June, 2011
This document is a Final Agency/Interagency Science Discussion draft. It has not been
formally released by the U.S. Environmental Protection Agency and should not at this stage be
construed to represent Agency position on this chemical. It is being circulated for review of its
technical accuracy and science policy implications.
0199
Trichloroethylene; CASRN 79-01-6; 00/00/0000
Human health assessment information on a chemical substance is included in IRIS only
after a comprehensive review of toxicity data by U.S. EPA health scientists from several
program offices, regional offices, and the Office of Research and Development. Sections I
(Health Hazard Assessments for Noncarcinogenic Effects) and II (Carcinogenicity Assessment
for Lifetime Exposure) present the positions that were reached during the review process.
Supporting information and explanations of the methods used to derive the values given in IRIS
are provided in the guidance documents located on the IRIS website at
http://www.epa.gov/iris/backgr-d.htm.
STATUS OF DATA FOR Trichloroethylene
File First On-Line 03/31/1987
Category (section) Status
Last Revised
Chronic Oral RfD Assessment (I. A.) on-line
00/00/0000
Chronic Inhalation RfC Assessment (I.B.) on-line
00/00/0000
Carcinogenicity Assessment (II.) on-line
00/00/0000
I. HEALTH HAZARD ASSESSMENTS FOR NONCARCINOGENIC EFFECTS
I.A. REFERENCE DOSE (RfD) FOR CHRONIC ORAL EXPOSURE
Substance Name - Trichloroethylene
CASRN-79-01-6
Section I. A. Last Revised — 00/00/0000
1
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The RfD is an estimate (with uncertainty spanning perhaps an order of magnitude) of a
daily oral exposure to the human population (including sensitive subgroups) that is likely to be
without an appreciable risk of deleterious effects during a lifetime. The RfD is intended for use
in risk assessments for health effects known or assumed to be produced through a nonlinear
(presumed threshold) mode of action. It is expressed in units of mg/kg-day. Please refer to the
guidance documents at http://www.epa.gov/iris/backgr-d.htm for an elaboration of these
concepts. Because RfDs can be derived for the noncarcinogenic health effects of substances that
are also carcinogens, it is essential to refer to other sources of information concerning the
carcinogenicity of this chemical substance. If the U.S. EPA has evaluated this substance for
potential human carcinogenicity, a summary of that evaluation will be contained in Section II of
this file.
There was no previous RfD for trichloroethylene on the IRIS database.
I.A.1. CHRONIC ORAL RfD SUMMARY
Critical Effect Point of UF Chronic RfD**
Departure*
Multiple (see below) Multiple (see Multiple 0.0005 mg/kg/day
below) (see below)
Decreased thymus weight in female HEDyyjovii.: 100
B6C3F1 mice 0.048 mg/kg/day (Candidate
RfD =
30 week drinking water study 0.00048
mg/kg/day)
Keil et al. (2009)
Decreased PFC response (3 and 8
weeks), increased delayed-type
hypersensitivity in B6C3F1 mice
Drinking water exposure from GD0
to 3- or 8-weeks of age
Peden-Adams et al. (2006)
Increased fetal cardiac
malformations in Sprague-Dawley
rats
Drinking water exposure from GDI
to GD22
Johnson et al. (2003)
*Conversion Factors and Assumptions - For Keil et al. (2009). the HE Poo mui is the 99th
percentile (due to human toxicokinetic uncertainty and variability) human equivalent dose to the
mouse LOAEL of 0.35 mg/kg/day, using the internal dose metric of TCE metabolized/kg3/Vday.
2
LOAEL: 1000
0.37 mg/kg/day (Candidate
RfD =
0.00037
mg/kg/day)
HED993MDL: 10
0.0051 mg/kg/day (Candidate
RfD =
0.00051
mg/kg/day)
-------
For Peden-Adams et al. (2006). there were no conversion factors. For Johnson et al. (2003). the
HED99 ij\[|)i is the 99th percentile (due to human toxicokinetic uncertainty and variability) human
equivalent dose to the rat internal dose BMDLoi of 0.0142 mg TCE oxidized/kg '/day. Details of
the methods used are presented in Section 5.1.3 of the ToxicolozicalReview of Trichloroethylene
(U.S. EPA. 2011)
** As a whole, the estimates support a RfD of 0.0005 mg/kg/d. This estimate is within 20% of
the estimates for the critical effects—0.0004 mg/kg/d for developmental immunotoxicity
(decreased PFC and increased delayed-type hypersensitivity) in mice, and 0.0005 mg/kg/d both
for heart malformations in rats and for decreased thymus weights in mice.
I.A.2. PRINCIPAL AND SUPPORTING STUDIES
The Toxicological Review of Trichloroethylene (TCE) reviews and summarizes the
available data on non-cancer effects caused by TCE (for summary, see U.S. EPA, 2011, Section
4.11.1). Adverse non-cancer effects associated with oral TCE exposure include decreased body
weight, liver and kidney effects, and neurological, immunological, reproductive, and
developmental effects. Candidate RfD values were developed for all endpoints on the basis of
applied dose (U.S. EPA, 2011, Section 5.1.2) and for the more sensitive endpoints on the basis of
PBPK model-derived internal dose (U.S. EPA, 2011, Section 5.1.3). The most sensitive
observed adverse effects, which were used as the principal bases of the RfD, were those affecting
the immune system and the developing fetus. Additional support for the RfD was based on
adverse effects in the kidney.
In particular, multiple candidate RfDs for the principal and supporting effects from oral
studies are in the relatively narrow range of 0.0003-0.0008 mg/kg/d, at the low end of the
overall range of candidate RfDs for all adverse effects. Given the somewhat imprecise nature of
the individual candidate RfD values, and the fact that multiple effects/studies lead to similar
candidate RfD values, the approach taken in this assessment is to select a RfD supported by
multiple effects/studies. The advantages of this approach, which is only possible when there is a
relatively large database of studies/effects and when multiple candidate values happen to fall
within a narrow range at the low end of the overall range, are that it leads to a more robust RfD
(less sensitive to limitations of individual studies) and that it provides the important
characterization that the RfD exposure level is similar for multiple noncancer effects rather than
being based on a sole explicit critical effect.
Three principal (Johnson et al.. 2003; Keil et al.. 2009; Peden-Adams et al.. 2006) and
two supporting (NTP, 1988; Woolhiser et al., 2006) studies/effects have been chosen as the basis
of the RfD for TCE noncancer effects (see Table below). Two of the lowest candidate RfDs for
the primary dose metrics—0.0008 mg/kg/d for increased kidney weight in rats and 0.0005
mg/kg/d for both heart malformations in rats and decreased thymus weights in mice—are derived
using the PBPK model for inter- and intraspecies extrapolation, and a third—0.0003 mg/kg/d for
increased toxic nephropathy in rats—is derived using the PBPK model for inter- and intraspecies
extrapolation as well as route-to-route extrapolation from an inhalation study. The other of these
lowest values—0.0004 mg/kg/d for developmental immunotoxicity (decreased PFC response and
increased delayed-type hypersensitivity) in mice—is based on applied dose.
There is medium confidence in the candidate RfDs for decreased thymus weights (U.S.
EPA, 2011, Section 5.1.2.5) and heart malformations (U.S. EPA, 2011, Section 5.1.2.8) and
developmental immunological effects (U.S. EPA, 2011, Section 5.1.2.8), and these effects are
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considered the critical effects used for deriving the RfD. For developmental effects, although the
available study has important limitations, the overall weight of evidence supports an effect of
TCE on cardiac development. For adult and developmental immunological effects, there is high
confidence in the evidence for an immunotoxic hazard from TCE. However, the available dose-
response data for immunological effects preclude application of BMD modeling.
For kidney effects (U.S. EPA, 2011, Section 5.1.2.2), there is high confidence in the
evidence for a nephrotoxic hazard from TCE. Moreover, the two lowest candidate RfDs for
kidney effects (toxic nephropathy and increased kidney weight) are both based on BMD
modeling and one is derived from a chronic study. However, as discussed in U.S. EPA (2011,
Section 3.3.3.2), there remains substantial uncertainty in the extrapolation of GSH conjugation
from rodents to humans due to limitations in the available data. In addition, the candidate RfD
value for toxic nephropathy had greater dose-response uncertainty since the estimation of its
POD involved extrapolation from high response rates (>60%). Therefore, kidney effects are
considered supportive but are not used as a primary basis for the RfD.
As a whole, the estimates support a RfD of 0.0005 mg/kg/d. This estimate is within 20%
of the estimates for the critical effects—0.0004 mg/kg/d for developmental immunotoxicity
(decreased PFC and increased delayed-type hypersensitivity) in mice, and 0.0005 mg/kg/d both
for heart malformations in rats and for decreased thymus weights in mice. This estimate is also
within approximately a factor of two of the supporting effect estimates of 0.0003 mg/kg/d for
toxic nephropathy in rats and 0.0008 mg/kg/d for increased kidney weight in rats. Thus, there is
strong, robust support for a RfD of 0.0005 mg/kg/d provided by the concordance of estimates
derived from multiple effects from multiple studies. The estimates for kidney effects, thymus
effects, and developmental heart malformations are based on PBPK model-based estimates of
internal dose for interspecies and intraspecies extrapolation, and there is sufficient confidence in
the PBPK model and support from mechanistic data for one of the dose metrics (total oxidative
metabolism for the heart malformations). There is high confidence that the amount of
bioactivated DCVC would be an appropriate dose metric to use for kidney effects, but there is
substantial quantitative uncertainty in the PBPK model predictions for this dose metric in
humans (U.S. EPA, 2011, Section 5.1.3.1). Note that there is some human evidence of
developmental heart defects from TCE exposure in community studies (U.S. EPA, 2011, Section
4.8.3.1.1) and of kidney toxicity in TCE-exposed workers (U.S. EPA, 2011, Section 4.4.1).
