiment Le is less than the natural lifespan of the test animal
L, the slope qls or more generally the exponent g(d), is increased by multi-
plying a factor (L/l_e)3. We assume that if the average dose d is continued,
the age-specific rate of cancer will continue to increase as a constant function
9-50
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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 demon-
strated 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
-ff Q
q-^ [or g(d)] would have been increased by at least (L/L ) .
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)J.
9.3.4 Unit Risk for a Compound
9.3.4.1 Definition of Unit Risk—This section deals with the unit risk for
PCE in air and water and the potency of PCE relative to other carcinogens that
the CAG has evaluated. The unit risk estimate for an air or water pollutant is
defined as the incremental lifetime cancer risk occurring in a hypothetical
population in which all individuals are exposed continuously from birth
throughout their lifetimes to a concentration of 1 pg/m^ of the agent in the
air they breathe, or to 1 pg/L in the water they drink. This calculation is
done to estimate in quantitative terms the impact of the agent as a carcinogen.
Unit risk estimates are used for two purposes: 1) to compare the carcinogenic
potency of several agents with each other, and 2) to give a crude indication
of the population risk which might be associated with air or water exposure to
these agents, if the actual exposures are known.
9-51
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For most cases of interest to risk assessment, the 95% upper-limit risk
associated with dose units/day as obtained by means of the GLOBAL83 computer
program (Howe, 1983), can be adequately approximated by P(d) = 1 exp [-(q^)].
9.3.4.2 Unit Risk Estimates and Interpretation—The unit risk estimate based
on animal bioassays is only an approximate indication of the absolute risk in
populations exposed to known carcinogen concentrations. There may be impor-
tant differences in target site susceptibility, immunological responses, hor-
mone function, dietary factors, and disease. In addition, 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 compared with other carcinogens. The comparative potency of
different agents is more reliable when the comparison is based on studies in
the same test species, strain, and sex.
The quantitative aspect of carcinogen risk assessment is included here
because it may be of use in the regulatory decision-making process, e.g., set-
ting regulatory priorities, evaluating the adequacy of technology-based controls,
etc. However, it should be recognized that the estimation of cancer risks to
humans at low levels of exposure is uncertain. At best, the linear extrapola-
tion model used here provides a rough but plausible estimate of the upper-limit
of risk; i.e., it is not likely that the true risk would be much more than the
estimated risk, but it could very well be considerably lower. The risk esti-
mates presented in subsequent sections should not be regarded as immutable
representations of the true cancer risks; however, the estimates presented may
be factored into regulatory decisions to the extent that the concept of upper
risk limits is found to be useful. Table 9-10 gives the upper-limit slope
estimates, q-^, from the linearized multistage model for PCE. The slope
9-52
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estimate can be used to compare the relative carcinogenic potency of PCE to
that of other potential human carcinogens. The slope is also used to calculate
upper-bound incremental risks at low levels of exposure.
9.3.5 Alternative Low-Dose Extrapolation Models
In addition to the multistage model, two more models, referred to as the
probit and the Weibull, are often employed by the CAG for low-dose extrapolation
for purposes of comparison. However, calculations made with these two alter-
native models are not presented in this document, since no meaningful compari-
sons can be made on the basis of these calculations. For example, the 95%
upper-bound estimate of cancer risk at the metabolized dose 0.1 mg/kg/day is
predicted to be 1 by either the probit or the Weibull model, while the corre-
sponding risk estimated by the multistage model is only 0.02.
9.4 CALCULATION OF UNIT RISK
9.4.1 Potency (Slope) Estimate on the Basis of Animal Data
The dose metabolized, which is defined as the total uptake minus the amount
expired unchanged, is used as the dosimetry in the risk assessment. The cancer
potency (slope) calculated on this basis can be converted back to the administered
dose by using the dose-metabolism relationship estimated from experimental data.
