Unite*::
Environmental Protection
Agency
Office of Water
Regulations and Standards
Criteria and Standards Division
Washington DC 20460
EPA 440-5-80-041
October 1980
Ambient
Water Quality
Criteria for
Dichloroethylenes
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AMBIENT WATER QUALITY CRITERIA FOR
DICHLOROETHYLENES
Prepared By
U.S. ENVIRONMENTAL PROTECTION AGENCY
Office of Water Regulations and Standards
Criteria and Standards Division
Washington, D.C.
Office of Research and Development
Environmental Criteria and Assessment Office
Cincinnati, Ohio
Carcinogen Assessment Group
Washington, D.C.
Environmental Research Laboratories
Corvalis, Oregon
Duluth, Minnesota
Gulf Breeze, Florida
Narragansett, Rhode Island
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DISCLAIMER
This report has been reviewed by the Environmental Criteria and
Assessment Office, U.S. Environmental Protection Agency, and approved
for publication. Mention of trade names or commercial products does not
constitute endorsement or recommendation for use.
AVAILABILITY NOTICE
This document is available to the public through the National
Technical Information Service, (NTIS), Springfield, Virginia 22161.
ii
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FOREWORD
Section 304 (a)(1) of the Clean Water Act of 1977 (P.L. 95-217),
requires the Administrator of the Environmental Protection Agency to
publish criteria for water quality accurately reflecting the latest
scientific knowledge on the kind and extent of all identifiable effects
on health and welfare which may be expected from the presence of
pollutants in any body of water, including ground water. Proposed water
quality criteria for the 65 toxic pollutants listed under section 307
(a)(1) of the Clean Water Act were developed and a notice of their
availability was published for public comment on March 15, 1979 (44 FR
15926), July 25, 1979 (44 FR 43660), and October 1, 1979 (44 FR 55628).
This document is a revision of those proposed criteria based upc.i a
consideration of comments received from other Federal Agencies, State
agencies, special interest groups, and individual scientists. The
criteria contained in this document replace any previously published EPA
criteria for the 65 pollutants. This criterion document is also
published in satisifaction of paragraph 11 of the Settlement Agreement
in Natural Resources Defense Council, et. al. vs. Train, 8 ERC 2120
(D.D.C. 1976), modified, 12 ERC 1833 (D.D.C. 1979).
The term "water quality criteria" is used in two sections of the
Clean Water Act, section 304 (a)(1) and section 303 (c)(2). The term has
a different program impact in each section. In section 304, the term
represents a non-regulatory, scientific assessment of ecological ef-
fects. The criteria presented in this publication are such scientific
assessments. Such water quality criteria associated with specific
stream uses when adopted as State water quality standards under section
303 become enforceable maximum acceptable levels of a pollutant in
ambient waters. The water quality criteria adopted in the State water
quality standards could have the same numerical limits as the criteria
developed under section 304. However, in many situations States may want
to adjust water quality criteria developed under section 304 to reflect
local environmental conditions and human exposure patterns before
incorporation into water quality standards. It is not until their
adoption as part of the State water quality standards that the criteria
become regulatory.
Guidelines to assist the States in the modification of criteria
presented in this document, in the development of water quality
standards, and in other water-related programs of this Agency, are being
developed by EPA.
STEVEN SCHATZOW
Deputy Assistant Administrator
Office of Water Regulations and Standards
iii
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ACKNOWLEDGEMENTS
Aquatic Life Toxicology
William A. Brungs, ERL-Narragansett
U.S. Environmental Protection Agency
Mammalian Toxicity and Human Health Effects
Richard Bull, HERL (author)
U.S. Environmental Protection Agency
Steven D. Lutkenhoff, (doc. mgr.)
ECAO-Cin
U.S. Environmental Protection Agency
Jerry F. Stara (doc. mgr.), ECAO-Cin
U.S. Environmental Protection Agency
Herbert Cornish
University of Michigan
C. Hiremath, CAG
U.S. Environmental Protection Agency
Benjamin Van Duuren
New York University Medical Center
Technical Support Services Staff: D.J. Reisman, M.A. Garlough, B.L. Zwayer,
P.A. Daunt, K.S. Edwards, T.A. Scandura, A.T. Pressley, C.A. Cooper,
M.M. Denessen.
Clerical Staff: C.A. Haynes, S.J. Faehr, L.A. Wade, D. Jones, B.J. Bordicks,
B.J. Quesnel1, P. Gray, R. Rubinstein.
*CAG Participating Members:
Elizabeth L. Anderson, Larry Anderson, Dolph Arnicar, Steven Bayard,
David L. Bayliss, Chao W. Chen, John R. Fowle III, Bernard Haberman,
Charalingayya Hiremath, Chang S. Lao, Robert McGaughy, Jeffrey Rosen-
blatt, Dharm V. Singh, and Todd W. Thorslund.
David J. Hansen, ERL-Gulf Breeze
U.S. Environmental Protection Agency
Roy E. Albert*
Carcinogen Assessment Group
U.S. Environmental Protection Agency
James Bruckner
University of Texas Medical School
Jacqueline Carr
U.S. Environmental Protection Agency
Patrick Durkin
Syracuse Research Corporation
Kris Khanna, ODW
U.S. Environmental Protection Agency
iv
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TABLE OF CONTENTS
Page
Introduction A-l
Aquatic Life Toxicology B-l
Introduction B-l
Effects B-l
Acute Toxicity B-l
Chronic Toxicity B-2
Plant Effects B-3
Miscellaneous B-3
Summary B-3
Criteria B-4
References B-9
Mammalian Toxicity and Human Health Effects C-l
Exposure C-l
Ingestion from Water C-l
Ingestion from Food C-l
Dermal C-3
Pharmacokinetics C-3
Absorption C-3
Distribution C-4
Metabolism C-4
Excretion C-42
Effects C-l2
Acute, Subacute and Chronic Toxicity C-l2
Synergism and/or Antagonism C-l6
Teratogenicity C-l 7
Mutagenicity C-18
Carcinogenicity C-l9
Criterion Formulation C-25
Existing Guidelines and Standards C-25
Basis and Derivation of Criterion C-26
References C-30
Appendix C-39
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CRITERIA DOCUMENT
DICHLOROETHYLENES
CRITERIA
Aquatic Life
The available data for dichloroethylenes indicate that acute toxicity to
freshwater aquatic life occurs at concentrations as low as 11,600 yg/1 and
would occur at lower concentrations among species that are more sensitive
than those tested. No definitive data are available concerning the chronic
toxicity of dichloroethylenes to sensitive freshwater aquatic life.
The available data for dichloroethylenes indicate that acute and chronic
toxicity to saltwater aquatic life occur at concentrations as low as 224,000
ug/1 and would occur at lower concentrations among species that are more
sensitive than those tested. No data are available concerning the chronic
toxicity of dichloroethylenes to sensitive saltwater aquatic life.
Human Health
1,1-Dichloroethylene
For the maximum protection of human health from the potential carcino-
genic effects due to exposure of 1,1-dichloroethylene through ingestion of
contaminated water and contaminated aquatic organisms, the ambient water
concentrations should be zero based on the non-threshold assumption for this
chemical. However, zero level may not be attainable at the present time.