In summary, the RfD is 0.0005 mg/kg/d based on the critical effects of heart
malformations (rats), adult immunological effects (mice), and developmental immunotoxicity
(mice), all from oral studies. This RfD value is further supported by results from an oral study
for the effect of toxic nephropathy (rats) and route-to-route extrapolated results from an
inhalation study for the effect of increased kidney weight (rats).
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Summary of critical studies, effects, PODs, and UFs used to derive the RfD
Keil et al. (2009)—Decreased thymus weight in female B6C3F1 mice exposed for 30 weeks by drinking
water.
• Internal dose POD = 0.139 mg TCE metabolized/kg Yd. which is the PBPK model-predicted
internal dose at the applied dose LOAEL of 0.35 mg/kg/d (continuous) (no BMD modeling due
to inadequate model fit caused by supralinear dose-response shape) (U.S. EPA, 2011, Appendix
F, Section F.6.4).
• HED99 = 0.048 mg/kg/d (lifetime continuous exposure) derived from combined interspecies and
intraspecies extrapolation using PBPK model.
• UF = 100.
• Primary candidate RfD = HED99/UF = 0.048/100 = 0.00048 mg/kg/d.
Peden-Adams et al. (2006)—Decreased PFC response (3 and 8 weeks), increased delayed-type
hypersensitivity (8 weeks) in pups exposed from GD 0 to 3- or 8-weeks-of-age through drinking water
(placental and lactational transfer, and pup ingestion).
• POD = 0.37 mg/kg/d is the applied dose LOAEL (estimated daily dam dose) (no BMD modeling
due to inadequate model fit caused by supralinear dose-response shape). No PBPK modeling was
attempted due to lack of appropriate models/parameters to account for complicated fetal/pup
exposure pattern (U.S. EPA, 2011, Appendix F, Section F.6.6).
• UF = 1000.
• Primary candidate RfD = HED99/UF = 0.37/1000 = 0.00037 mg/kg/d.
Johnson et al. (2003)—fetal heart malformations in S-D rats exposed from GD 1-22 by drinking water
• Internal dose POD = 0.0142 mg TCE metabolized by oxidation/kg Yd. which is the BMDL from
BMD modeling using PBPK model-predicted internal doses, with highest-dose group (1,000-fold
higher than next highest-dose group) dropped, pup as unit of analysis, BMR =1% (due to
severity of defects, some of which could have been fatal), and a nested Log-logistic model to
account for intralitter correlation (U.S. EPA, 2011, Appendix F, Section F.6.5).
• HED99 = 0.0051 mg/kg/d (lifetime continuous exposure) derived from combined interspecies and
intraspecies extrapolation using PBPK model.
• UF = 10
• Primary candidate RfD = HED99/UF = 0.0051/10 = 0.00051 mg/kg/d.
GD = gestation day.
Summary of supporting studies, effects, PODs, and UFs for the RfD
NTP (1988)—Toxic nephropathy in female Marshall rats exposed for 104 weeks by oral gavage (5 d/wk).
• Internal dose POD = 0.0132 mg DCVC bioactivatcd/kg Yd. which is the BMDL from BMD
modeling using PBPK model-predicted internal doses, BMR = 5% (clearly toxic effect), and Log-
logistic model (U.S. EPA, 2011, Appendix F, Section F.6.1).
• HED99 = 0.0034 mg/kg/d (lifetime continuous exposure) derived from combined interspecies and
intraspecies extrapolation using PBPK model.
• UF = 10.
• Supporting candidate RfD = HED99/UF = 0.0034/10 = 0.00034 mg/kg/d.
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Woolhiser et al. (2006)—Increased kidney weight in female S-D rats exposed for 4 weeks by inhalation
(6 h/d, 5 d/wk).
• Internal dose POD = 0.0309 mg DCVC bioactivated/kg3/4/d, which is the BMDL from BMD
modeling using PBPK model-predicted internal doses, BMR = 10%, and Hill model with
constant variance (U.S. EPA, 2011, Appendix F, Section F.6.3).
• HED99 = 0.0079 mg/kg/d (lifetime continuous exposure) derived from combined interspecies and
intraspecies extrapolation using PBPK model.
• UF = 10.
• Supporting candidate RfD = HED99/UF = 0.0079/10 = 0.00079 mg/kg/d.
I.A.3. UNCERTAINTY FACTORS
Uncertainty factors are used to address differences between study conditions and conditions of
human environmental exposure (U.S. EPA, 2002). These include
(a) Extrapolating from laboratory animals to humans: If a POD is derived from
experimental animal data, it is divided by an UF to reflect pharmacokinetic and
pharmacodynamic differences that may make humans more sensitive than laboratory
animals. For oral exposures, the standard value for the interspecies UF is 10, which
breaks down (approximately) to a factor of three for pharmacokinetic differences (which
is removed if the PBPK model is used) and a factor of three for pharmacodynamic
differences. For inhalation exposures, ppm equivalence across species is generally
assumed, in which case pharmacokinetic differences are considered to be negligible, and
the standard value used for the interspecies UF is 3, which is ascribed to
pharmacodynamic differences. These standard values were used for all the candidate
RfCs and RfDs based on laboratory animal data in this assessment.
(b) Human (intraspecies) variability: RfCs and RfDs apply to the human population,
including sensitive subgroups, but studies rarely examine sensitive humans. Sensitive
humans could be adversely affected at lower exposures than a general study population;
consequently, PODs from general-population studies are divided by an UF to address
sensitive humans. Similarly, the animals used in most laboratory animal studies are
considered to be "typical" or "average" responders, and the human (intraspecies)
variability UF is also applied to PODs from such studies to address sensitive subgroups.
The standard value for the human variability UF is 10, which breaks down
(approximately) to a factor of three for pharmacokinetic variability (which is removed if
the PBPK model is used) and a factor of three for pharmacodynamic variability. This
standard value was used for all the PODs in this assessment with the exception of the
PODs for a few immunological effects that were based on data from a sensitive
(autoimmune-prone) mouse strain; for those PODs, an UF of 3 was used for human
variability.
(c) Uncertainty in extrapolating from subchronic to chronic exposures: RfCs and RfDs
apply to lifetime exposure, but sometimes the best (or only) available data come from
less-than-lifetime studies. Lifetime exposure can induce effects that may not be apparent
or as large in magnitude in a shorter study; consequently, a dose that elicits a specific
level of response from a lifetime exposure may be less than the dose eliciting the same
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level of response from a shorter exposure period. Thus, PODs based on subchronic
exposure data are generally divided by a sub chronic-to-chronic UF, which has a standard
value of 10. If there is evidence suggesting that exposure for longer time periods does
not increase the magnitude of an effect, a lower value of three or one might be used. For
some reproductive and developmental effects, chronic exposure is that which covers a
specific window of exposure that is relevant for eliciting the effect, and subchronic
exposure would correspond to an exposure that is notably less than the full window of
exposure.
(d) Uncertainty in extrapolating from LOAELs to NOAELs: PODs are intended to be
estimates of exposure levels without appreciable risk under the study conditions so that,
after the application of appropriate UFs for interspecies extrapolation, human variability,
and/or duration extrapolation, the absence of appreciable risk is conveyed to the RfC or
RfD exposure level to address sensitive humans with lifetime exposure. Under the
NOAEL/LOAEL approach to determining a POD, however, adverse effects are
sometimes observed at all study doses. If the POD is a LOAEL, it is divided by an UF to
better estimate a NOAEL. The standard value for the LOAEL-to-NOAEL UF is 10,
although sometimes a value of three is used if the effect is considered minimally adverse
at the response level observed at the LOAEL or even one if the effect is an early marker
for an adverse effect. For one POD in this assessment, a value of 30 was used for the
LOAEL-to-NOAEL UF because the incidence rate for the adverse effect was >90% at the
LOAEL.
(e) Additional database uncertainties: A database UF of 1, 3 or 10 is used to reflect the
potential for deriving an underprotective toxicity value as a result of an incomplete
characterization of the chemical's toxicity. No database UF was used in this assessment.
See U.S. EPA (2011, Section 5.1.4.1) for additional discussion of the uncertainties
associated with the overall database for TCE.
Specific UFs used in the principal and supporting studies for the RfD are summarized in the
following tables. (Note that UF values of "3" actually represent VlO, and, when 2 such values are
multiplied together, the result is 10 rather than 9.)
Summary of critical studies, effects, and UFs used to derive the RfD
Keil et al. (2009)—Decreased thymus weight in female B6C3F1 mice exposed for 30 weeks by drinking
water.
• UFcomp0site 100.
• UFloaei =10 because POD is a LOAEL for an adverse effect.
• UF1S = 3 because the PBPK model was used for interspecies (is) extrapolation.
• UFh = 3 because the PBPK model was used to characterize human (h) toxicokinetic variability.
Peden-Adams et al. (2006)—Decreased PFC response (3 and 8 weeks), increased delayed-type
hypersensitivity (8 weeks) in pups exposed from GD 0 to 3- or 8-weeks-of-age through drinking water
(placental and lactational transfer, and pup ingestion).
• UFcomposite = 1000.
• UFloaei = 10 because POD is a LOAEL for multiple adverse effects.
• UF1S = 10 for interspecies extrapolation because PBPK model was not used.
• UFh = 10 for human variability because PBPK model was not used.
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Johnson et al. (2003)—fetal heart malformations in S-D rats exposed from GD 1-22 by drinking water
• UFcomp0site 10
• UF1S = 3 because the PBPK model was used for interspecies extrapolation.