Table 9-10 presents the carcinogenic potency for humans along with the
metabolized doses calculated by two approaches and tumor incidence rates in the
NCI (1977a) bioassay. The detailed derivations of the metabolized doses are
presented in Tables 9-8 and 9-9. To account for the high mortality rates in
the dosed groups, the animals that died before the occurrence of the first
hepatocellular carcinomas are excluded from the denominators of tumor incidence
rates. Another approach would be to use the time-to-death data for risk calcu-
lation. This approach yields a slightly higher risk value than that calculated
9-53
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TABLE 9-10. DOSE-RESPONSE DATA AND POTENCY (SLOPE) ESTIMATES
Potency estimate0
Data base for
calculating
body burden
Method 1: Use
total metabolized
dose (Schumann
et al., 1980)
Method 2: Use
dose metabolized
in urinary
excretion (Buben
and O'Flaherty,
1985)
(animal )
(mg/kg/day)
Male mice:
0
56.41
112.82
Female mice:
0
40.62
81.24
Male mice:
0
37.73
52.11
Female mice:
0
31.07
45.39
Tumor
incidence13
2/20
32/48
27/45
0/20
19/48
19/45
2/20
32/48
27/45
0/20
19/48
19/45
(mg/kg/day )-•>-
qi(M) q*(A)
1.8x10-1 4.4x10-2
1.6x10-1 4.0x10-2
3.4x10-1 6.8x10-2
2.5x10-1 5.1x10-2
aSee Tables 9-8 and 9-9 for the animal lifetime average exposures (LAE).
bThe denominators are the number of animals that survived at the time the first
hepatocellular carcinoma occurred in each study. The first tumor occurred at
2£ weeks for the male mice and at 41 weeks for the female mice.
cqi(M) is the potency estimate expressed in terms of the metabolized dose.
q*(A) is the potency estimate expressed in terms of the administered dose.
9-54
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on the basis of incidence data with a modified denominator. However, the risk
calculated from the time-to-death data is difficult to interpret because the
majority of animals with tumors were discovered at the terminal sacrifice.
Therefore, only the potency estimates calculated on the basis of incidence data
are presented.
Two sets of potency estimates are presented in the last two columns of
Table 9-10, using either the total dose metabolized (Method 1) or the dose meta-
bolized and eliminated in urine (Method 2). These estimates are calculated by
the linearized multistage model, using the dose-response data in Table 9-10.
The values q^(M) and q^(A) are potency estimates expressed, respectively,
in terms of the metabolized and the administered doses.
The carcinogenic potency q^(M) for humans is calculated by multiplying
the animal potency q-^(M) by a factor (W^/W^)1/ , where WH and W^ are, respec-
tively, the body weights for humans and animals. The underlying assumption for
this adjustment is that the metabolized dose per surface area is equally effec-
tive in inducing cancer among species. To convert from q^(M) to q-^(A), the
relationship between administered dose (A) and the amount metabolized (M) is
utilized. For the estimates calculated according to Method 2, the relationship
M = 0.20A is used. This relationship can be obtained from the dose-metabolism
relationship derived by Buben and O'Flaherty (1985) when the administered dose
A is small. For the estimates calculated by Method 1, the relationship M =
0.25A is used, as in Buben and O'Flaherty (1985), and considering the fact that
the urinary metabolites account for about 80% of the total metabolites. It is
worthwhile to note that these relationships are consistent with the empirical
observations of Pegg et al. (1979), in which approximately 29% of the dose was
metabolized after rats were given a small dose (1 mg/kg or 0.25 mg per animal).
Because the body burdens calculated by Methods 1 and 2 result in compa-
9-55
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rable potency estimates, and because the available human body burden via inha-
lation is given in terms of urinary metabolites, it is recommended that the
potency estimates on the basis of Method 2 be used to represent the carcino-
genic potency of PCE. Since more reliable dose-response data exist for female
mice than for male mice, the carcinogenic potency calculated from these data is
used. The slope, q^, estimate for PCE is:
qj(A) - 5.1 x ID'2 mg/kg/day
expressed in terms of the administered dose, and
q£(M) - 2.5 x 10"1 mg/kg/day
expressed in terms of the metabolized dose in urinary excretion.
9.4.2 Risk Associated with 1 ug/L of PCE in Drinking Water
The daily PCE intake (mg/kg/day) for a 70 kg person consuming 2 L of
water contaminated with 1 Mg/L of PCE is:
d = (1 ug/L) x (2 L/day) x (1Q-3 mg/pg)/70 kg
= 2.9 x 1CT5 mg/kg/day
Therefore, the upper-bound estimate of the incremental lifetime risk due to con
suming water contaminated with 1 pg/L of PCE is:
P = 5.1 x ID'2 x 2.9 x ID"5 = 1.5 x 1CT6
9.4.3 Risk Associated with 1 U9/m3 of PCE in Air
In order to estimate the risk due to 1 yg/m3 of PCE in air, it is neces-
sary to estimate the amount metabolized that accompanies exposure to this con-
centration in air. For PCE, the body burden can be estimated from both human
9-56
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and animal data. Since only the metabolized dose in urine is available for
humans, the potency q^(M) = 2.5 x 10'1 mg/kg/day in terms of urinary metabo-
lites will be used in the risk calculation.