Therefore, the levels which may result in incremental increase of cancer
C C J
risk over the lifetime are estimated at 10 , 10 , and 10" . The
corresponding recommended criteria are 0.33 ug/1, 0.033 ug/1, and 0.003
ug/1, respectively. If the above estimates are made for consumption of
aquatic organisms only, excluding consumption of water, the levels are 18.5
ug/1, 1.85 ug/1, and 0.185 ug/1, respectively.
vi
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1,2-Dichloroethylene
Using the present guidelines, a satisfactory criterion cannot be derived
at this time due to the insufficiency in the available data for 1,2-di-
chloroethylene.
vii
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INTRODUCTION
The dichloroethylenes consist of three isomers: 1,1-dichloroethylene
(1,1-DCE), c i s-1,2-d i ch1oroethy1ene (cis-1,2-DCE) and trans-1,2-dichloro-
ethylene (trans-I,2-DCE). As of the early 1960's, neither of the 1,2-di-
chloroethylenes (1,2-DCEs) had wide industrial usage (Patty, 1963). How-
ever, annual production of 1,1-DCE now approximates 120,000 metric tons
(Arthur 0. Little, Inc., 1976). It is used as a chemical intermediate in
the synthesis of methyl chloroform and in the production of polyvinylidene
chloride copolymers (PVDCs). Among other monomers used with 1,1-DCE in co-
polymer production are vinyl chloride, acrylonitrile, and alky! acrylates.
The impermeability of PVDCs make them useful primarily as barrier coatings
in the packaging industry. Polymers with high vinylidene chloride (1,1-DCE)
content such as Saran are widely used in the food packaging industry. The
heat-seal characteristics of saran coatings make them useful in the manufac-
ture of non-flammable synthetic fiber. 1,1-DCE polymers have also been used
extensively as interior coatings for ship tanks, railroad cars, and fuel
storage tanks, and for coating of steel pipes and structures (Wessling and
Edwards, 1970).
Dichloroethylenes are clear colorless liquids with the molecular formula
C2H2C12 and a molecular weight of 96.96. 1,1-DCE has a water solubil-
ity of 2,500 wg/ml, a vapor pressure of 591 rim Hg, and a melting point of
-122.1°C. The cis isomer of 1,2-dichloroethylene has a water solubility of
3,500 yg/ml, a vapor pressure of 208 mm Hg, and a melting point of -80.5'C;
trans 1,2-dichloroethylene has a water solubility of 6,300 /g/ml, a vapor
presssure of 324 mm Hg and a melting point of -50°C (Patty, 1963; Wessling
A-l
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and Edwards, 1970). The octanol/water partition coefficient for 1,1,-OCE
was reported as 5.37 (log P - 0.73) by Radding, et al. (1977), indicating it
should not accumulate significantly in animals.
1,1-DCE is not known to occur in nature and ambient levels have not been
determined. This chemical is expected to be short-lived in water because of
its volatilization to the atmosphere (Hushon and Kornreich, 1976). The
1,2-DCEs have been measured in a limited number of U.S. drinking water sup-
plies (U.S. EPA, 1978). The population most exposed to 1,1-DCE consists of
workers in industries manufacturing or using 1,1-DCE (Hushon and Kornreich,
1976; Arthur D. Little, Inc., 1976). 1,1-DCE was identified as a cocontami-
nant with vinyl chloride monomer in the working environment of polyvinyl-
chloride productions plants (Kramer and Mutchler, 1972).
A-2
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REFERENCES
Arthur 0, Little, Inc. 1976. Vinylidene chloride monomer emissions from
the monomer, polymer, and polymer processing industries. U.S. Environ.
Prot. Agency, Research Triangle Park, North Carolina.
Hushon, J. and M. Kornreich. 1976. Air pollution assessment of vinylidene
chloride. Contract 68-02-1495. U.S. Environ. Prot, Agency.
Kramer, C. and J. Mutchler. 1972. The correlation of clinical and environ-
mental measurements for workers exposed to vinyl chloride. Am. Ind. Hyg.
Assoc. Jour. 33: 19.
Patty, F.A. 1963. Aliphatic halogenated hydrocarbons, Ind. Hyg. Toxicol.
2: 1307.
Radding, S., et al. 1977. Review of environmental fate of selected chemi-
cals. Contract 68-01-2681. U.S. Environ. Prot. Agency.
U.S. EPA. 1978. Statement of basis and purpose for an amendment to the
national interim primary drinking water regulations on a treatment technique
for synthetic organics. Off. Drinking Water Criteria Stand. Div., U.S.
Environ. Prot. Agency, Washington, D.C.
Wessling, R. and F. G. Edwards. 1970. Poly (Vinylidene Chloride) Iru H.F.
Mark, et al. (eds.), Kirk-Othmer Encyclopedia of Chemical Technology.
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Aquatic Life Toxicology*
INTRODUCTION
All of the available data for dichloroethylenes with one exception are
for 1,1-dichloroethylene. The bluegill has been tested (U.S. EPA, 1978)
with both 1,1-dichloroethylene and 1,2-dichloroethylene under similar condi-
tions, and the 96-hour LC^q values under static conditions were 73,900 and
135,000 wg/1, respectively. Apparently, the location of the chlorine atoms
on the molecule does not affect the acute toxicity of dichloroethylenes very
much.
All of the data on the effects of dichloroethylenes on saltwater organ-
isms are for 1,1-dichloroethylene.
EFFECTS
Acute Toxicity
Two 48-hour tests were conducted with Oaphnia magna with 1,1-dichloro-
ethylene and the ECjq values were 11,600 yg/1 (Dill, et al. manuscri Pt)
and 79,000 wg/1 (U.S. EPA, 1978) (Table 1). The cause of this difference
could not be ascertained.
Dill, et al. (manuscript) tested the fathead minnow and they ooserved
that the 96-hour LCg0 was 169,000 yg/1 using static techniques and 108,000
wg/l using flow-through techniques with measured concentrations. The value
from the static test gave a lower estimate of toxicity than the flow-through
test. The only additional data are for the bluegill, which was discussed
*The reader is referred to the Guidelines for Deriving Water Quality Cri-
teria for the Protection of Aquatic Life and Its Uses in order to oetter un-
derstand the following discussion and recommendation. The following tables
contain the appropriate data that were found in the literature, and at the
bottom of each table are calculations for deriving various measures of toxi-
city as described in the Guidelines.
Q-l
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above; this species is about as sensitive to 1,1-dichloroethylene as is the
fathead minnow.
The data on acute static tests with bluegill under similar conditions
(U.S. EPA, 1978) in this and other documents on structurally related chemi-
cals show a correlation between degree of chlorination and toxicity. The
96-hour LC,jq values for this species are 73,900 and 135,000 ug/1 for 1,1-
and 1,2-dichloroethylene, respectively, 44,700 ug/1 for trichloroethylene,
and 12,900 yg/1 for tetrachloroethylene. These results indicate increase in
the lethal effect on bluegills with an increase in chlorine content. The
correlation with Daphnia magna toxicity is not as clear. The 48-hour values
from the same investigator (U.S. EPA, 1978) are 79,000, 85,200, and 17,700
ug/1 for 1,1-dichloroethylene, trichloroethylene, and tetrachloroethylene,
respectively.
The 96-hour LC^q for the mysid shrimp and 1,1-dichloroethylene is
224,000 ug/1 (Table 1). The 96-hour LC^g for the same species under
similar test conditions (U.S. EPA, 1978) is 10,200 yg/1 for tetrachloroethy-
lene. As was suggested in the freshwater part of this document, acute toxi-
city of these structurally related compounds increases with increasing
degree of chlorination. The 96-hour LC,-0 values (Table 1) for the sheeps-
head minnow (U.S. EPA, 1978) and the tidewater silversides (Dawson, et al.
1977) are 249,000 and 250,000 yg/1, respectively.
Chronic Toxicity
An embryo-larval test has been conducted (U.S. EPA, 1978) with the fat-
head minnow and 1,1-dichloroethylene (Table 2). The range of experimental
concentrations was such that no adverse effects were observed at the highest
test concentration of 2,800 ug/1. The flow-through 96-hour LC^q for this
species was 108,000 ug/1 (DiH» et al. manuscript), suggesting that the dif-
B-2
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ference in concentration between acute and chronic effects for the fathead
minnow is less than 40 times.