• UFh = 3 because the PBPK model was used to characterize human toxicokinetic variability.
GD = gestation day.
Summary of supporting studies, effects, and UFs for the RfD
NTP (1988)—Toxic nephropathy in female Marshall rats exposed for 104 weeks by oral gavage (5 d/wk).
• UFCOmposite 10.
• UF1S = 3 because the PBPK model was used for interspecies extrapolation.
• UFh = 3 because the PBPK model was used to characterize human toxicokinetic variability.
Woolhiser et al. (2006)—Increased kidney weight in female S-D rats exposed for 4 weeks by inhalation
(6 h/d, 5 d/wk).
• UFcomp0site 10 •
• UFSC = 1 because Kjellstrand et al. (1983) reported that in mice, kidney effects after exposure for
120 d was no more severe than those after 30 d exposure.
• UF1S = 3 because the PBPK model was used for interspecies extrapolation.
• UFh = 3 because the PBPK model was used to characterize human toxicokinetic variability.
I.A.4. ADDITIONAL STUDIES/COMMENTS
I.A.5. CONFIDENCE IN THE CHRONIC ORAL RfD
Study - High-medium/medium/low-medium (for each endpoint individually)
Data Base - High
RfD - High
For adult and developmental immunological effects, there is high confidence in the
evidence of immunotoxic hazard from TCE. However, the available dose-response data for the
most sensitive for immulological effects (Keil et al., 2009; Peden-Adams et al„ 2006) precluded
application of BMD modeling. There are inadequate data on the active moiety for TCE-induced
immulogical effects, so PBPK modeling applied to Kiel et al. (2009) used a generic dose metric.
The PBPK model could not be applied to Peden-Adams et al. (2006) due to a lack of data on
gestational and lactational transfer. Thus, due to the high confidence in the immunotoxic hazard
coupled with the quantitative uncertainties in the dose-response assessment, the confidence in
candidate RfDs derived from these studies is characterized as medium-to-high.
For developmental cardiac effects, although the available study (Johnson et al.. 2003) has
important limitations, the overall weight of evidence supports an effect of TCE on cardiac
development. Both BMD and PBPK modeling could be applied to there data. With respect to
PBPK modeling, data suggest that oxidative metabolites are involved in TCE-induced cardiac
malformations, lending greater confidence in the appropriateness of the selected dose metric.
Thus, due to the important limitations of the critical study coupled with the higher confidence in
the dose-response analysis, the confidence in the candidate RfD derived from this study is
characterized as medium.
For kidney effects, there is high confidence in the evidence of nephrotoxic hazard from
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TCE. Both BMD and PBPK modeling could be applied to the most sensitive studies for this
endpoint (NTP, 1988; Woolhiser et al., 2006), and one of these studies is of chronic duration
rNTP. 1988). However, although there is high confidence in the conclusion that GSH
conjugation metabolites are involved in TCE nephrotoxicity, there remains substantial
uncertainty in the extrapolation of GSH conjugation from rodents to humans due to limitations in
the available data. In addition, BMD modeling of the NTP (1988) data involved extrapolation
from response rates much higher than the chosen BMR. Therefore, due to the high qualitative
confidence coupled with the low quantitative confidence, the overall confidence in candidate
RfDs derived from these studies is characterized as low-to-medium.
The RfD is supported by three principal studies (whose candidate RfDs are characterized
as being of medium-to-high/medium confidence) and two supporting studies (whose candidate
RfDs are characterized as being of low-to-medium confidence). Morever, the multiple candidate
RfDs from these studies fall within a narrow range, providing robust support for the final RfD.
In addition, numerous studies were available for other potential candidate critical effects, which
were also considered. Thus, overall, confidence in both the database and the RfD is
characterized as high.
I.A.6. EPA DOCUMENTATION AND REVIEW OF THE CHRONIC ORAL RfD
Source Document — U.S. EPA (2011)
This document has been reviewed by EPA scientists, interagency reviewers from other
federal agencies and White House offices, and the public, and peer reviewed by independent
scientists external to EPA. A summary and EPA's disposition of the comments received from
the independent external peer reviewers and from the public is included in Appendix I of the
ToxicologicalReview of Trichloroethylene (U.S. EPA, 2011).
Agency Completion Date —
I.A.7. EPA CONTACTS
Please contact the IRIS Hotline for all questions concerning this assessment or IRIS, in
general, at (202) 566-1676 (phone), (202) 566-1749 (fax), or hotline.iris@epa.gov (email
address).
I.B. REFERENCE CONCENTRATION (RfC) FOR CHRONIC INHALATION
EXPOSURE
Substance Name - Trichloroethylene
CASRN-79-01-6
Section I.B. Last Revised — 00/00/0000
The RfC is an estimate (with uncertainty spanning perhaps an order of magnitude) of a
continuous inhalation exposure to the human population (including sensitive subgroups) that is
likely to be without an appreciable risk of deleterious effects during a lifetime. The RfC
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considers both toxic effects of the respiratory system (portal-of-entry) and effects peripheral to
the respiratory system (extrarespiratory effects). The inhalation RfC (generally expressed in
"3
units of mg/m ) is analogous to the oral RfD and is similarly intended for use in risk assessments
for health effects known or assumed to be produced through a nonlinear (presumed threshold)
mode of action.
Inhalation RfCs are derived according to Methods for Derivation of Inhalation Reference
Concentrations and Application of Inhalation Dosimetry (U.S. EPA, 1994). Because RfCs can
also be derived for the noncarcinogenic health effects of substances that are carcinogens, it is
essential to refer to other sources of information concerning the carcinogenicity of this chemical
substance. If the U.S. EPA has evaluated this substance for potential human carcinogenicity, a
summary of that evaluation will be contained in Section II of this file.
There was no previous RfC for trichloroethylene on the IRIS database.
I.B.I. CHRONIC INHALATION RfC SUMMARY
Critical Effect Point of UF Chronic RfC**
Departure*
Multiple (see below)
Multiple (see
Multiple
0.002 mg/m3
below)
(see below)
(0.0004 ppm)
Decreased thymus weight in female
HECg^LOAEL:
100
B6C3F1 mice
0.19 mg/m3
(Candidate
(0.033 ppm)
RfC =
30 week drinking water study
0.0019
mg/m3
Keil et al. (2009)
[0.00033
ppm])
Increased fetal cardiac
HEC993MDL:
10
malformations in Sprague-Dawley
0.021 mg/m3
(Candidate
rats
(0.0037 ppm)
RfC =
0.0021
Drinking water exposure from GDI
mg/m3
to GD22
[0.00037
ppm])
Johnson et al. (2003)
*Conversion Factors and Assumptions - For Keil et al. (2009), the HECgg^oAEL is the route-to-
route extrapolated 99th percentile (due to human toxicokinetic uncertainty and variability) human
equivalent concentration to the mouse LOAEL of 0.35 mg/kg/day, using the internal dose metric
of TCE metabolized/kg'Vday. For Johnson et al. (2003), the HEC99,BMDLis the route-to-route
extrapolated 99th percentile (due to human toxicokinetic uncertainty and variability) human
equivalent concentration to the rat internal dose BMDLoi of 0.0142 mg TCE oxidized/kg '/day.
Details of the methods used are presented in Section 5.1.3 of the Toxicological Review of
Trichloroethylene (U.S. EPA, 2011)
** As a whole, the estimates support a RfC of 0.0004 ppm (0.4 ppb or 2 (j,g/m3). This estimate
essentially reflects the midpoint between the similar candidate RfC estimates for the two critical
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effects (0.00033 ppm for decreased thymus weight in mice and 0.00037 ppm for heart
malformations in rats), rounded to one significant figure.
I.B.2. PRINCIPAL AND SUPPORTING STUDIES
The Toxicological Review of Trichloroethylene (TCE) reviews and summarizes the
available data on non-cancer effects caused by TCE (for summary, see U.S. EPA, 2011, Section
4.11.1). Adverse non-cancer effects associated with TCE exposure by inhalation include hepatic,
renal, neurological, immunological, reproductive, and developmental effects. Candidate RfC
values were developed for all endpoints on the basis of applied dose (U.S. EPA, 2011, Section
5.1.2) and for the more sensitive endpoints on the basis of PBPK model-derived internal dose
(U.S. EPA, 2011, Section 5.1.3). The most sensitive observed adverse effects, which were used
as the principal bases of the RfC, were those affecting the immune system and the developing
fetus. Additional support for the RfC was based on adverse effects in the kidney.
In particular, multiple candidate RfCs for the principal and supporting effects are in the
relatively narrow range of 0.0003-0.0006 ppm, at the low end of the overall range of candidate
RfCs for all adverse effects. Given the somewhat imprecise nature of the individual candidate
RfC values, and the fact that multiple effects/studies lead to similar candidate RfC values, the
approach taken in this assessment is to select a RfC supported by multiple effects/studies. The
advantages of this approach, which is only possible when there is a relatively large database of
studies/effects and when multiple candidate values happen to fall within a narrow range at the
low end of the overall range, are that it leads to a more robust RfC (less sensitive to limitations of
individual studies) and that it provides the important characterization that the RfC exposure level
is similar for multiple noncancer effects rather than being based on a sole explicit critical effect.
Two principal (Johnson et al.. 2003; Keil et al.. 2009) and one supporting CNTP. 1988)
studies/effects have been chosen to as the basis of the RfC for TCE noncancer effects (see Table
below). Each of these lowest candidate RfCs, ranging from 0.0003-0.0006 ppm, for
developmental, immunologic, and kidney effects, are values derived from route-to-route
extrapolation using the PBPK model. The lowest candidate RfC estimate (for a primary dose
metric) from an inhalation studies is 0.001 ppm for kidney effects, which is higher than the
route-to-route extrapolated candidate RfC estimate from the most sensitive oral study. For each
of the candidate RfCs, the PBPK model was used for inter- and intraspecies extrapolation, based
on the preferred dose metric for each endpoint.