9.4.3.1 Unit Risk Calculated on the Basis of Body Burden Derived from Human
Data — In an experimental human exposure to PCE vapor, Fernandez et al. (1976)
observed that in 72 hours, the quantity of TCA metabolized and eliminated in
urine was about 25 mg following 150 ppm exposure for 8 hours. Assuming that
the amount metabolized in urine is linearly proportional both to the air
concentration and to the duration of exposure, the observations of Fernandez et
al. (1976) imply that a continuous (24 hours) exposure to 1 ppm of PCE produce"
25 x 3/150 = 0.5 mg/day of urinary metabolites. Since 1 ppm = 6,908 yg/m3, a
continuous exposure to 1 yg/m3 of PCE yields a metabolized dose of 7.3 x 10~5
mg/day. For chronic and continuous exposures, the metabolized dose would be
7.3 x 1Q-5/70 = 1.0 x 10~6 mg/kg/day. Therefore, the incremental lifetime risk
due to 1 yg/m3 of PCE in air is:
P = 2.5 x 10-1 x i.o x 10-6 = 2.5 x 10~7
A study by Bolanowska and Golacka (1972) also provides some information
on the relationship between the air concentration and the amount metabolized
in urinary excretion. In this study, five subjects were exposed to 390,000
yg/m3 of PCE (approximately 50 ppm) for 6 hours. The concentration of meta-
bolites in urine (mg/hr) is presented graphically over the 20-hour period
during and after exposure. The total amount of metabolites in urine (area
under curve) is estimated to be about 13 mg. The area under the curve from 20
hours to infinity is calculated by C x ti/2/0.693, where C is the concentration
at the last sampling time (i.e., 20 hours) and ty?, the half-life, is assumed
to be 100 hours. Under the assumption that the amount metabolized is linearly
9-57
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proportional to the air concentration and the duration of exposure, the amount
metabolized, associated with 1 pg/m3 of PCE in air, is
(13 mg/390,000) x (24 hours/6 hours) - 1.33 x 1Q-4 mg/day
Equivalently, the amount metabolized due to continuous exposure to 1 yg/m3 of
PCE in air would be 1.33 x 10'4 mg/day for a 70-kg man, which would be 1.9 x
ID'6 mg/kg/day. Thus, the cancer risk due to 1 yg/m3 of PCE in air is esti-
mated to be
P - 2.5 x 10-1 x 1.9 x ID'6 = 4.8 x 10"7
This risk estimate is about two times higher than that calculated on the
basis of Fernandez et al. (1976). Since the study by Bolanowska and Golacka
(1972) was conducted at lower concentrations, the risk calculation on the
basis of this study is preferred. It is therefore recommended that 4.8 x 10~?
be used as the unit risk estimate for PCE in air.
9.4.3.2 Unit Risk Calculated on the Basis of Body Burden Derived from Animal
Data—Table 9-6 contains some animal data that can be used to infer human
metabolized dose when exposed to 1 yg/m3 of PCE in air. The dispositions of
14C-PCE radioactivity in rats and mice 72 hours after exposure to 10 ppm of PCE
in air for a period of 6 hours are as follows:
Rats Mice
(mg/animal) (mg/animal)
Expired unchanged 1.008 0.048
Total metabolized 0.467 (32%) 0.356 (88%)
Metabolites in urinary
excretion 0.275 (19%) 0.285 (71%)
Total uptake 1.475 0.404
9-58
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These data indicate that the percentage of PCE that is metabolized de-
creases with an increase of body weight and that the total uptake, not the
amount metabolized, is approximately proportional to the body surface area
between the two species. To estimate the amount metabolized in human urinary
excretion after PCE exposure, it is assumed that the total uptake is propor-
tional to the body surface area among species, including humans, and the
relationship between the percentage metabolized (M) and the body weight (W)
follows the equation M = AWK. The parameters A and K are estimated to be
A = 0.146 and K = 0.420, using the animal data presented above. Thus, for a
70-kg person, the percentage metabolized is predicted to be 2.5%. This is
consistent with the empirical observation in humans (Monster, 1979). On this
basis, if humans were exposed to 10 ppm of PCE in air, as were the animals,
the amount metabolized would be
1.48 x (70/0.25)2/3 x 0.025 = 1.58 mg
where 1.48 x (70/0.25)2/3 js the estimated total uptake using the surface area
correction factor (70/0.25)2/3.