No other chronic tests have been conducted on other freshwater or any
saltwater species.
Plant Effects
The 96-hour ECgg values (Table 3), based on chlorophyll a and cell
numbers of the freshwater alga, Selenastrum capricornutum, were greater than
798,000 ug/1 (U.S. EPA, 1978).
There was no effect of 1,1-dichloroethylene on the saltwater alga,
Skeletonema costatum, at test concentrations as high as 712,000 ug/1 (Table
3).
Miscellaneous
Dili, et al. (manuscript) extended their testing with the fathead min-
now and observed a 13-day LC5Q of 29,000 ug/1 for 1,1-dichloroethylene
(Table 4). Their related 96-hour LC5Q value (Table 1) was 108,000 ug/1
which indicates that significant additional mortality occurred between 96
hours and 13 days.
Summary
Most of the data for dichloroethylenes and freshwater organisms are for
1,1-dichloroethylene and Paphnia magna, the fathead minnow, and the blue-
gill, with 50 percent effect concentrations that range from 11,600 to
169,000 ug/1. The toxicity of 1,1- and 1,2-dichloroethylene to the bluegi11
was similar. No chronic effects were observed for the fathead minnow at
concentrations of 1,1-dichloroethylene as high as 2,800 ug/1. The tested
freshwater alga was relatively insensitive to 1,1-dichloroethylene with no
effects at concentrations as high as 798,000 ug/1.
All of the saltwater data are for 1,1-dichloroethylene. The range of
LCgg and EC,-q values for the mysid shrimp, sheepshead minnow, tidewater
B-3
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si 1versides, and an alga is from 224,000 to greater than 712,000 ug/1. No
data are available for any dichloroethylene to estimate chronic toxicity.
CRITERIA
The available data for dichloroethylenes indicate that acute toxicity
to freshwater aquatic life occurs at concentrations as low as 11,600 ug/1
and would occur at lower concentrations among species that are more
sensitive than those tested. No definitive data are available concerning
the chronic toxicity of dichloroethylenes to sensitive freshwater aquatic
life.
The available data for dichloroethylenes indicate that acute toxicity
to saltwater aquatic life occurs at concentrations as low as 224,000 yg/1
and would occur at lower concentrations among species that are more
sensitive than those tested. No data are available concerning the chronic
toxicity of dichloroethylenes to sensitive saltwater aquatic life.
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Table 1. Acuta values for dlchloroethylanas
Spacles Method* Chemical
LC50/EC50
Species Acute
Value (uq/D Reference
FRESHWATER SPECIES
Cladoceran, S, U 1,1~dIch loro-
Daphnla magna ethyIene
Cladoceran, S, U I, I -d I ch loro-
Daphnla magna ethylene
Fathead nlnnow, S, U 1,1-dlchloro-
Plmephales promelas ©thy lane
Fathead minnow, FT, M 1,l-dlchloro-
Plmephales prone I as ethylene
Blueglll, S, U 1,1-dIchloro-
Lepomls macrochlrus ethy lane
Blueglll, S, U 1,2-dIchloro-
Lepomls macrochlrus ethylene
SALTWATER
11,600
79,000
169,000
108,000
73,900
135,000
SPECIES
Mysld shrimp, S, U I,I-dIchloro-
Mysldopsls bah Ia ethylene
Sheepshead minnow, S, U I, I-dlchloro-
Cyprlnodon varlegatus ethylene
Tidewater si Ivers ides, S, U 1,1-dlchloro-
Menldla beryIllna et hy I e ne
224,000
249,000
250,000
30,300
108,000
73,900
135,000
224,000
249,000
250,000
Dili, et al.
Manuscript
U.S. EPA, 1978
01II, et al.
ManuscrIpt
Dl11, et al.
ManuscrIpt
U.S. EPA, 1978
U.S. EPA, 1978
U.S. EPA, 1978
U.S. EPA, 1978
Dawson, et al. 1977
* S = static, FT = t low-through, U a unmeasured, M = measured
No Final Acute Values are calculable since the minimum data base requirements are not met.
B-5
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Table 2* Chronic values for dlchtoroethyl«n«s (U.S. EPA, I97B)
Chronic
Units Value
Specie* Method* Chemical (wg/l) (ug/l)
FRESHWATER SPEOES
Fathead minnow, E-L 1,1-d ichloro- >2,800
Ptwephales prowelas ethylena
* E-L * embryo-larvaI
No acute-chronic ratio Is calculable.
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Table 3. Plant values for dlchloroethylanes (U.S, EPA, 1978)
Species
Chemical
Effect
Alga,
Selenastrum caprlcornutum
Alga,
Selenastrum caprlcornutum
FRESHWATER SPECIES
1,1-dlchloro-
ethy lene
1,1-dlch loro-
ethy lene
EC50 96-hr
ch lorophy 11 _a
EC50 96-hr
ce11 count
SALTWATER SPECIES
Alga,
SkeIetonema costatuw
Alga,
SkeIetonema costatum
1,1-dlchloro-
ethy lene
I,1-dlchloro-
ethy lene
EC50 96-hr
ch lorophy 11 _a
EC50 96-hr
eel I count
Result
(tig/1)
>798,000
>798,000
>712,000
>712,000
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Table 4* Othar dots for dlchloroathylana* (Dill, at al. Manuscript)
Spec I as
Fathead minnow,
PI met) hales promelas
ChawleaI Duration
FRESHWATER SPECIES
1,1-dIch loro- 15 days
athylone
RosuIt
Effact (uq/D
LC50 29,000
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REFERENCES
Dawson, G.W., et al. 1977. The acute toxicity of 47 industrial chemicals
to fresh and saltwater fishes. Jour. Hazardous Mater. 1: 303.
Dill, D.C., et al. Toxicity of 1,1-dichloroethylene (vinylidene chloride)
to aquatic organisms. Dow Chemical Company. (Manuscript)
U.S. EPA. 1978. In-depth studies on health and environmental impacts of
selected water pollutants. U.S. Environ. Prot. Agency. Contract No. 68-01-
4646.
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Mammalian Toxicology and Human Health Effects
EXPOSURE
Ingestion from Water
The National Organic Monitoring Survey (U.S.EPA, 1978a) re-
ported detecting, but did not quantify occurrence of one of the
three dichloroethylenes (DCE) in finished drinking water, 1,1-di-
chloroethylene (vinylidene chloride). Fishbein (1976) indicates
there is no information concerning the amount of 1,1-DCE which may
migrate into water from discarded materials made from vinylidene
chloride polymers. Another source of 1,1-DCE is from the decom-
position of 1,1,1-trichloroethylene (McConnell, et al. 1975) which
has been occasionally detected in drinking water (U.S. EPA, 1978b)
usually at low concentrations (about 1.0 ug/1). The other di-
chlor oethylenes cis-1,2-dichloroethylene (cis-1,2-DCE) and trans-
1,2-dichloroethylene {trans-1,2-DCE) were found at concentrations
of 16 and 1 ug/1 in Miami drinking water (U.S. EPA, 1975, 1978b)
Concentrations of 0.1 ug/1 cis-1,2-DCE were observed in Cincinnat
and Philadelphia drinking waters as well, but were absent fr
seven other drinking waters included in this survey. Much t
little information is available to estimate and assess curr<
exposure of DCE to humans via drinking water.