There is medium confidence in the lowest candidate RfC for developmental effects (heart
malformations) (U.S. EPA, 2011, Section 5.1.2.8) and the lowest candidate RfC estimate for
immunological effects (U.S. EPA, 2011, Section 5.1.2.5), and these are considered the critical
effects used for deriving the RfC. For developmental effects, although the available study has
important limitations, the overall weight of evidence supports an effect of TCE on cardiac
development. For immunological effects, there is high confidence in the evidence for an
immunotoxic hazard from TCE, but the available dose-response data preclude application of
BMD modeling.
For kidney effects (U.S. EPA, 2011, Section 5.1.2.2), there is high confidence in the
evidence for a nephrotoxic hazard from TCE. Moreover, the lowest candidate RfC for kidney
effects (toxic nephropathy) is derived from a chronic study and is based on BMD modeling.
However, as discussed in U.S. EPA (2011, Section 3.3.3.2), there remains substantial uncertainty
in the extrapolation of GSH conjugation from rodents to humans due to limitations in the
available data. In addition, the p-cRfC for toxic nephropathy had greater dose-response
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uncertainty since the estimation of its POD involved extrapolation from high response rates
(>60%). Therefore, toxic nephropathy is considered supportive but is not used as a primary basis
for the RfC. The other sensitive candidate RfCs for kidney effects were all within a factor of 5
of that for toxic nephropathy; however, these values similarly relied on the uncertain inter-
species extrapolation of GSH conjugation.
As a whole, the estimates support a RfC of 0.0004 ppm (0.4 ppb or 2 (J,g/m3). This
estimate essentially reflects the midpoint between the similar candidate RfC estimates for the
two critical effects (0.00033 ppm for decreased thymus weight in mice and 0.00037 ppm for
heart malformations in rats), rounded to one significant figure. This estimate is also within a
factor of two of the candidate RfC estimate of 0.00006 ppm for the supporting effect of toxic
nephropathy in rats. Thus, there is robust support for a RfC of 0.0004 ppm provided by
estimates for multiple effects from multiple studies. The estimates are based on PBPK model-
based estimates of internal dose for interspecies, intraspecies, and route-to-route extrapolation,
and there is sufficient confidence in the PBPK model and support from mechanistic data for one
of the dose metrics (TotOxMetabBW34 for the heart malformations). There is high confidence
that ABioactDCVCBW34 and AMetGSHBW34 would be appropriate dose metrics for kidney
effects, but there is substantial uncertainty in the PBPK model predictions for these dose metrics
in humans (U.S. EPA, 2011, Section 5.1.3.1). Note that there is some human evidence of
developmental heart defects from TCE exposure in community studies (U.S. EPA, 2011, Section
4.8.3.1.1) and of kidney toxicity in TCE-exposed workers (U.S. EPA, 2011, Section 4.4.1).
"3
In summary, the RfC is 0.0004 ppm (0.4 ppb or 2 (J,g/m ) based on route-to-route
extrapolated results from oral studies for the critical effects of heart malformations (rats) and
immunotoxicity (mice). This RfC value is further supported by route-to-route extrapolated
results from an oral study of toxic nephropathy (rats).
Summary of critical studies, effects, PODs, and UFs used to derive the RfC
Keil et al. (2009)—Decreased thymus weight in female B6C3F1 mice exposed for 30 weeks by drinking
water.
• Internal dose POD = 0.139 mg TCE metabolized/kg Vd. which is the PBPK model-predicted
internal dose at the applied dose LOAEL of 0.35 mg/kg/d (continuous) (no BMD modeling due
to inadequate model fit caused by supralinear dose-response shape) (U.S. EPA, 2011, Appendix
F, Section F.6.4).
• HEC99 = 0.033 ppm (lifetime continuous exposure) derived from combined interspecies,
intraspecies, and route-to-route extrapolation using PBPK model.
• UF = 100.
• Principal candidate RfC = HEC99/UF = 0.033/100 = 0.00033 ppm (2 (.ig/nr1).
lohnson et al. (2003)—fetal heart malformations in S-D rats exposed from GD 1-22 by drinking water.
• Internal dose POD = 0.0142 mg TCE metabolized by oxidation/kg Yd. which is the BMDL from
BMD modeling using PBPK model-predicted internal doses, with highest-dose group (1,000-fold
higher than next highest-dose group) dropped, pup as unit of analysis, BMR =1% (due to
severity of defects, some of which could have been fatal), and a nested Log-logistic model to
account for intralitter correlation (U.S. EPA, 2011, Appendix F, Section F.6.5).
• HEC99 = 0.0037 ppm (lifetime continuous exposure) derived from combined interspecies,
intraspecies, and route-to-route extrapolation using PBPK model.
• UF = 10.
• Principal candidate RfC = HEC99/UF = 0.0037/10 = 0.00037 ppm (2 (.ig/nr1).
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GD = gestation day.
Summary of supporting study, effect, POD, and UFs for the RfC
NTP (1988)—Toxic nephropathy in female Marshall rats exposed for 104 weeks by oral gavage (5 d/wk).
• Internal dose POD = 0.0132 mg DCVC bioactivatcd/kg Yd. which is the BMDL from BMD
modeling using PBPK model-predicted internal doses, BMR = 5% (clearly toxic effect), and log-
logistic model (U.S. EPA, 2011, Appendix F, Section F.6.1).
• HEC99 = 0.0056 ppm (lifetime continuous exposure) derived from combined interspecies,
intraspecies, and route-to-route extrapolation using PBPK model.
• UF = 10.
• Supporting candidate RfC = HEC99/UF = 0.0056/10 = 0.00056 ppm (3 (.ig/nr1).
I.B.3. UNCERTAINTY FACTORS
General discussion of uncertainty factors is presented above in I.A.3. Specific UFs used in the
principal and supporting studies for the RfC are summarized in the following tables. (Note that
UF values of "3" actually represent VlO, and, when 2 such values are multiplied together, the
result is 10 rather than 9.)
Summary of critical studies, effects, and UFs used to derive the RfC
Keil et al. (2009)—Decreased thymus weight in female B6C3F1 mice exposed for 30 weeks by drinking
water.
• UFCOmposite 100.
• UFioaei =10 because POD is a LOAEL for an adverse effect.
• UF1S = 3 because the PBPK model was used for interspecies extrapolation.
• UFh = 3 because the PBPK model was used to characterize human toxicokinetic variability.
Johnson et al. (2003)—fetal heart malformations in S-D rats exposed from GD 1-22 by drinking water.
• UFCOmposite 10 •
• UF1S = 3 because the PBPK model was used for interspecies extrapolation.
• UFh = 3 because the PBPK model was used to characterize human toxicokinetic variability.
GD = gestation day.
Summary of supporting study, effect, and UFs for the RfC
NTP (1988)—Toxic nephropathy in female Marshall rats exposed for 104 weeks by oral gavage (5 d/wk).
• UFcomp0site 10 •
• UF1S = 3 because the PBPK model was used for interspecies extrapolation.
• UFh = 3 because the PBPK model was used to characterize human toxicokinetic variability.
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I.B.4. ADDITIONAL STUDIES/COMMENTS
I.B.5. CONFIDENCE IN THE CHRONIC INHALATION RfC
Study - High-medium/medium/low-medium (for each endpoint individually)
Data Base - High
RfC -- High
For adult immunological effects, there is high confidence in the evidence of immunotoxic
hazard from TCE. However, the available dose-response data for the most sensitive for
immulological effects (Keil et al.. 2009) precluded application of BMD modeling. There are
inadequate data on the active moiety for TCE-induced immulogical effects, so PBPK modeling
applied to Kiel et al. (2009) used a generic dose metric. Thus, due to the high confidence in the
immunotoxic hazard coupled with the quantitative uncertainties in the dose-response assessment,
the confidence in the candidate RfC derived from this study is characterized as medium-to-high.
For developmental cardiac effects, although the available study (Johnson et al., 2003) has
important limitations, the overall weight of evidence supports an effect of TCE on cardiac
development. Both BMD and PBPK modeling could be applied to there data. With respect to
PBPK modeling, data suggest that oxidative metabolites are involved in TCE-induced cardiac
malformations, lending greater confidence in the appropriateness of the selected dose metric.
Thus, due to the important limitations of the critical study coupled with the higher confidence in
the dose-response analysis, the confidence in the candidate RfC derived from this studies is
characterized as medium.
For kidney effects, there is high confidence in the evidence of nephrotoxic hazard from
TCE. Both BMD and PBPK modeling could be applied to the most sensitive study for this
endpoint (NTP, 1988), which is of chronic duration. However, although there is high confidence
in the conclusion that GSH conjugation metabolites are involved in TCE nephrotoxicity, there
remains substantial uncertainty in the extrapolation of GSH conjugation from rodents to humans
due to limitations in the available data. In addition, BMD modeling of the NTP (1988) data
involved extrapolation from response rates much higher than the chosen BMR. Therefore, due to
the high qualitative confidence coupled with the low quantitative confidence, the overall
confidence in the candidate RfCs derived from these studies is characterized as low-to-medium.
The RfC is supported by two principal studies (whose candidate RfCs are characterized
as being of medium-to-high/medium confidence) and one supporting study (whose candidate
RfC is characterized as being of low-to-medium confidence). Morever, the multiple candidate
RfCs from these studies fall within a narrow range, providing robust support for the final RfC.
In addition, numerous studies were available for other potential candidate critical effects, which
were also considered. Thus, overall, confidence in both the database and the RfC is
characterized as high.