Under the assumption that the amount metabolized (in urine) is linearly
proportional to both the air concentration and the duration of exposure, the
amount metabolized due to a continuous (24-hour) exposure to 1 ppm of PCE in
air is 1.58 x 4/10 = 0.63 mg/day. Equivalently. the body metabolic load is 0.63/
6,908 = 9.2 x 10~5 mg for a 24-hour exposure when the air concentration is
1 yg/m3, using the fact that 1 ppm = 6,908 ^g/m3. For chronic and continuous
exposure, the amount metabolized per day due to 1 yg/m3 of PCE in air is 9.2
x ID'5 mg/70 kg/day = 1.3 x 10'6 mg/kg/day. Therefore, the incremental lifetime
risk due to 1 pg/m3 of PCE is:
9-59
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P = 2.5x10-1 x 1.3x10-6 = 3.3x10-7
This estimate is about the same as that calculated on the basis of direct ob-
servations regarding human metabolized dose. Since data based on direct obser-
vation of humans is preferred to the animal data, it is recommended that the
value 4.8 x 10~7, calculated on the basis of actual human metabolized dose, be
used as the unit risk estimate for PCE in air.
9.4.3.3 Unit Risk Calculated on the Basis of a Negative Inhalation Study--
For comparison, a negative inhalation study on rats, sponsored by the Dow
Chemical Company (Rampy et al., 1978) is used to calculate an upper-bound
estimate of risk for PCE. Ninety-four animals of each sex were exposed to
either 300 ppm or 600 ppm of PCE, 6 hours/day, 5 days/week for 52 weeks.
Animals were allowed to survive until the terminal sacrifice at 31 months.
A shortcoming of this study was that the animals were exposed to PCE for only
52 weeks. The use of data from such a study may seriously underestimate the
risk. Among the four PCE-exposed groups, the data from the female high-dose
group yield the smallest upper-bound estimate of risk, and thus will be used
to calculate the unit risk for PCE in air. Forty-nine female rats in the high-
dose (600 ppm) group survived beyond 2 years. The 95% upper-bound estimate of
the probability of cancer among these 49 animals with no cancer response is Pu
= 0.059. The corresponding lifetime average exposure is
d = 600 ppm x 5 days x 6 h°urs x 52 weeks = 53.57 ppm
7 days 24 hours 104 weeks
or
d - 370,071 yg/m3
9-60
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Thus, the carcinogenic potency of PCE, defined as the slope in the one-hit
model, is calculated to be
b = [-ln(l-Pu)]/d = 1.6 x 10-7/yg/m3
for animals.
To calculate the risk to humans associated with 1 yg/m3 of PCE in air,
it is necessary to determine the human exposure concentration C(yg/m3)
equivalent to 1 yg/m3 of PCE in air for rats. Assuming that dose in rela-
tion to surface area is equally effective among species, the human exposure
concentration C(yg/m3) satisfies the equation
C(yg/m3) x (20 m3/day)/(70 kg)2/3
= 1 yg/m3 x (0.224 m3/day)/(0.35 kg)2/3
where 20 m3/day and 0.224 m3/day are the daily air intakes, respectively, for
a 70-kg person and an 0.35-kg rat. Thus, C(yg/m3) = 0.39 yg/m3. Therefore,
the human risk due to 1 yg/m3 of PCE in air is
P = 1.6 x 10-7/0.39 - 4.1 x 10~7
This risk estimate is again comparable to those calculated previously.
9.4.3.4 Discussion
To the extent possible, the available metabolism and pharmacokinetic
information for PCE has been used in the risk calculations. Nevertheless, the
use of pharmacokinetic information may reduce some of the ambiguity associated
with three major areas of risk assessment uncertainty listed below:
1. Extrapolation from high dose-response in animals to low dose-response
9-61
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in animals (i.e., low-dose extrapolation),
2. Extrapolation from low dose-response in animals to low dose-response
in humans (i.e., species conversion), and
3. Extrapolation from gavage to inhalation (i.e., route-to-route
extrapolation).