Ingestion from Food
Food wrappers are commonly made of 1,1-DCE copolymers,
ever, there appear to be no data as to the extent to which tht
treated monomer migrates into foods (Fishbein, 1976). The
are very volatile compounds and are degraded in air with a
life of eight weeks (Pearson and McConnell, 1975).
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A bioconcentration factor {BCF) relates the concentration of
a chemical in aquatic animals to the concentration in the water in
which they live. The steady-state BCFs for a lipid-soluble com-
pound in the tissues of various aquatic animals seem to be propor-
tional to the percent lipid in the tissue. Thus, the per capita
ingestion of a lipid-soluble chemical can be estimated from the
per capita consumption of fish and shellfish, the weighted average
percent lipids of consumed fish and shellfish, and a steady-state
BCF for the chemical.
Data from a recent survey on fish and shellfish consumption
in the United States were analyzed by SRI International (U.S. EPA,
19805. These data were used to estimate that the per capita con-
sumption of freshwater and estuarine fish and shellfish in the
United States is 6.5 g/day (Stephan, 1980). In addition, these
data were used with data on the fat content of the edible portion
of the same species to estimate that the weighted average percent
lipids for consumed freshwater and estuarine fish and shellfish is
3.0 percent.
No measured steady-state bioconcentration factor (BCF) is
ivailable for 1,1-dichloroethylene, but the equation "Log BCF =
0.85 Log P) - 0.70* can be used (Veith, et al. 1979) to estimate
le steady-state BCF for aquatic organisms that contain about 7.6
rcent lipids (Veith, 1980) from the octanol/water partition co-
Eicient (P). The measured log P value was obtained from Dec, et
^Manuscript). When no measured value could be found, a calcu-
sd log P value was obtained using the method described in
ch and Leo { 1979). The adjustment factor of 3 .0/7.6 = 0.395
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Is used to adjust the estimated BCF from the 7.6 percent lipids on
which the equation is based to the 3.0 percent lipids that is the
weighted average for consumed fish and shellfish, in order to ob-
tain the weighted average for the edible portion of all freshwater
and estuarine aquatic organisms consumed by Americans.
Estimated Weighted
Chemical Log P Steady-State Average
Meas. Calc. BCF BCF
1.1-Dichloroethylene 2.18 14.2 5.61
1.2-Dichloroethylene 1.53 4.0 1.58
(cis and trans)
Dermal
There are no data indicating the extent and significance of
human exposure to dichloroethylenes via the skin.
PHARMACOKINETICS
Absorption
No animal or human studies appear to exist which specifically
document the degree of systemic absorption of the DCEs by any
route. However, related compounds such as trichloroethylene and
tetrachloroethylene are absorbed to near the theoretical maximum
for both the inhalation and ingestion routes (Daniel, 1963; Mon-
ster# ¦et al. 1976) . On this basis 35 to 50 percent of inhaled DCE
and virtually 100 percent of ingested DCE may be absorbed system-
icaliy. Such a high degree of absorption of 1,1-DCE seems implied
by the studies of McKenna, et al. (1977a,b) even if not specific-
ally quantified.
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Distribution
Distribution of 1,1-DCE in various organs of the rat has been
studied after inhalation exposure (Jaeger, et al. 1977a) using
14C-labeled 1,1-DCE. The distribution of total radioactivity
30 minutes after cessation of exposure was similar in fed and
fasted rats, although the absolute level of label was considerably
greater in fasted animals. The largest concentrations were found
in kidney, followed by liver, spleen, heart, and brain. Blood
concentrations were high relative to tissue concentrations. A
similar distribution was observed for trichloroacetic acid-in-
soluble (TCA-insoluble), labeled material (presumably bound to
macromolecules).
Subcellular distribution of radio-labeled 1,1-DCE metabolites
30 minutes following a 2-hour exposure to 8 ,000 mg/m3 was ob-
served in the microsomal, mitochondrial, and cytosolic compart-
ments of the liver (Jaeger, et al. 1977a). However, a large
amount of radioactivity in the cytosol was TCA- and CHCI3-
soluble, whereas that in the mitochondria and microsomes was TCA-
insoluble and CHCl3-soluble. These data suggest substantial
binding of 1,1-DCE metabolites to macromolecules and association
with the lipid present in the latter two fractions, respectively.
The turnover of bound TCA-insoluble radioactivity derived from
1,1-DCE has a half-life of 2 to 3 months.
Distribution data on the 1,2-DCE isomers are not available.
Metabolism
The comparative metabolism of chloroethylenes has been ex-
tensively studied. Leibman and Ortiz (1977) have postulated th
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various metabolic pathways for 1,1-DCE, cis-1,2-DCE, and trans-
1,2-DCE as shown in Figure 1.
In addition, there is evidence that the 1,1-DCE metabolites
are conjugated with glutathione, which presumably represents a
detoxification step (McKenna, et al. 1977a). Bonse, et al. (1975)
identified both the chloroethanol and chloroacetic acid deriva-
tives of cis-1,2-DCE and trans-1,2-DCE in a perfused rat liver.
The metabolism of the cis-isomer relative to the amount taken up
by the liver was much greater than the trans-isomer. Leibman and
Ortiz (1977) identified formation of chloroacetic acid from
1,1-DCE, but as yet the dichloroacetaldehyde has not been unequiv-
ocally identified as a metabolite of 1,1-DCE. Inhibition of epox-
ide hydrase resulted in a stimulation of chloroacetic acid forma-
tion from 1,1-DCE, leading to the conclusion that the glycol
intermediate is relatively unimportant in the conversion of
1,1-DCE to chloroacetic acid (Leibman and Ortiz, 1977). The
essential feature of each of these metabolic pathways is that all
chloroethylenes appear to be metabolized through expoxide inter-
mediates which are reactive and may form covalent bonds with tis-
sue raacromolecules (Henschler, 1977a).
In the intact animal, a large portion of the systemically
absorbed 1,1-DCE is metabolically converted. Jaeger, et al.
(1977a) found 36.7 + 3.3 percent of the absorbed dose in the urine
of rats within 26 hours. Disposition of l^C-activity from
radio-labeled 1,1-DCE in mice and rats over a 48-hour period
(McKenna, et al. 1977b) is shown in Table 1.
C-5
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a a os qs
\ / \ i % /
xCs»c c-—c —ci c—c — ci
/ \ / i i \
ci a ci a so a
monochloro- monochloro—
acetyl chloride acetic acid
\ X
1,1-DCE CI s cl
\ t 1
c —C 2—G—G
/\ /I L \
cx o a ci a
epoxide dichloroacetaldehyda
I
intermediate
cl a o a o a
1 1 * ' ~ 1 1
C1 i i / i j i
oh oa cl a ao a
glycol monochloro- monochloro-
intermediate acetyl chloride acetic acid
a a
a c ——c^° ci c
I \ *" I ' \
OL CL a OH
monochloro— monochloroacetic
acetyl chloride acid
/
\ Av /
epoxide
Intermediate
C£s-1,2-DCE CL
1 /
a—c —c'
i Na
a
dichloroacetaldehyda.
Cl—c —
» \
epoxide CL Cl 3 QH
£rans-l,2-DCE intermediate monochloro— monochloroacetic
acetyl chloride acid
FIGURE 1
Proposed Metabolic Pathways of 1,1-DCE
and the 1,2-DCE Isomers
Source; Leibman and Ortiz, 1977
C-6
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TABLE 1
Disposition of 1,1-DCE Following Inhalation of
40 mg/m^ Over a 40-hour Period*
Mice Rats
Expired 1,1-DCE %
0.65
+
0 .07
1.63
+
0 .14
Expired ^^C02
4.64
+
0 .17
8.74
+
3.72
Body Burden %
Urine
80.83
+
1.68
74 .72
+
2.30
Feces
6.58
+
0 .81
9.73
+
0 .10
Carcass
5.46
+
0 .41
4.75
+
0 .78
Cage Wash.