I.B.6. EPA DOCUMENTATION AND REVIEW OF THE CHRONIC INHALATION
RfC
Source Document - U.S. EPA (2011)
This document has been reviewed by EPA scientists, interagency reviewers from other
federal agencies and White House offices, and the public, and peer reviewed by independent
14
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scientists external to EPA. A summary and EPA's disposition of the comments received from
the independent external peer reviewers and from the public is included in Appendix I of the
ToxicologicalReview of Trichloroethylene (U.S. EPA. 2011).
Agency Completion Date —
I.B.7. EPA CONTACTS
Please contact the IRIS Hotline for all questions concerning this assessment or IRIS, in
general, at (202) 566-1676 (phone), (202) 566-1749 (fax), or hotline.iris@epa.gov (email
address).
II. CARCINOGENICITY ASSESSMENT FOR LIFETIME EXPOSURE
Substance Name - Trichloroethylene
CASRN-79-01-6
Section II. Last Revised — 00/00/0000
This section provides information on three aspects of the carcinogenic assessment for the
substance in question: the weight-of-evidence judgment of the likelihood that the substance is a
human carcinogen, and quantitative estimates of risk from oral and inhalation exposure. Users
are referred to Section I of this file for information on long-term toxic effects other than
carcinogenicity.
The rationale and methods used to develop the carcinogenicity information in IRIS are
described in the Guidelines for Carcinogen Risk Assessment and the Supplemental Guidance for
Assessing Susceptibility from Early-Life Exposure to Carcinogens (U.S. EPA. 2005b) (U.S.
EPA, 2005c). The quantitative risk estimates are derived from the application of a low-dose
extrapolation procedure, and are presented in two ways to better facilitate their use. First, route-
specific risk values are presented. The "oral slope factor" is a plausible upper bound on the
estimate of risk per mg/kg-day of oral exposure. Similarly, a "unit risk" is a plausible upper
bound on the estimate of risk per unit of concentration, per (j,g/m3 air breathed (see Section
II.C.l).
A previous cancer assessment for trichloroethylene is not available on the IRIS database.
II.A. EVIDENCE FOR HUMAN CARCINOGENICITY
II.A.1. WEIGHT-OF-EVIDENCE CHARACTERIZATION
Following U.S. EPA (2005b) Guidelines for Carcinogen Risk Assessment, TCE is
characterized as "Carcinogenic to Humans" by all routes of exposure. This conclusion is based
on convincing evidence of a causal association between TCE exposure in humans and kidney
cancer. The kidney cancer association cannot be reasonably attributed to chance, bias, or
confounding. The human evidence of carcinogenicity from epidemiologic studies of TCE
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exposure is strong for non-Hodgkin lymphoma (NHL), but less convincing than for kidney
cancer, and more limited for liver and biliary tract cancer. In addition to the body of evidence
pertaining to kidney cancer, NHL, and liver cancer, the available epidemiologic studies also
provide more limited evidence of an association between TCE exposure and other types of
cancer, including bladder, esophageal, prostate, cervical, breast, and childhood leukemia.
Differences between these sets of data and the data for kidney cancer, NHL, and liver cancer are
observations from fewer numbers of studies, a mixed pattern of observed risk estimates, and the
general absence of exposure-response data from the studies using a quantitative TCE-specific
exposure measure.
There are several lines of supporting evidence for TCE carcinogenicity in humans. First,
TCE induces site-specific tumors in rodents given TCE by oral gavage and inhalation. Second,
toxicokinetic data indicate that TCE absorption, distribution, metabolism, and excretion are
qualitatively similar in humans and rodents. Finally, there is sufficient weight of evidence to
conclude that a mutagenic MOA is operative for TCE-induced kidney tumors, and this MOA is
clearly relevant to humans. MO As have not been established for other TCE-induced tumors in
rodents, and no mechanistic data indicate that any hypothesized key events are biologically
precluded in humans.
II.A.2. HUMAN CARCINOGENICITY DATA
The available epidemiologic studies provide convincing evidence of a causal association
between TCE exposure and cancer. The strongest epidemiologic evidence consists of reported
increased risks of kidney cancer, with more limited evidence for NHL and liver cancer, in
several well-designed cohort and case-control studies (discussed below). The summary
evaluation below of the evidence for causality is based on guidelines adapted from Hill (1965)
by U.S. EPA (2005b), and focuses on evidence related to kidney cancer, NHL, and liver cancer.
(a) Consistency of observed association. Elevated risks for kidney cancer have been observed
across many independent studies. Eighteen studies in which there is a high likelihood of TCE
exposure in individual study subjects (e.g., based on job-exposure matrices or biomarker
monitoring) and which were judged to have met, to a sufficient degree, the standards of
epidemiologic design and analysis, were identified in a systematic review of the epidemiologic
literature. Of the 15 of these studies reporting risks of kidney cancer (Anttila et al.. 1995;
Axelson et al., 1994; Boice et al., 1999; Briming et al., 2003; Charbotel et al., 2006; Dosemeci et
al„ 1999; Greenland etal., 1994; Hansen et al„ 2001; Moore et al., 2010; Morgan et al., 1998;
Pesch et al„ 2000; Raaschou-Nielsen et al., 2003; Radican et al„ 2008; Siemiatycki, 1991; Zhao
et al., 2005), most estimated relative risks between 1.1 and 1.9 for overall exposure to TCE (U.S.
EPA, 2011, Sections 4.1 and 4.4.2). Six of these 15 studies reported statistically significant
increased risks either for overall exposure to TCE (Briming et al., 2003; Dosemeci et al„ 1999;
Moore et al„ 2010; Raaschou-Nielsen et al„ 2003) or for one of the highest TCE exposure group
(Charbotel et al„ 2006; Moore et al„ 2010; Raaschou-Nielsen et al., 2003; Zhao et al„ 2005).
Thirteen other cohort, case-control, and geographic based studies were given less weight because
of their lesser likelihood of TCE exposure and other study design limitations that would decrease
statistical power and study sensitivity (U.S. EPA, Sections 4.1. and 4.4.2).
The consistency of association between TCE exposure and kidney cancer is further
supported by the results of the meta-analyses of the 15 cohort and case-control studies of
sufficient quality and with high probability TCE exposure potential to individual subjects. These
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analyses observed a statistically significant increased summary relative risk estimate (RRm) for
kidney cancer of 1.27 (95% CI: 1.13, 1.43) for overall TCE. The summary relative risk were
robust and did not change appreciably with the removal of any individual study or with the use
of alternate relative risk estimates from individual studies. In addition, there was no evidence for
heterogeneity or publication bias.
The consistency of increased kidney cancer relative risk estimates across a large number
of independent studies of different designs and populations from different countries and
industries argues against chance, bias or confounding as the basis for observed associations.
This consistency, thus, provides substantial support for a causal effect between kidney cancer
and TCE exposure.
Some evidence of consistency is found between TCE exposure and NHL and liver
cancer. In a weight-of-evidence review of the NHL studies, 17 studies in which there is a high
likelihood of TCE exposure in individual study subjects (e.g., based on job-exposure matrices or
biomarker monitoring) and which met, to a sufficient degree, the standards of epidemiologic
design and analysis were identified. These studies generally reported excess relative risk
estimates for NHL between 0.8 and 3.1 for overall TCE exposure (U.S. EPA, 2011, Section 4.1
and 4.6.1.2). Statistically significant elevated relative risk estimates for overall exposure were
observed in two cohort (Hansen et al., 2001; Raaschou-Nielsen et al., 2003) and one case-control
(Hardell et al.. 1994) studies. The other 14 identified studies reported elevated relative risk
estimates with overall TCE exposure that were not statistically significant (Anttila et al., 1995;
Axelson et al.. 1994; Boice et al.. 1999; Cocco et al.. 2010; Greenland et al.. 1994; Miligi et al..
2006; Morgan et al„ 1998; Nordstrom et al„ 1998; Persson and Fredrikson, 1999; Purdue et al.,
2011; Radican et al., 2008; Siemiatvcki, 1991; Wang et al., 2009; Zhao et al., 2005). Fifteen
additional studies were given less weight because of their lesser likelihood of TCE exposure and
other design limitations that would decrease study power and sensitivity (U.S. EPA, 2011,
Sections 4.1 and 4.6.1.2). The observed lack of association with NHL in these studies likely
reflects study design and exposure assessment limitations and is not considered inconsistent with
the overall evidence on TCE and NHL.
Consistency of the association between TCE exposure and NHL is further supported by
the results of meta-analyses. These meta-analyses found a statistically significant increased
summary relative risk estimate for NHL of 1.23 (95% CI: 1.07, 1.42) for overall TCE exposure.
This result and its statistical significance were not overly influenced by most individual studies.
Some heterogeneity was observed across the 17 studies of overall exposure, though it was not
statistically significant (p = 0.16). Analyzing the cohort and case-control studies separately
resolved most of the heterogeneity, but the result for the summary case-control studies was only
about a 7% increased relative risk estimate and was not statistically significant. The sources of
heterogeneity are uncertain but may be the result of some bias associated with exposure
assessment and/or disease classification, or from differences between cohort and case-control
studies in average TCE exposure. In addition, there is some evidence of potential publication
bias in this data set; however, it is uncertain that this is actually publication bias rather than an
association between standard error and effect size resulting for some other reason, e.g., a
difference in study populations or protocols in the smaller studies. Furthermore, if there is
publication bias in this data set, it does not appear to account completely for the finding of an
increased NHL risk.