In calculating the dose-response relationship for PCE, the amount meta-
bolized is considered to be an effective dose. The use of this surrogate
effective dose may not eliminate the uncertainty associated with the low-dose
extrapolation because the dose actually reaching the receptor sites may not be
linearly proportional to the total amount metabolized, and the shape of the
dose-response relationship is still unknown. However, it seems reasonable to
expect that the uncertainty with regard to the low-dose extrapolation would be
somewhat reduced by the use of the metabolized dose because the metabolized
dose better reflects the dose-response relationship, particularly in the
high-dose region.
To extrapolate from animals to humans, the metabolized dose per body sur-
face area is assumed to be equivalent (i.e., equally potent) among species.
This assumption is by no means supported by the empirical data. An alternative
approach for animal-to-human extrapolation would be to assume that mg metabolized
dose/kg/day is equivalent among species. If this assumption is made, the
slope factor ql5 as calculated for PCE and expressed in terms of (mg/kg/day)-1,
would be reduced by approximately 10 times.
The carcinogenic potency calculated from the gavage study and on the basis
of metabolized dose does provide a better basis for extrapolating the risk
estimate from gavage to inhalation. This is so because the relationship between
the ambient air concentration and the amount metabolized for PCE is available
9-62
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for both humans and animals. Therefore, the uncertainty due to route-to-route
extrapolation should be greatly reduced by the use of metabolized dose.
9.5 RISK PREDICTION AGAINST HUMAN EXPERIENCE
Although there is no sound epidemiologic data that could be used to esti-
mate the carcinogenic potential of PCE, efforts are made herein to determine
whether or not the risk estimate of 4.8 x 10~? per yg/m3 is reasonable in
view of the fact that none of the epidemiologic studies on dry-cleaners showed
significant elevation of cancer risk,, The major human experience in PCE expo-
sure has been on the part of dry-cleaning workers. Among them, the machine
operators are the workers most heavily exposed, with a time-weighted average
concentration of about 22 ppm. For other dry-cleaners, the mean exposure
concentration is approximately 3 ppm (Ludwig et al., 1983).
According to Campbell et al . (1980), the number of persons employed as
machine operators in commercial, industrial, and coin-operated dry-cleaning
facilities in the United States is estimated to be 27,700. A question to be
asked is whether a significant increase of cancer deaths should be expected
among dry-cleaning workers, and in particular, among the machine operators.
As demonstrated below, a cohort study on dry-cleaning workers is unlikely to
show an elevation of cancer risk.
At an ambient air concentration of 22 ppm (for machine operators) and 3
ppm (for other dry-cleaners), workers who work 8 hours/day, 250 days/year for
one-third of their lifetime would have an average daily exposure of
d = 22 ppm x 6908 (Mg/m3)/ppm x 8 hours x 25° days x I
24 hours 365 days 3
= 11,565.90 ug/m3
9-63
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for machine operators and, similarly,
d = 1577.17 Mg/m3
for other dry-cleaners. Therefore, the incremental lifetime risk is
P = 4.8 x lO"7 x 11,565.9 = 5.6 x 10'3
for machine operators, and
P = 4.8 x 10-7 x 1577.17 = 7.6 x 1(H
for other dry-cleaners.
The incremental lifetime risks for machine operators (5.6 x 10"3) and
other workers (7.6 x ICT4) are not likely to be detected by a cohort study
because of the large sample sizes required. The sample sizes required to
show a significant elevation of cancer risk at a level of 0.05 (one-sided test)
and a power of 0.80 are as follows, using the formula presented in Schlesselman
(1974).
Background Sample size required
1 ifetime
cancer rates
0.01
0.05
0.10
Machine operators
5,000
25,000
36,263
Other workers
218,000
509,000
1,705,000
The sample size required depends on the background cancer rate. Since
no specific cancers are known to be associated with PCE exposure, three hypo-
thetical background rates are used in the calculations, which should cover most
of the rates for specific cancers in the U.S. population.
9-64
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9.6 COMPARISON OF POTENCY WITH OTHER COMPOUNDS
One of the uses of the concept of unit risk is to compare the relative
potencies of carcinogens. Figure 9-6 is a histogram representing the frequency
distribution of potency indices for 54 suspect carcinogens evaluated by the
CAG. The actual data summarized by the histogram are presented in Table 9-11.
The potency index is derived from q-^, the 95% upper bound of the linear compo-
nent in the multistage model, and is expressed in terms of (mmol/kg/day)~1.