1.83
+
0 .84
0.44
+
0 .28
Body Burden mg. eq.
1,1-DCE/kg
5.30
+
0.75
2.89
+
0 .24
Total Metabolized
mg . eq. 1,1-DCE/kg
5.27
+
0 .74
2.84
+
0 .26
*Source: McKenna, et al. 1977b.
C-7
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It is clear from these data that the mouse develops a higher
body burden of 1,1-DCE than the rat at identical exposure levels.
The disposition of 1,1,-DCE appears quite similar in the two
species. However, as a result of the overall greater rate of
metabolism, covalently bound 1,1-DCE metabolites are higher in the
mouse than in the rat as shown in Table 2. The substantial dif-
ference in distribution may be related to the differing sensitiv-
ity of the species to carcinogenic effects of 1,1-DCE (Hathaway,
1977) .
In a study where mice and rats were administered an oral dose
(50 mg/kg) of 1,1-DCE, mice were found to metabolize a greater
proportion of DCE (Jones and Hathaway, 1978). Although DCE was
metabolized in much the same way in both species (Table 3), mice
formed considerably more of the N-acetyl-S-cysteinyl acetyl deriv-
ative, and excrete a small amount of N-acetyl-S-(2-carboxymethyl)-
cysteine, which was not found in the rat.
Jones and Hathaway suggested that the efficiency of DCE
metabolism in rats and mice follows the activity of cytochrome
P-450 in the organs of these animals, and that "real exposure"
(expressed in the amount of DCE metabolized) is relatively higher
for orally dosed mice than rats.
It is notable that in both species, covalently bound metabo-
lites of 1,1-DCE are highest in kidney followed by liver. Expo-
sure of rats by inhalation to a higher concentration (8,000
mg/m3) for a shorter period (two hours) appears to give somewhat
differing results. In this case, a high degree of TCA-insoluble
l^C-activity is observed in the liver. These differences
C-8
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TABLE 2
Covalently Bound l^C-Activity in Rat and Mouse Following
Exposure to 40 mg/ra^ ^4C-1,1-DCE*
^C-l,1 ,-DCE, ug eq/mg Protein
Liver Kidney
Mice 22.29 + 3.77 79.55 + 19.11
Rats 5.28 + 0.14 13.14 + 1.15
~Source: McKenna, et al. 1977b.
C-9
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TABLE 3
Relative Proportion of (-^C) Excretory Products After Oral
Administration of 50 mg/kg of (1-14C)DCE to Rodentst
Expressed
as % of Dose
(14q Excretory Products Mice* Rats*
Unchanged DCE^pulmonary 6 28
CO2 Jexcretion 3 3.5
Chloroacetic acid 0 1
Thiodiglycollic acid 3 22
Thioalycollic acid 5 3
Dithioglycollic acid 23 5
Thioglycollyloxalic acid 3 2
N-Acetyl-S-cysteinyl acetyl
derivative 50 28
N-Acetyl-S-(2-carboxymethyl)
cysteine 4 0
Urea 3 3.5
tSource: Jones and Hathaway, 1978.
*Alderly Park strains.
C-10
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between long- and short-term inhalation exposures must be kept in
mind when evaluating the long-term toxicity of 1,1-DCE.
The relationship of metabolism of DCEs to their toxicity is
not well understood. Differing results are obtained with inducers
and inhibitors of microsomal enzymes depending upon the age or
body weight of experimental animals (Anderson and Jenkins, 1977).
Pretreatment with dithiocarbamates, which presumably interfere
with 1,1-DCE metabolism, protects against lethal and hepatotoxic
damage (from 1,1-DCE) and reduces covalent binding of metabolites
to tissue macromolecules (Short, et al. 1977a) On the other hand,
Aroclor 1254 or phenobarbital pretreatment, which would increase
microsomal enzyme activity, also decreases the hepatotoxicity of
1,1-DCE (Reynolds, et al. 197 5; Jenkins, et al. 1972) . However,
Carlson and Fuller (1972) reported increased mortality from
1,1-DCE in rats following phenobarbital pretreatment. The possi-
bilities that these pretreatments either intereact with more than
one metabolic pathway or that differing mechanisms account for
hepatotoxicity and lethality of 1,1-DCE, or both, could account
for the inconsistent responses (Anderson and Jenkins, 1977; Rey-
nolds and Moslen, 1977). It is clear that the hepatotoxicity of
1,1-DCE does increase with decreasing concentrations of hepatic
glutathione (Jaeger, et al. 1973). This seems involved with the
greater sensitivity of fasted animals to 1,1-DCE-induced hepato-
toxicity (Jaeger, et al. 1974). In regard to this issue, Hathaway
(1977) has identified thiodiglycollic acid, thioglycollic acid,
dithioglycollic acid, and an N-acetyl-S-cysteinyl-acetyl deriva-
tive as products of 1,1-DCE metabolism in rats and mice.
C-ll
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Excretion
There appear to be no data relating to the rate at which any
of the DCEs are cleared from the body. In the case of 1,1-DCE one
can guess that the rate of elimination is relatively rapid in that
substantial fractions of the total absorbed dose may be recovered
in urine within 26 or 72 hours (Jaeger, et al. 1977a; McKenna, et
al. 1977a). However, these data do not allow a more precise esti-
mate of turnover. Disappearance of covalently bound metabolites
of 1,1-DCE (measured as TCA-insoluble fractions) appears to be
fairly rapid as well, with a reported half-life of 2 to 3 hours
(Jaeger, et al. 1977a). These data were obtained from experiments
of insufficient duration and resolution of the nature of the bind-
ing to rule out the possibility of a residual binding to tissue
macromolecules of an irreversible or slowly reversible character.
No data exist concerning the excretion of the cis- or trans-
isaners of 1,2-DCE.
EFFECTS
Acute, Subacute, and Chronic Toxicity
Like other members of the chlorinated ethylene series, the
DCEs possess anesthetic properties. Unlike those compounds of the
group which were employed as general anesthetics, very little
effort has gone into characterizing either short- or long-term CNS
toxicity of the DCEs. Irish (1962) referenced unpublished data
concerning the CNS depressant activity of 1,1-DCE in humans. A
concentration of 16,000 mg/m^ was estimated as sufficient to
rapidly produce a state resembling drunkenness and was felt likely
C-12
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to result in unconsciousness if exposure was continued. A single
report has been made associating sensory disturbances in the area
of the face controlled by the trigeminal nerve and accidental ex-
posure to 1,1-DCE (Broser, et al. 1970). Subsequent evaluation,
however, suggested that the toxic agent was either mono- or di-
chloroacetylene which was a contaminant of 1,1-DCE (Henschler, et
al. 1970). No further information exists concerning the central
nervous system toxicity of these compounds.
In recent years, considerable attention has focused on the
liver and kidney damage produced by 1,1-DCE. Prendergast, et al.