There are fewer studies on liver cancer than for kidney cancer and NHL. Of nine studies,
all of them cohort studies, in which there is a high likelihood of TCE exposure in individual
study subjects (e.g., based on job-exposure matrices or biomarker monitoring) and which met, to
17
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a sufficient degree, the standards of epidemiologic design and analysis in a systematic review
(Anttila et al., 1995; Axelson et al., 1994; Boice et al., 1999; Boice et al., 2006; Greenland et al.,
1994; Morgan et al.. 1998; Radican et al.. 2008) (Hansen et al.. 2001; Raaschou-Nielsen et al..
2003), most reported relative risk estimates for liver and gallbladder cancer between 0.5 and 2.0
for overall exposure to TCE (U.S. EPA, 2011, Sections 4.1 and 4.5.2). Relative risk estimates
were generally based on small numbers of cases or deaths, with the result of wide confidence
intervals on the estimates, except for one study (Raaschou-Nielsen et al.. 2003). This study has
almost 6 times more cancer cases than the next largest study and observed a statistically
significant elevated liver and gallbladder cancer risk with overall TCE exposure (relative risk
[RR] = 1 .35 [95% CI: 1.03, 1.77]). Ten additional studies were given less weight because of
their lesser likelihood of TCE exposure and other design limitations that would decrease
statistical power and study sensitivity (U.S. EPA, 2011, Sections 4.1 and 4.5.2).
Consistency of the association between TCE exposure and liver cancer is further
supported by the results of meta-analyses. These meta-analyses found a statistically significant
increased summary relative risk estimate for liver and biliary tract cancer of 1.29 (95% CI: 1.07,
1. 56) with overall TCE exposure. Although there was no evidence of heterogeneity or
publication bias and the summary estimate was fairly insensitive to the use of alternative relative
risk estimates, the statistical significance of the summary estimate depends heavily on the one
large study by Raaschou-Nielsen et al. (2003). However, there were fewer adequate studies
available for meta-analysis of liver cancer (9 versus 17 for NHL and 15 for kidney), leading to
lower statistical power, even with pooling. Moreover, liver cancer is comparatively rarer, with
age-adjusted incidences roughly half or less those for kidney cancer or NHL; thus, fewer liver
cancer cases are generally observed in individual cohort studies.
(b) Strength of the observed association. In general, the observed associations between TCE
exposure and cancer are modest, with relative risks or odds ratios for overall TCE exposure
generally less than 2.0, and higher relative risks or odds ratios for high exposure categories.
Among the highest statistically significant relative risks were those reported for kidney cancer in
the studies by Henschler et al. (1995) (7.97 [95% CI: 2.59, 8.59]) and Vamvakas et al. (1998)
(10.80 [95% CI: 3.36, 34.75]). As discussed in U.S. EPA (2011, Section 4.5.3), risk magnitude
in both studies is highly uncertain due, in part, to possible selection biases, and neither was
included in the meta-analyses. However, the findings of these studies were corroborated, though
with lower reported relative risks, by later studies which overcame many of their deficiencies,
such as Brnning et al. (2003) (2.47 [95% CI: 1.36, 4.49]), Charbotel et al. (2006; 2009) (2.16
[95% CI: 1.02, 4.60] for the high cumulative exposure group), and Moore et al. (2010) (2.05
[95%> CI: 1.13, 3.73] for high confidence assessment of TCE). In addition, the very high
apparent exposure in the subjects of Henschler et al. (1995) and Vamvakas et al. (1998) et al.
may have contributed to their reported relative risks being higher than those in other studies.
Exposures in most population case-control studies are of lower overall TCE intensity compared
to exposures in Briining et al. (2003) and Charbotel et al. (2006; 2009), and, as would be
expected, observed relative risk estimates are lower (1.24 [95% CI: 1.03, 1.49]), Pesch et al.,
2000a; 1.30 [95% CI: 0.9, 1.9], (Dosemeci etal., 1999). A few high-quality cohort and case-
control studies reported statistically significant relative risks of approximately 2.0 with highest
exposure, including Zhao et al. (2005) (4.9 [95% CI: 1.23, 19.6] for high TCE score),
Raaschou-Nielsen et al. (2003) (1.7 [95% CI: 1.1, 2.4] for >5 year exposure duration, subcohort
with higher exposure]), Charbotel et al. (2006) (2.16 [95% CI: 1.02, 4.60] for high cumulative
exposure and 2.73 [95% CI: 1.06, 7.07] for high cumulative exposure plus peaks) and Moore et
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al. (2010) (2.23 [95% CI: 1.07, 4.64] for high cumulative exposure and 2.41 [95% CI: 1.05, 5.56]
for high average intensity TCE exposure).
Among the highest statistically significant relative risks reported for NHL were those of
Hansen et al. (2001) (3.1 [95% CI: 1.3, 6.1]), Hardell et al. (1994) (7.2 [95% CI: 1.3, 42]), the
latter a case-control study whose magnitude of risk is uncertain because of self-reported
occupational TCE exposure. A similar magnitude of risk was reported in Purdue et al. (2011) for
highest exposure (3.3 [95% CI: 1.1, 10.1], >234,000 ppm-hr, and 7.9 [95% CI: 1.8, 34.3], >360
ppm-hr/week). Observed relative risk estimates for liver cancer and overall TCE exposure are
generally more modest.
The strength of association between TCE exposure and cancer is modest with overall
TCE exposure. Large relative risk estimates are considered strong evidence of causality;
however, a modest risk does not preclude a causal association and may reflect a lower level of
exposure, an agent of lower potency, or a common disease with a high background level (U.S.
EPA, 2005b). Modest relative risk estimates have been observed with several well-established
human carcinogens such as benzene and secondhand smoke. Chance cannot explain the
observed association between TCE and cancer; statistically significant associations are found in a
number of the studies that contribute greater weight to the overall evidence, given their design
and statistical analysis approaches. In addition, other known or suspected risk factors can not
fully explain the observed elevations in kidney cancer relative risks. All kidney cancer case-
control studies included adjustment for possible confounding effects of smoking, and some
studies included body mass index, hypertension, and co-exposure to other occupational agents
such as cutting or petroleum oils. Cutting oils and petroleum oils, known as metalworking
fluids, have not been associatied with kidney cancer (Mirer. 2010: NIOSH, 1998). and potential
confounding by this occupational co-exposure is unable to explain the observed assocation with
TCE. Additionally, the associations between kidney cancer and TCE exposure remained in these
studies after statistical adjustment for possible known and suspected confounders. Charbotel et
al. (2005) observed a nonstatistically significantly kidney cancer risk with exposure to TCE
adjusted for cutting or petroleum oil exposures (1.96 [95% CI: 71, 5.37] for the high- cumulative
exposure group and 2.63 [95% CI: 0.79, 8,83] for high-exposure group with peaks).
All kidney cancer case-control studies adjusted for smoking except the Moore et al. (2010) study,
which reported that smoking did not significantly change the overall association with TCE
exposure. Although direct examination of smoking and other suspected kidney cancer risk
factors is usually not possible in cohort studies, confounding is less likely in Zhao et al. (2005).
given their use of an internal referent group and adjustment for socioeconomic status, an indirect
surrogate for smoking, and other occupational exposures. In addition, the magnitude of the lung
cancer risk in Raaschou-Nielsen et al. (2003) suggests a high smoking rate is unlikely and cannot
explain their finding on kidney cancer. Last, a meta-analysis of the nine cohort studies that
reported kidney cancer risks found a summary relative risk estimate for lung cancer of 0.96 (95%
CI: 0.76, 1.21) for overall TCE exposure and 0.96 (95% CI: 0.72, 1.27) for the highest exposure
group. These observations suggest that confounding by smoking is not an alternative
explanation for the kidney cancer meta-analysis results.
Few risk factors are recognized for NHL, with the exception of viruses and suspected
factors such as immunosuppression or smoking, which are associated with specific NHL
subtypes. Associations between NHL and TCE exposure are based on groupings of several NHL
subtypes. Three of the seven NHL case-control studies adjusted for age, sex and smoking in
statistical analyses (Miligi et al„ 2006; Wang et al., 2009), two others adjusted for age, sex and
education (Cocco et al„ 2010: Purdue et al„ 2011), and the other three case-control studies
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adjusted for age only or age and sex (Hardell et al.. 1994; Nordstrom et al.. 1998; Persson and
Fredrikson, 1999). Like for kidney cancer, direct examination of possible confounding in cohort
studies is not possible. The use of internal controls in some of the higher quality cohort studies
is intended to reduce possible confounding related to lifestyle differences, including smoking
habits, between exposed and referent subjects.
Heavy alcohol use and viral hepatitis are established risk factors for liver cancer, with
severe obesity and diabetes characterized as a metabolic syndrome associated with liver cancer.
Only cohort studies for liver cancer are available, and they were not able to consider these
possible risk factors.
(c) Specificity of the observed association. Specificity is generally not as relevant as other
aspects forjudging causality. As stated in the U.S. EPA Guidelines for Carcinogen Risk
Assessment (U.S. EPA. 2005b). based on our current understanding that many agents cause
cancer at multiple sites, and cancers have multiple causes, the absence of specificity does not
detract from evidence for a causal effect. Evidence for specificity could be provided by a
biological marker in tumors that was specific to TCE exposure. There is some evidence
suggesting particular VHL mutations in kidney tumors may be caused by TCE, but uncertainties
in these data preclude a definitive conclusion.
(d) Temporal relationship of the observed association. Each cohort study was evaluated for
the adequacy of the follow-up period to account for the latency of cancer development. The
studies with the greatest weight based on study design characteristics (e.g., those used in the
meta-analysis) all had adequate follow-up to assess associations between TCE exposure and
cancer. Therefore, the findings of those studies are consistent with a temporal relationship.