Where human data are available for an agent, they have been used to calculate
the index. Where no human data are available, animal oral studies are 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 PCE has been calculated as 8 x 10^. This figure is
* o
derived by multiplying the slope q-j_ = 5.1 x 10 (mg/kg/day) and the molecular
weight of PCE, 165.8. This places the potency index for PCE in the fourth
quartile of the 54 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 a low-dose linear extrapolation
model. 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 may be more appropriate.
9-65
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4th
QUARTILE
3rd
QUARTILE
2nd
QUARTILE
1st
QUARTILE
10'
2 x 1CT3
-1
0
123456
LOG OF POTENCY INDEX
Figure 9-6. Histogram representing the frequency distribution of
the potency indices of 54 suspect carcinogens evaluated by the
Carcinogen Assessment Group.
9-66
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TABLE 9-11. RELATIVE CARCINOGENIC POTENCIES AMONG 54 CHEMICALS EVALUATED BY THE CARCINOGEN ASSESSMENT GROUP
AS SUSPECT HUMAN CARCINOGENS
UJ>
i
Level
of evidence3
Compounds
Aery! onitrile
Aflatoxin B1
Al drin
Ally"! chloride
Arsenic
B[a]P
Benzene
Benzidene
Beryl! ium
1,3-Butadiene
Cadmi urn
Carbon tetrachloride
Chlordane
CAS Number
107-13-1
1162-65-8
309-00-2
107-05-1
7440-38-2
50-32-8
71-43-2
92-87-5
7440-41-7
106-99-0
7440-43-9
56-23-5
57-74-9
Humans
L
L
I
S
I
S
S
L
I
L
I
I
Animal s
S
S
L
I
S
S
S
S
S
S
S
L
Groupi ng
based on
IARC
criteria
2A
2A
2B
1
2B
1
1
2A
2B
2A
2B
3
SI opeb
(mg/kg/day)-1
0.24(W)
2900
11.4
1.19xlO-2
15(H)
11.5
2.9xlO-2(W)
234(W)
2.6
1.0xlO~1(I)
6.1(W)
1.30X10-1
1.61
Molecul ar
weight
53.1
312.3
369.4
76.5
149.8
252.3
78
184.2
9
54.1
112.4
153.8
409.8
! __.__!• _ — — i -
Potency
index0
lxlO+1
9xlO+5
4x1 0+3
9x10-!
2xlO+3
3xlO+3
2x10°
4xlO+4
2X10+1
5x10°
7xlO+2
2xlO+1
7x10+2
~ TT f _ 1 ^ . ~
Order of
magnitude
(log10
index)
+1
+6
+4
0
+3
+3
0
+5
+1
+1
+3
+1
+3
-------
TABLE 9-11. (continued)
Compounds
Chlorinated ethanes
1,2-Dichloroethane
hexachloroethane
Level
of evidence3
CAS Number Humans
107-06-2 I
67-72-1 I
1,1,2,2-Tetrachloroethane 79-34-5 I
1,1,2-Trichloroethane
Chloroform
Chromium VI
f DDT
01
00
Dichlorobenzidine
1,1-Dichloroethylene
(Vinylidene chloride)
Dichloromethane
(Methylene chloride)
Dieldrin
2,4-Dinitrotoluene
Diphenylhydrazine
Epichlorohydrin
Bis(2-chloroethyl)ether
79-00-5 I
67-66-3 I
7440-47-3 S
50-29-3 I
91-94-1 I
75-35-4 I
75-09-2 I
60-57-1 I
121-14-2 I
122-66-7 I
106-89-8 I
111-44-4 I
Animal s
S
L
L
L
S
S
S
s
L
L
S
S
S
S
S
Grouping
based on
IARC
criteri a
2B
3
3
3
2B
1
2B
2B
3
3
2B
2B
2B
2B
2B
SI opeb
(mg/kg/day)-1
9.1xlO-2
1.42xlO-2
0.20
5.73x10-2
7x10-2
41(W)
0.34
1.69
1.16(1)
6.3xlO-4(I)
30.4
0.31
0.77
9.9xlO-3
1.14
Molecul ar
weight
98.9
236.7
167.9
133.4
119.4
100
354.5
253.1
97
84.9
380.9
182
180
92.5
143
Potency
index0
9x10°
3x10°
3xlO+1
8x10°
8x10°
4xlO+3
1x10+2
4x10+2
1x10+2
5x10-2
1x10+4
6xlO+1
lxlO+2
9x10-!