(1967) documented morphological damage to kidney and liver of rats
and guinea pigs exposed to 1,1,-DCE at 189 mg/'m^. These effects
were associated with an increase liver lipid concentration in
rats. In guinea pigs, continuous 90—day exposure to concentra-
tions as low as 20 mg/m- produced increased mortality, whereas
intermittent exposures (30 exposures, 8 hours/day, 5 days/week) at
395 jng/in- produced no increase in mortality. A similar differ—
a i m t' m ^ a v*fn i ^ n V av 11 f i mm a m a m . a ^ J J _
X A A VCL gua w«Jiiv.i.llUUU9 wad VJIJ L veu 111
i> U/\ m am r.i vrt i 9 0 A/4 1*1 An* f a 1 *i 4>«* v.*« a J
ill jr ill WL wa o WU a CZ I. V CS U WJ.UJ1 tUllt-iHUUU5
a vn ^enra a +• a r\r»r»o n b »-» h l rsn rs-f 10 1 m/i /m 3 xta m A v «*
W A^/ V/4U4i w M V4. A V A f ill + ^ ^ 111 WL WQX X W
e ui fh fho c^nio in i a vnrsen ro nfi 1 i >70^ -P av mi 1
U V fe V »* * W*4 wtitv w»a V WW Wk V Wl U * A A ttV W i. Vb ^ u
Results of a 90—day toxicity study in rats given 1,1-DCE in
water or in air revealed cytoplasmic vacuolization in liver cells
of rats ingesting 200 ppm or inhaling 25 ppm of the compound. The
hepatocellular changes were interpreted to be of a reversible
character (Quast, et al. 1977).
C-13
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Oral administration of single doses of 1,1-DCE at 200 and
400 mg/kg decreases liver glucose-6-phosphatase, increases liver
alkaline phosphatase, liver tyrosine transaminase, plasma alkaline
phosphatase, and plasma alanine transaminase (Jenkins, et al.
1972) 20 hours after administration. The LD50 of 1,1-DCE was
decreased in adrenalectomized rats (1,550 mg/kg in sham-operated,
84 mg/kg in adrenalectomized animals) by a factor of 20. This
effect was not clearly related to the hepatotoxicity of 1,1-DCE
(Jenkins, et al. 1972).
The hepatotoxic effects of 1,1-DCE appear to be potentiated
by depletion of hepatic glutathione concentrations, whether re-
sulting from normal diurnal variation (Jaeger, et al. 1973) or by
fasting (Jaeger, et al. 1974, 1975). In the latter case, it was
possible to demonstrate increased concentrations of 1,1-DCE in
blood and liver in the fasted animals. Thyriodectomy decreased
whereas thyroxine administration exacerbated the hepatotoxic
effects of 1,1-DCE (Szabo, et al. 1977; Jaeger, et al. 1977b). At
the same time thyroidectomy increased, and thyroxine decreased
liver glutathione concentrations. Jaeger (1977) has suggested
that the hepatotoxic effects of 1,1-DCE are secondary to a reduc-
tion in mitochondrial glutathione (and sulfhydryl enzymes) and
marked inhibition of mitochondrial respiration.
Jenkins, et al (1972) found both cis- and trans-1,2-DCE to be
considerably less potent than 1,1-DCE as a hepatotoxin. Freundt,
et al. (1977) indicated that repeated inhalation exposures of 800
mg/m^ (8 hours/day, 5 days/week, 16 weeks) of the trans-1,2-DCE
produces fatty degeneration of the liver.
C-14
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Less attention has been paid to the renal toxicity of the
DCEs despite the occurence of histologically demonstrated damage
at 1,1-DCE exposures equal to or less than those required for
hepatotoxicity (Prendergast, et al. 1967; Short, et al. 1977a).
The degree of covalent binding of 1,1-DCE metabolites has been
shown to be higher in kidney than in liver i n both. rats and mice
(McK.enna, et al, 1977b; Short, et al, 1977a) suggesting that renal
damage resulting from 1,1-DCE should be more carefully examined in
the future. No quantitative data appear available on the nephro-
toxicity of 1,2-DCEs. Inhalation of 1,1-DCE at high concentra-
tions (102,000 mg/m3) for short periods of time (10 minutes) has
been found to sensitize the myocardium to arrhythmias produced by
inj ection of epinephrine (Siletchnik and Carlson, 1974) . Unlike
hepatotoxicity, the cardiac sensitizing effects of 1,1-DCE were
enhanced by phenobarbital pretreatment. This implies that a
metabolite of 1,1-DCE may be involved. Similar effects have been
observed in human poisoning with trichloroethylene. The 1,2-DCE
isomers have not been investigated with respect to this effect.
Only one epidemiological study has been published which ex-
amined workers exposed to 1,1-DCE uncomplicated by exposures to
other solvents (Ott, et al. 1976). At the time of this prelimi-
nary report, no abnormal findings could be associated with 1,1-DCE
exposure in a population of 138 workers. Measured concentrations
in the workplaces of these individuals ranged from 9 to 280
mg/m3 (time-weighted averages).
As reported with a number of other chlorinated hydrocarbon
solvents, 1,1,-DCE increases bile-duct pancreatic fluid flow in
C—15
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fasted rats (Hamada and Peterson, 1977) . The composition of the
fluid did not differ significantly from that in control animals.
The significance of this finding remains to be established.
Synergism and/or Antagonism
DCEs are metabolically converted to reactive epoxide inter-
mediates (Bonse, et al, 1975; Hathaway, 1977; McKenna, et al.
1977b). Consequently, it would be predicted that compounds which
increase or decrease the rate of DCE metabolism would affect toxi-
city. However, interactions at this level do not at present lend
themselves to simple prediction. Compounds which decrease coval-
ent binding of 1,1-DCE metabolites, such as disulfiram, protect
against lethality and hepatotoxicity resulting from 1,1-DCE expo-
sure (Short, et al. 1977a). On the other hand, pretreatment of
animals with inducers of microsomal enzyme systems, such as Aro-
clor 1254 or phenobarbital, appear to decrease the hepatotoxicity
due to 1,1-DCE (Reynolds, et al. 1975; Jenkins, et al. 1972), but
increase mortality (Carlson and Fuller, 1972). These conflicting
findings suggest that metabolism of the DCEs is not sufficiently
understood to allow straightforward predictions. It is likely
that the degree of conjugation of reactive intermediates with glu-
tathione may be the factor not yet taken into account in attempt-
ing to predict the impact of altering microsomal enzyme activities
on toxicity (Hathaway, 1977) .
Although tissue glutathione concentrations affect the hepato-
toxicity of 1,1-DCE (Jaeger, et al. 1973, 1977b), decreased tissue
C-16
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glutathione is associated with greater toxicity and elevated glu-
tathione with decreased toxicity in response to 1,1-DCE challenge.
Thyroidectomy decreases and adrenalectomy greatly potentiates
the toxicity of 1,1-DCE (Szabo, et al. 1977; Jenkins, et al.
1972), by altering glutathione levels.
Teratogenicity
Murray, et al. (1979) evaluated the teratogenic potential of
inhaled or ingested 1,1-DCE in Sprague-Dawley rats and New Zealand
white rabbits. Inhalation exposure for both species was for 7
hours/day at 20 (rats only), 80, and 160 ppm. In the ingestion
study, rats were given drinking water with 200 ppm DCE, or ap-
proximately 40 mg/kg/day. Administration for rats was on days 6
to 15 of gestation, and for rabbits, the 6 th to 18th day. In
rats, inhalation of 80 to 160 ppm of DCE produced significant
maternal effects including decreased weight gain, decreased food
consumption, increased water consumption, and increased liver
weight (160 ppm only). In the offspring, there was a significant-
ly increased incidence of skeletal alterations at 80 and 160 ppm;
these alterations included wavy ribs and delayed ossification of
various bones. In rabbits, 160 ppm caused a significant increase
in resorptions in the dams, and in the offspring, a significant
change in several minor skeletal variations. In both rats and
rabbits exposed to 1,1-DCE by inhalation, Murray, et al. (1979)
noted that concentrations which caused little evidence of maternal
toxicity (20 ppm in the rat and 80 ppm in the rabbit) caused no
adverse effect on embryonal or fetal development.
C-17
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In rats receiving DCE by ingestion, the only significant
effect noted was an increase in mean fetal crown rump length
(Murray, et al. 1979).