(e) Biological gradient (exposure-response relationship). Exposure-response relationships are
examined in the TCE epidemiologic studies only to a limited extent. Many studies examined
only overall "exposed" versus "unexposed" groups and did not provide exposure information by
level of exposure. Others do not have adequate exposure assessments to confidently distinguish
between levels of exposure. For example, many studies used duration of employment as an
exposure surrogate; however, this is a poor exposure metric given subjects may have differing
exposure intensity with similar exposure duration (NRC, 2006).
Three studies of kidney cancer reported a statistically significant trend of increasing risk
with increasing TCE exposure, Zhao et al. (2005) (p = 0.023 for trend with TCE score),
Charbotel et al. (2005; 2007) (p = 0.04 for trend with cumulative TCE exposure) and Moore et
al. (2010) (p = 0.02 for trend with cumulative TCE exposure). Charbotel et al. (2007) was
specifically designed to examine TCE exposure and had a high-quality exposure assessment and
the Moore et al. (2010) exposure assessment considered detailed information on jobs using
solvents. Zhao et al. (2005) also had a relatively well-designed exposure assessment. A positive
trend was also observed in one other study (Raaschou-Nielsen et al., 2003) with employment
duration).
Biological gradient is further supported by meta-analyses for kidney cancer using only
the highest exposure groups and accounting for possible reporting bias, which yielded a higher
summary relative risk estimate (1.58 [95% CI: 1.28, 1.96]) than for overall TCE exposure (1.27
[95% CI: 1.13, 1.43]). Although this analysis uses a subset of studies in the overall TCE
exposure analysis, the finding of higher risk in the highest exposure groups, where such groups
were available, is consistent with a trend of increased risk with increased exposure.
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The NHL case-control study of Purdue et al. (2011) reported a statistically significant
trend with TCE exposure (p = 0.02 for trend with average-weekly TCE exposure), and NHL risk
in Boice et al. (1999) appeared to increase with increasing exposure duration (p = 0.20 for
routine-intermittent exposed subjects). The borderline trend with TCE intensity in the case-
control studies of Wang et al. (2009) (p = 0.06) and Purdue et al. (2011) (p = 0.08 for trend with
cumulative TCE exposure) is consistent with their findings for average weekly TCE exposure.
As with kidney cancer, further support was provided by meta-analyses using only the highest
exposure groups, which yielded a higher summary relative risk estimate (1.43 [95% CI: 1.13,
1.82]) than for overall TCE exposure (1.23 [95% CI: 1.07, 1.42]). For liver cancer, the meta-
analyses using only the highest exposure groups yielded a lower, and nonstatistically significant,
summary estimate (1.28 [95% CI: 0.93, 1.77]) than for overall TCE exposure (1.29 [95% CI:
1.07, 1.56]). There were no case-control studies on liver cancer and TCE, and the cohort studies
generally had few liver cancer cases, making it more difficult to assess exposure-response
relationships. The one large study (Raaschou-Nielsen et al., 2003) used only duration of
employment, which is an inferior exposure metric.
(f) Biological plausibility. TCE metabolism is similar in humans, rats, and mice and results in
reactive metabolites. TCE is metabolized in multiple organs and metabolites are systemically
distributed. Several oxidative metabolites produced primarily in the liver, including CH, TCA
and DC A, are rodent hepatocarcinogens. Two other metabolites, DCVC and DCVG, which can
be produced and cleared by the kidney, have shown genotoxic activity, suggesting the potential
for carcinogenicity. Kidney cancer, NHL, and liver cancer have all been observed in rodent
bioassays (see below). The laboratory animal data for liver and kidney cancer are the most
robust, corroborated in multiple studies, sexes, and strains, although each has only been reported
in a single species and the incidences of kidney cancer are quite low. Lymphomas were only
reported to be statistically significantly elevated in a single study in mice, but one additional
mouse study reported elevated lymphoma incidence and one rat study reported elevated leukemia
incidence. In addition, there is some evidence both in humans and laboratory animals for kidney,
liver and immune system noncancer toxicity from TCE exposure. Several hypothesized modes
of action have been presented for the rodent tumor findings, although there are insufficient data
to support any one mode of action, and the available evidence does not preclude the relevance of
the hypothesized modes of action to humans. Activation of macrophages, natural killer cells,
and cytokine production (e.g., tumor necrosis factor), may also play an etiologic role in
carcinogenesis, and so the immune-related effects of TCE should also be considered. In
addition, the decreased in lymphocyte counts and subsets, including CD4+ T cells, and decreased
lymphocyte activation seen in TCE-exposed workers (Lan et al., 2010) also support the
biological plausibility of a role of TCE exposure in NHL.
(g) Coherence. Coherence is defined as consistency with the known biology. As discussed
under biological plausibility, the observance of kidney and liver cancer, and NHL in humans is
consistent with the biological processing and toxicity of TCE.
(h) Experimental evidence (from human populations). Few experimental data from human
populations are available on the relationship between TCE exposure and cancer. The only study
of a "natural experiment" (i.e., observations of a temporal change in cancer incidence in relation
to a specific event) notes that childhood leukemia cases appeared to be more evenly distributed
throughout Woburn, MA, after closure of the two wells contaminated with trichloroethylene and
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other organic solvents (MDPH. 1997).
(i) Analogy. Exposure to structurally related chlorinated solvents such as tetrachloroethylene
and dichloromethane have also been associated with kidney, lymphoid, and liver tumors in
humans, although the evidence for TCE is considered stronger.
Conclusion. In conclusion, based on the weight-of-evidence analysis for kidney cancer and in
accordance with U.S. EPA guidelines, TCE is characterized as "Carcinogenic to Humans." This
hazard descriptor is used when there is convincing epidemiologic evidence of a causal
association between human exposure and cancer. Convincing evidence is found in the
consistency of the kidney cancer findings. The consistency of increased kidney cancer relative
risk estimates across a large number of independent studies of different designs and populations
from different countries and industries provides compelling evidence given the difficulty, a
priori, in detecting effects in epidemiologic studies when the relative risks are modest, the
cancers are relatively rare, and therefore, individual studies have limited statistical power. This
strong consistency argues against chance, bias, and confounding as explanations for the elevated
kidney cancer risks. In addition, statistically significant exposure-response trends are observed
in high-quality studies. These studies were designed to examine kidney cancer in populations
with high TCE exposure intensity. These studies addressed important potential confounders and
biases, further supporting the observed associations with kidney cancer as causal. In a meta-
analysis of the 15 studies that met the inclusion criteria, a statistically significant summary
relative risk estimate was observed for overall TCE exposure (RRm: 1.27 [95% CI: 1.13, 1.43]).
The summary relative risk estimate was greater for the highest TCE exposure groups (RRm: 1.58
[95% CI: 1.28, 1.96]; n = 13 studies). Meta-analyses investigating the influence of individual
studies and the sensitivity of the results to alternate relative risk estimate selections found the
summary relative risk estimates to be highly robust. Furthermore, there was no indication of
publication bias or significant heterogeneity. It would require a substantial amount of negative
data from informative studies (i.e., studies having a high likelihood of TCE exposure in
individual study subjects and which meet, to a sufficient degree, the standards of epidemiologic
design and analysis in a systematic review) to contradict this observed association.
The evidence is strong but less convincing for NHL, where issues of (non-statistically
significant) study heterogeneity, potential publication bias, and weaker exposure-response results
contribute greater uncertainty. The evidence is more limited for liver cancer mainly because
only cohort studies are available and most of these studies have small numbers of cases. In
addition to the body of evidence described above pertaining to kidney cancer, NHL, and liver
cancer, the available epidemiologic studies also provide suggestive evidence of an association
between TCE exposure and other types of cancer, including bladder, esophageal, prostate,
cervical, breast, and childhood leukemia. Differences between these sets of data and the data for
kidney cancer, NHL, and liver cancer are fewer studies, a mixed pattern of observed risk
estimates and the general absence of exposure-response data from the studies using a quantitative
TCE-specific cumulative exposure measure.
II.A.3. ANIMAL CARCINOGENICITY DATA
Additional evidence of TCE carcinogenicity consists of increased incidences of tumors
reported in multiple chronic bioassays in rats and mice. In total, this database identifies some of
the same target tissues of TCE carcinogenicity also seen in epidemiological studies, including the
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kidney, liver, and lymphoid tissues.
Of particular note is the site-concordant finding of TCE-induced kidney cancer in rats. In
particular, low, but biologically and sometimes statistically significant, increases in the incidence
of kidney tumors were observed in multiple strains of rats treated with TCE by either inhalation
or corn oil gavage (Maltoni etal., 1986; NTP. 1988. 1990a). For instance, Maltoni et al. (1986)
reported that although only 4/130 renal adenocarcinomas in rats in the highest dose group, these
tumors had never been observed in over 50,000 Sprague-Dawley rats (untreated, vehicle-treated,
or treated with different chemicals) examined in previous experiments in the same laboratory In
addition, the gavage study by NCI (1976) and two inhalation studies by Henschler et al. (1980).
and Fukuda et al. (1983) each observed one renal adenoma or adenocarcinoma in some dose
groups and none in controls. The largest (but still small) incidences were observed in treated
male rats, only in the highest dose groups. However, given the small numbers, an effect in
females cannot be ruled out. Several studies in rats were limited by excessive toxicity,
accidental deaths, or deficiencies in reporting (NCI, 1976; NTP, 1988, 1990a). Individually,
therefore, these studies provide only suggestive evidence of renal carcinogenicity. Overall,
given the rarity of these types of tumors in the rat strains tested and the repeated similar results
across experiments and strains, these studies taken together support the conclusion that TCE is a
kidney carcinogen in rats, with males being more sensitive than females. No other tested
laboratory species (i.e., mice and hamsters) have exhibited increased kidney tumors, although
high incidences of kidney toxicity have been reported in mice (Maltoni et al„ 1986; NCI, 1976;
NTP, 1990a). The GSH-conjugation-derived metabolites suspected of mediating TCE-induced
kidney carcinogenesis have not been tested in a standard 2-year bioassay, so their role cannot be
confirmed definitively. However, it is clear that GSH conjugation of TCE occurs in humans and
that the human kidney contains the appropriate enzymes for bioactivation of GSH conjugates.