2x10+2
Order of
magnitude
(Iog10 '
index)
+1
0
+ 1
+ 1
+ 1
+4
+2
+3
+2
-1
+4
+2
+2
0
+2
(continued on the following page)
-------
TABLE 9-11. (continued)
Compounds
Bis(chloromethyl )ether
Ethylene dibromide (EDB)
Ethylene oxide
Heptachlor
Hexachlorobenzene
Hexachlorobutadiene
Hexachlorocyclohexane
technical grade
alpha isomer
beta isomer
gamma isomer
Hexachlorodibenzodioxi n
Nickel
Nitros amines
Dimethyl nitrosamine
Diethylnitrosamine
Dibutylnitrosamine
N-nitrosopyrrolidine
N-nitroso-N-ethylurea
CAS Number
542-88-1
106-93-4
75-21-8
76-44-8
118-74-1
87-68-3
319-84-6
319-85-7
58-89-9
34465-46-8
7440-02-0
62-75-9
55-18-5
924-16-3
930-55-2
759-73-9
of
Level
evidence3
Humans Animals
S
I
L
I
I
I
I
I
I
I
L
I
I
I
I
I
S
S
S
S
S
L
S
L
L
S
S
S
S
S
S
S
Groupi ng
based on
IARC
criteria
1
2B
2A
2B
2B
3
2B
3
2B
2B
2A
28
2B
2B
2B
2B
SI opeb
(mg/kg/day)-l
9300(1)
41
3.5x10-1(1)
3.37
1.67
7.75x10-2
4.75
11.12
1.84
1.33
6.2x10+3
1.15(W)
25.9(not by
43.5(not by
5.43
2.13
32.9
Molecul ar
weight
115
187.9
44.1
373.3
284.4
261
290.9
290.9
290.9
290.9
391
58.7
qf) 74.1
qf) 102.1
158.2
100.2
117.1
Potency
indexc
1x10+6
8x10+3
2xlO+1
1x10+3
5x10+2
2xlO+1
1x10+3
3x10+3
5x10+2
4x10+2
2xlO+6
7xlO+1
2x10+3
4xlO+3
9x10+2
2x10+2
4x10+3
i: u^. £~^i i _. .
Order of
magnitude
(Iog10
index)
+6
+4
+ 1
+3
+3
+1
+3
+ 3
+ 3
+ 3
+6
+2
+3
+4
+ 3
+ 2
+4
-------
TABLE 9-11. (continued)
Compounds CAS Number
N-nitroso-N-methyl urea
N-nitroso-diphenyl amine
PCBs
Phenol s
2,4,6-Trichl orophenol
Tetrachl orodibenzo-
^ p-dioxin (TCDD)
o
Tetrachl oroethyl ene
Toxaphene
Trichloroethylene
Vinyl chloride
684-93-5
86-30-6
1336-36-3
88-06-2
1746-01-6
127-18-4
8001-35-2
79-01-6
75-01-4
Level
of evidencea
Humans
I
I
I
I
I
I
I
I
S
Animal s
S
S
S
S
S
L
S
L/S
S
Grouping
based on
IARC
criteria
2B
2B
2B
2B
2B
3
2B
3/2B
1
SI opeb
(mg/kg/day)"1
302.6
4.92xlO-3
4.34
1.99x10-2
1.56xlO+5
5.1x10-2
1.13
1.1x10-2
1.75x10-2(1)
Molecul ar
weight
103.1
198
324
197.4
322
165.8
414
131.4
62.5
Potency
indexc
3x1 0+4
1x10°
lxlO+3
4x10°
5xlO+7
8x10°
5x10+2
1x10°
1x10°
Order of
magnitude
(Iog10
index)
+4
0
+3
+1
+8
+1
+3
0
0
dS = Sufficient evidence; L = Limited evidence; I = Inadequate evidence.
^Animal slopes are 95% upper-bound slopes based on the linearized multistage 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 nonthreshold model. 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. The slope value is an upper bound in the sense that the true value (which
is unknown) is not likely to exceed the upper bound and may be much lower, with a lower bound approaching zero.
Thus, the use of the slope estimate in risk evaluations requires an appreciation for the implication of the upper
bound concept as well as the "weight of evidence" for the likelihood that the substance is a human carcinogen.
cThe potency index is a rounded-off slope in (mmol/kg/day)'1 and is calculated by multiplying the slopes in
(nig/kg/day)-! by the molecular weight of the compound.