Mutagenicity
1,1-DCE has been shown to be mutagenic in Salmonella typhi-
murium strains TA1530 and TA100 (Bartsch, et al. 1975) and E. coli
K12 (Greim, et al. 1975) . In both systems mutagenic activity
required microsomal activation. Pretreatment with phenobarbital
increased the mutagenic activity produced by microsomal fractions
derived from liver, kidney, or lung (Bartsch, et al. 1975).
Microsomal preparations, particularly liver, were considerably
more active when derived from mice than from rats (Bartsch, et al.
1975) .
Both the cis- and trans-isomers of 1,2-DCE were nonmutagenic
when assayed with E^. coli K12 at similar concentrations used for
1,1-DCE (Greim, et al. 1975). Henschler (1977a) and his associ-
ates have suggested that the mutagenic and presumably carcinogenic
activities of the chloroethylene series are related to the unsym-
metrical chlorine substitution or the respective epoxide inter-
mediates. Such substitution would result in less stable and more
reactive intermediates than symmetricaly substituted epoxides.
These data support this hypothesis, at least with respect to muta-
genesis in the E. coli K12 system. However, generalization of
this hypothesis to carcinogenesis in intact animals is not yet
possible and requires modification to account for the demonstrated
carcinogenic activity of tetrachloroethylene [National Cancer In-
stitute (NCI), 1977]. In addition, both 1,1,-DCE and cis-l,2-DCE
C-18
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were found mutagenic in Salmonella tester strains (Cerna and
Kypenova, 1977). The trans-1,2-DCE isomer was found inactive. Of
the three DCEs, only cis-1,2-DCE promoted chromosomal aberrations
in cytogenic analysis of bone marrow cells with repeated i.p.
injections (Cerna and Kypenova, 1977).
The finding of increased mutation rates in bacterial systems
has not been confirmed in mammalian systems. Adult CD male rats
exposed to 220 mg 1, l-DCE/m^ for 6 hours/day, 5 days/week for 11
weeks failed to produce dominant lethal mutations (Short, et al.
1977b). Similar results have been reported by Anderson, et al.
(1977) in dominant lethal studies in CD-I mice.
Carcinogenicity
The carcinogenicity of 1,1-DCE is currently being evaluated
in studies sponsored by the National Cancer Institute (1978). No
results are yet available. Maltoni, et al. (1977) and Maltoni
(1977) have reported preliminary results with inhalation exposures
to 1,1-DCE. Exposure conditions were 4 hours/day, 4 to 5 days/
week, for 52 weeks to 100 mg 1, l-DCE/m^. Animals at the time of
the report had been observed for a total of 98 weeks. At this
concentration 17 of these mice had developed kidney adenocarcino-
mas. No kidney adenocarcinomas had been observed in the control
animals. The majority of tumors were observed in male mice as
shown in Table 4. In this same study, no kidney adenocarcinomas
were observed in Sprague-Dawley rats at exposure up to 800 mg/m^
1 ,1,-DCE.
C-19
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TABLE 4
Kidney Adenocarcinomas in Swiss Mice Exposed to 100 mg/m3
1,1-DCE Starting at 9 Weeks of Age*
Total Animals No. Animals with
1,1-DCE
Sex
at Risk**
Adenocarcinoma
%
None
Male
54
0
0
None
Female
49
0
0
100 mg/m3
Male
78
16
20 .5
100 mg/m3
Female
65
1
1.5
*Source: Maltoni, et al. 1977.
**Defined here as the number of animals that had died since the
appearance of the first tumor; or (survivors at the time of the
first tumor) - (survivors at the time these preliminary data
were reported).
C-20
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Maltoni, et al (1977) also observed a significant increase in
mammary adenocarcinomas in Swiss mice inhaling 100 mg 1,1-DCE/m^
and Sprague-Dawley rats exposed to 600 mg/m^ under the same con-
ditions. The mouse data are presented in Table 5 and the rat data
in Table 6. Experiments in female Sprague-Dawley rats exposed to
20 mg 1,1-DCE by gavage 4 to 5 days/week for 52 weeks resulted in
a 42 percent incidence of mammary tumors, whereas control animals
had a 34 percent incidence. These latter data were not analyzed
statistically. Finally, these authors also found that hamsters
exposed to the same conditions as the Swiss mice failed to exhibit
an increased tumor incidence.
In another study, Lee, et al. (1977) observed a small in-
crease in hepatic hemangiosarcomas in mice exposed to 220 mg/m3
1,1-DCE, 6 hours/day, 5 days/week for 7 to 12 months. Although
kidney pathology was observed, no mention was made of kidney
adenocarcinomas.
Rampy, et al. (1977) exposed male and female Sprague-Dawley
rats to 200 mg 1,1-DCE/l in drinking water (two years) or 100 and
300 mg/m-* 1,1-DCE by inhalation (6 hours/day, 5 days/week for 18
months). In their interim report, there was no evidence (based on
total tumor count) of increased tumor incidence in animals treated
with 1,1-DCE. An unpublished final report of the Rampy study by
Humiston, et al. (1977) agreed with the interim conclusion.
The only human data concerning possible carcinogenic effects
of 1,1,-DCE in man appeared in the epidemiological study of
approximately 30 employees by Ott, et al. (1976). No associations
C-21
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TABLE 5
Incidence of Mammary Adenocarcinomas in Female Swiss Mice
Receiving 1,1-DCE via Inhalationt
0 mg/n»3
40 mg/m^
0 mg/m^
100 mg/m^
Age at start
of test
(weeks)
16
16
9
9
Number with
tumor s/num-
ber at risk*
2/52
1/20
0/49
7/65
Incidence
3.8
5.0
0
*—•
o
•
03
p-values**
0 .63
0.017
tSource; Maltoni, et al. 1977.
*Number of risk defined here as the number of animals that had
died since the appearance of the first tumor (a kidney adenocar-
cinoma); or (survivors at time of the first tumor) - (survivors
at the time these preliminary data were reported).
**Calculated using Fisher exact test and matched control inci-
dence .
C-22
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Incidence
of Mammary
TABLE 6
Tumors in
Rats Inhaling 1,1-DCE*
0
40
100
200
400
600
Number with tumors
32/100
15/30
12/30
15/30
18/30
35/60
Number initially
Incidence
0.32
0.05
0.40
0.50
0.60
0.58
p-Values**
0.058
0.27
0.058
0.0058
0.001
~Source: Maltoni, et al. 1?77.
~~Calculated using the Fisher exact test.
C-23
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could be made between cancer deaths and exposure to 1,1-DCE. The
population was too small to evaluate the carcinogenicity of
1,1-DCE.
C-24
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CRITERION FORMULATION
Existing Guidelines and Standards
Standards that have been established for the DCEs are applic-
able primarily to occupational exposures. The threshold limit
values (TLV) for these compounds presently established by the
American Conference of Governmental Industrial Hygienists (ACGIH)
are 40 mg/m3 (1,1-DCE) and 790 mg/m3 (1,2-DCE). These values
allow daily exposures of 286 mg 1,1-DCE and 5 ,643 mg 1,2-DCE.
These calculations are based on the assumption of a 50 m3/work
week of inhaled air averaged over a 7-day period (Stokinger and
Woodward, 1958). A separate standard has been established for
1,1-DCE, but the TLV does not distinguish between the two isomers
of 1,2-DCE.