-------
9.7 SUMMARY
9.7.1 Qualitative
Tetrachloroethylene (PCE) 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
PCE used was over 99% pure, but was estimated to contain epichlorohydrin con-
centrations of less than 500 ppm. It is unlikely that this response could be
attributed to the low concentration of epichlorohydrin.
No carcinogenic effects were observed in lifetime studies of Osborne-
Mendel rats given PCE by gavage or in Sprague-Dawley rats exposed via inhala-
tion 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 PCE in rats.
Intraperitoneal injection of PCE in strain A mice induced no statistically
significant incidence of pulmonary adenomas. In mouse skin initiation-promo-
tion experiments, PCE 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, these negative
results 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 PCE, it was found
that workers were at an elevated risk of colon cancer mortality; however,
as many as one half of these workers may have been exposed to petroleum dis-
tillates in their working history. Other epidemiologic studies that found
an association between cancer risk and employment in the dry-cleaning industry,
did not identify which dry-cleaning solvents the employees had been exposed to.
9-71
-------
9.7.2 Quantitative
Data on hepatocellular carcinomas in male and female mice from an NCI
(1977a) gavage study on PCE have been used to calculate the unit risk for
water and air. The unit risk is defined as the increased lifetime cancer risk
due to exposure to one unit of dose in drinking water (1 yg/L) or in air
(1 yg/m3). Evaluation of the metabolism data in mice, rats, and humans shows
that common principal end metabolites have been identified and that there is
no comparative experimental evidence that the metabolic pathways for PCE
qualitatively differ for these species. The use of relevant metabolism and
pharmacokinetic data in the development of unit risk calculations helps to
minimize some of the uncertainty associated with route to route extrapolation.
The upper-bound estimate of the incremental cancer risk due to 1 yg/L
of PCE in drinking water is 1.5 x 10~6. The upper-bound estimate of the
incremental cancer risk due to 1 yg/m3 of PCE in air is 4.8 x 10~7. The
fact that the weight-of-evidence for the carcinogenicity of PCE is limited
should be taken into consideration in the use of these estimates.
None of the epidemiologic studies reviewed in this report are adequate
for use in estimatin-g unit risks. If PCE does cause cancer in humans, it is
unlikely that a cohort study on dry cleaners would reveal this fact, since the
incremental lifetime risk to this limited number of workers is too small to be
detected.
The potency index for PCE is 8 x 10°, ranking it in the lowest quartile
of the 54 chemicals that the CAG has evaluated as suspect carcinogens. This
potency index is calculated on the basis of oral exposures.
9-72
-------
9.8 CONCLUSIONS
PCE has induced malignant tumors of the liver in both male and female
B6C3F1 mice. This constitutes a signal that PCE might be a carcinogen for
humans. The technical adequacy and the strong nature of the positive response
in the 1977 NCI bioassay makes it likely that a repeat bioassay would also be
positive. In fact, the recent NTP study, currently under audit, showed similar
positive results. This study, if validated, would influence the evidence for
carcinogenicity of PCE.
According to the Agency's Proposed Guidelines for Carcinogen Risk Assess-
ment (U.S. EPA, 1984), the evidence for PCE carcinogenicity in animals is
limited, and the epidemiologic data is inconclusive. The overall weight-of-
evidence for the human carcinogenic potential of PCE is Group C, i.e., a
possible human carcinogen.
According to a literal interpretation of the criteria of the International
Agency for Research on Cancer (IARC), the animal data supporting the carcino-
genicity of PCE would be classified as limited. Also, since existing epidemi-
ologic data for PCE are inconclusive, its overall IARC ranking would be
classified as Group 3, meaning that the data are inadequate, and PCE cannot be
classified as to its human carcinogenicity.
While the weight of evidence is limited for PCE carcinogenicity (a
possible human carcinogen by proposed EPA criteria and inadequate data for
assessing carcinogenicity by IARC criteria), an upper-bound incremental unit
risk has been estimated for both ingestion and inhalation exposure. This has
been done to help answer the "what-if" question—if PCE is carcinogenic in
humans what is the possible impact upon public health. The upper-bound
incremental unit risk for inhalation of 1 pg PCE/m^ is 4.8 x 10~? and the
9-73
-------
unit risk for ingestion of 1 ^g of PCE/L in drinking water is 1.5 x 10"^.
The potency index for PCE, based on NCI gavage data, places it in the lowest
quartile of the 54 chemicals that the CAG has evaluated as suspect or known
carci nogens.
9-74
-------
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4U.S. GOVERNMENT PRINTING OFFICE: 1986-646-116-40626
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