The standard for 1,2-DCE was established on the basis of no
measurable effects on growth, mortality, organ and body weights,
hematology, clinical chemistry, and gross and microscopic pathol-
ogy at doses of up to 4,000 mg/m3 for six months in rats, rab-
bits, guinea pigs and dogs (ACGIH, 1977) . However, more recent
data (Freundt, et al. 1977) indicate that 16 weeks of exposure at
the TLV of 790 mg/m3 of trans-1,2-DCE produces histological evi-
dence of fatty degeneration of the liver in rats.
in the case of 1,1-DCE, the standard was established pri-
marily on the basis of the work of Prendergast, et al. {1967) who
observed increased mortality as a result of continuous 90-day
exposures of 1,1-dichloroethylene to guinea pigs at 20 mg/m3 or
to monkeys at 101 mg/m3. Liver and kidney pathology were ob-
served at 189 mg/m3 in rats and guinea pigs. As can be seen
C-25
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these industrial TLVs allow very little safety factor for sensitive
populations. Moreover, recent data suggesting that 1,1-DCE is car-
cinogenic in mice (Maltoni, 1977) have not yet been taken into
account.
Basis and Derivation of Criterion
The use of TLV data assumes an 8-hour day, time-weighted aver-
age, occupational exposure in the work place with workers inhaling
the toxic substance throughout such period. Exposures for the gen-
eral population should be considerably less. Such worker-exposure
inhalation standards are inappropriate for the general population
since they presume an exposure limited to an 8-hour day, an age
bracket of the population that excludes the very young and the very
old, and a healthy worker prior to exposure. Ingestion data are
far superior to inhalation data when the risks associated with the
food and water environment are being considered. Recent data sug-
gest that the ACGIH estimate of noncarcinogenic risks resulting
from exposure to the DCEs may approximate effect levels for
1,1-DCE. Additionally, it is recognized that the ACGIH standards
apply primarily to healthy adult worker populations and do not
incorporate safety factors for sensitive populations. In order to
provide a wider margin of safety, calculations of acceptable con-
centrations of DCEs in drinking water as proposed by Stokinger and
Woodward (1958) include a safety factor of 100 and are illustrated
as follows:
1,1-DCE 40 mg/m^ x 50 m^/week x 0.40*
7 days/week x 100** = 1'14 m<3/daY
* Estimated coefficient of absorption via inhalation vs. ingestion.
**Safety factor for sensitive populations.
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The safety factor of 100 was based upon the rationale of the
National Academy of Sciences (NAS) for noncarcinogenic substances
where limited human data and valid animal studies exist (NAS,
1977) . The absorption coefficient is based on the data of Mon-
ster, et al. (1976) for the related compound, trichloroethylene.
Assuming a 2 liter daily consumption of drinking water, concentra-
tions of 1,1—DCE should be limited to 0.6 mg/1 (1.14 mg/day/2
liters) on the basis of noncarcinogenic risks.
However, under the Consent Decree in NRDC v. Train, criteria
are to state "recommended maximum permissible concentrations
(including where appropriate, zero) consistent with the protection
of aquatic organisms, human health, and recreational activities."
1,1-Dichloroethylene is suspected of being a human carcinogen.
Because there is no recognized safe concentration for a human car-
cinogen, the recommended concentration of 1,1-dichloroethylene in
water for maximum protection of human health is zero.
Because attaining a zero concentration level may be infeas-
ible in cases and in order to assist the Agency and States in the
possible future development of water quality regulations, the con-
centrations of 1,1-dichloroethylene corresponding to several in-
cremental lifetime cancer risk levels have been estimated. A can-
cer risk level provides an estimate of the additional incidence of
cancer that may be expected in an exposed population. A risk of
10~5 for example, indicates a probability of one additional
case of cancer for every 100,000 people exposed, a risk of
10"® indicates one additional case of cancer for every million
people exposed, and so forth.
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In the Federal Register notice of availability of draft am-
bient water quality criteria, EPA stated that it is considering
setting criteria for 1,1-dichloroethylene at an interim target
risk level of 10~5, 10~6, or 1G"7 as shown in the fol-
lowing table.
Risk Levels
and Corresponding Criteria (1)
10-7 10-6 10-5
Exposure Assumtpions
(per day)
2 liters of drinking
water and consumption
of 6.5 g fish and
shellfish. (2)
Consumption of fish and
shellfish only.
0.003 ug/1
0.185 ug/1
o.033 ug/1 0.33 ug/1
1.85 ug/1 18.5 ug/1
(1) Calculated by applying a linearized multistage model, as
described in the Human Health Methodology Appendices to
the October 1980 Federal Register notice which announced
the availability of this document, to the animal bio-
assay data presented in the Appendix and in Table 4.
Since the extrapolation model is linear at low doses,
the additional lifetime risk is directly proportional to
the water concentration. Therefore, water concentra-
tions corresponding to other risk levels can be derived
by multiplying or dividing one of the risk levels and
corresponding water concentrations shown in the table by
factors such as 10, 100 , 1 ,000 , and so forth.
(2) Approximately 2 percent of the 1,1-dichloroethylene ex-
posure results from the consumption of aquatic organisms
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which exhibit an average bioconcentration potential of
5.61-fold. The remaining 98 percent of 1,1-dichloro-
ethylene exposure results from drinking water.
Concentration levels were derived assuming a lifetime expo-
sure to various amounts of 1,1-dichloroethylene (1) occurring from
the consumption of both drinking water and aquatic life grown in
waters containing the corresponding 1,1-dichloroethylene concen-
trations and (2) occurring solely from consumption of aquatic life
grown in the waters containing the corresponding 1,1-dichloroethy-
lene concentrations. Because data indicating other sources of
1,1-dichloroethylene exposure and their contributions to total
body burden are inadequate for quantitative use, the figures re-
flect the incremental risks associated with the indicated routes
only.
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APPENDIX
Derivation of Criterion for 1,1-Dichloroethylene
Maltoni (1977) exposed Swiss mice via inhalation to 25 ppm
(group IV in his Table 7) of 1,1,-dichloroethylene for 4 hrs/day,
4.5 days/week for 52 weeks. The interim report summarized the re-
sults after 8 2 weeks. The number of males alive at the time of
the first kidney adenocarcinoma (55 weeks) were 98 and 70 in the
treated and control (group VII) groups, respectively. At 82 weeks
20 and 16 animals were alive in treated and control groups, re-
spectively, and the remaining animals were examined histologic-
ally. The results were that 16 of treated males had kidney adeno-
carcinomas, whereas none of the control animals had these tumors.
The average dose in mg/kg/day is calculated from the concen-
tration (25 ppm) and the breathing rate (assumed to be a standard
rate of 0.0375 m-^/day for 30 gm mice) as follows, where 1 ppm of
dichloroethylene is assumed to be equivalent to 4 mg/m^ in air:
The lifetime average dose is:
25 ppm x (4/24) x (4.5/7) x (52/82) = 1.699 ppm
Converting to mg/kg/day:
D = 1.699 ppm x (4 mg/m^ per ppm) x (0.0375 m^ of air/day
x (1/0.03 kg) = 8.502 mg/kg/day.
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The parameters of the risk extrapolation are:
Dose Incidence
mn /lrn /rlaul 3 I no rpsnnnHi nn/nn feshodl
—-or —of -r / r ,
Q,0 0/54
ft 16/78
le » 52 weeks w = 0,030 kg
Le = 82 weeks R = 5.61 1/kg
L =90 weeks
With these parameters the carcinogenic potency factor for
humans, qi*, is 1.04 (mg/kg/day)~1. This leads to the esti-
— ^ - —
mate that the water concentration should be less than 0.33 ug/1 in
order to keep the lifetime risk to 1,1-dichloroethylene less than
10~5.
aThis value of dose has already been adjusted for the length of
the experiment (i.e., it has been multiplied by le/Le).
* U. S GOVERNMENT PRINTING OFFICE : 720-016/5959
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