Therefore, the production of the active metabolites thought to be responsible for kidney tumor
induction in rats likely occurs in humans.
Statistically significant increases in TCE-induced liver tumors have been reported in
multiple inhalation and gavage studies with male Swiss mice and B6C3F1 mice of both sexes
(Anna et al„ 1994; Bull et al„ 2002; Herren-Freund et al„ 1987; Maltoni et al., 1986; NCI, 1976;
NTP, 1990a). In female Swiss mice, on the other hand, Fukuda et al. (1983), in CD-I (ICR,
Swiss-derived) mice, and Maltoni et al. (1986) both reported small, nonsignificant increases at
the highest dose by inhalation. Henschler et al. (1984; 1980) reported no increases in either sex
of Han:NMRI (also Swiss-derived) mice exposed by inhalation and ICR/HA (Swiss) mice
exposed by gavage. However, the inhalation study (Henschler et al„ 1980) had only 30 mice per
dose group and the gavage study (Henschler et al„ 1984) had dosing interrupted due to toxicity.
Studies in rats (Henschler et al., 1980; Maltoni et al„ 1986; NCI, 1976; NTP, 1988, 1990a) and
hamsters (Henschler et al., 1980) did not report statistically significant increases in liver tumor
induction with TCE treatment. However, several studies in rats were limited by excessive
toxicity or accidental deaths (NCI, 1976; NTP, 1988, 1990a), and the study in hamsters only had
30 animals per dose group. These data are inadequate for concluding that TCE lacks
hepatocarcinogenicity in rats and hamsters, but are indicative of a lower potency in these species.
Moreover, it is notable that a few studies in rats reported low incidences (too few for statistical
significance) of very rare biliary- or endothelial-derived tumors in the livers of some treated
animals (Fukuda et al., 1983; Henschler et al„ 1980; Maltoni et al., 1986). Further evidence for
the hepatocarcinogenicity of TCE is derived from chronic bioassays of the TCE oxidative
metabolites CH, TCA, and DCA in mice (e.g.. Bull et al„ 1990; DeAngelo et al., 1996;
DeAngelo et al., 2008; DeAngelo et al., 1999; George et al„ 2000; Leakey et al., 2003), all of
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which reported hepatocarcinogenicity. Very limited testing of these TCE metabolites has been
done in rats, with a single experiment reported in both Richmond et al. (1995) and DeAngelo et
al. (1996) finding statistically significant DCA-induced hepatocarcinogenicity. With respect to
TCA, DeAngelo et al. (1997), often cited as demonstrating lack of hepatocarcinogenicity in rats,
actually reported elevated adenoma multiplicity and carcinoma incidence from TCA treatment.
However, statistically, the role of chance could not be confidently excluded because of the low
number of animals per dose group (20-24 per treatment group at final sacrifice). Overall, TCE
and its oxidative metabolites are clearly carcinogenic in mice, with males more sensitive than
females and the B6C3F1 strain appearing to be more sensitive than the Swiss strain. Such strain
and sex differences are not unexpected, as they appear to parallel, qualitatively, differences in
background tumor incidence. Data in other laboratory animal species are limited. Thus, except
for DC A, which is carcinogenic in rats, inadequate evidence exists to evaluate the
hepatocarcinogenicity of these compounds in rats or hamsters. However, to the extent that there
is hepatocarcinogenic potential in rats, TCE is clearly less potent in the strains tested in this
species than in B6C3F1 and Swiss mice.
Additionally, there is more limited evidence for TCE-induced lymphatic cancers in rats
and mice, lung tumors in mice, and testicular tumors in rats. With respect to the lymphomas,
Henschler et al. (1980) reported statistically significant increases in lymphomas in female
Han:NMRI mice treated via inhalation. While Henschler et al. (1980) suggested these
lymphomas were of viral origin specific to this strain, subsequent studies reported increased
lymphomas in female B6C3F1 mice treated via corn oil gavage (NTP. 1990a) and leukemias in
male Sprague-Dawley and female August rats (Maltoni et al., 1986; NTP, 1988). However,
these tumors had relatively modest increases in incidence with treatment, and were not reported
to be increased in other studies. With respect to lung tumors, rodent bioassays have
demonstrated a statistically significant increase in pulmonary tumors in mice following chronic
inhalation exposure to TCE (Fukuda et al., 1983; Maltoni et al„ 1988; Maltoni et al., 1986).
Pulmonary tumors were not reported in other species tested (i.e., rats and hamsters) (Fukuda et
al., 1983; Henschler et al., 1980; Maltoni et al., 1988; Maltoni et al., 1986). Chronic oral
exposure to TCE led to a nonstatistically significant increase in pulmonary tumors in mice but,
again, not in rats or hamsters (Henschler et al„ 1984; Maltoni et al„ 1986; NCI, 1976; NTP,
1988, 1990a; Van Duuren et al., 1979). A lower response via oral exposure would be consistent
with a role of respiratory metabolism in pulmonary carcinogenicity. Finally, increased testicular
(interstitial cell and Leydig cell) tumors have been observed in rats exposed by inhalation and
gavage (Maltoni et al„ 1986; NTP, 1988, 1990b). Statistically significant increases were
reported in Sprague-Dawley rats exposed via inhalation (Maltoni et al., 1986) and Marshall rats
exposed via gavage (NTP, 1988). In three rat strains, ACI, August, and F344/N, a high (>75%)
control rate of testicular tumors was observed, limiting the ability to detect a treatment effect
(NTP, 1988, 1990a).
In summary, there is clear evidence for TCE carcinogenicity in rats and mice, with
multiple studies showing TCE to cause tumors at multiple sites. The apparent lack of site
concordance across laboratory animal species may be due to limitations in design or conduct in a
number of rat bioassays and/or genuine interspecies differences in sensitivity. Nonetheless, these
studies have shown carcinogenic effects across different strains, sexes, and routes of exposure,
and site-concordance is not necessarily expected for carcinogens. Of greater import is the
finding that there is support in experimental animal studies for the main cancers observed in
TCE-exposed humans—in particular, cancers of the kidney, liver, and lymphoid tissues.
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II.A.4. SUPPORTING DATA FOR CARCINOGENICITY
Additional evidence from toxicokinetic, toxicity, and mechanistic studies supports the
biological plausibility of TCE carcinogenicity in humans.
Toxicokinetic data indicates that TCE is well absorbed by all routes of exposure, and that
TCE absorption, distribution, metabolism, and excretion are qualitatively similar in humans and
rodents. There is evidence that TCE is systemically available, distributes to organs and tissues,
and undergoes systemic metabolism from all routes of exposure. Therefore, although the
strongest evidence from epidemiologic studies largely involves inhalation exposures, the
evidence supports TCE carcinogenicity being applicable to all routes of exposure. In addition,
there is no evidence of major qualitative differences across species in TCE absorption,
distribution, metabolism, and excretion. Extensive in vivo and in vitro data show that mice, rats,
and humans all metabolize TCE via two primary pathways: oxidation by CYPs and conjugation
with glutathione via GSTs. Several metabolites and excretion products from both pathways have
been detected in blood and urine from exposed humans as well as from at least one rodent
species. In addition, the subsequent distribution, metabolism, and excretion of TCE metabolites
are qualitatively similar among species. Therefore, humans possess the metabolic pathways that
produce the TCE metabolites thought to be involved in the induction of rat kidney and mouse
liver tumors, and internal target tissues of both humans and rodents experience a similar mix of
TCE and metabolites. See U.S. EPA (2011, Sections 3.1-3.4) for additional discussion of TCE
toxicokinetics. Quantitative interspecies differences in toxicokinetics do exist, and are addressed
through PBPK modeling (see U.S. EPA, 2011, Section 3.5 and Appendix A). Importantly, these
quantitative differences affect only interspecies extrapolations of carcinogenic potency, and do
not affect inferences as to the carcinogenic hazard for TCE.
Available mechanistic data do not suggest a lack of human carcinogenic hazard from
TCE exposure. In particular, these data do not suggest qualitative differences between humans
and test animals that would preclude any of the hypothesized key events in the carcinogenic
MO A in rodents from occurring in humans. For the kidney, the predominance of positive
genotoxicity data in the database of available studies of TCE metabolites derived from GSH
conjugation (in particular DCVC), together with toxicokinetic data consistent with their systemic
delivery to and in situ formation in the kidney, supports the conclusion that a mutagenic MOA is
operative in TCE-induced kidney tumors. While supporting the biological plausibility of this
hypothesized MOA, available data on the von Hippel-Lindau (VHL) gene in humans or
transgenic animals do not conclusively elucidate the role of VHL mutation in TCE-induced renal
carcinogenesis. Cytotoxicity and compensatory cell proliferation, similarly presumed to be
mediated through metabolites formed after GSH-conjugation of TCE, have also been suggested
to play a role in the MOA for renal carcinogenesis, as high incidences of nephrotoxicity have
been observed in animals at doses that induce kidney tumors. Human studies have reported
markers for nephrotoxicity at current occupational exposures, although data are lacking at lower
exposures. Nephrotoxicity is observed in both mice and rats, in some cases with nearly 100%
incidence in all dose groups, but kidney tumors are only observed at low incidences in rats at the
highest tested doses. Therefore, nephrotoxicity alone appears to be insufficient, or at least not
rate-limiting, for rodent renal carcinogenesis, since maximal levels of toxicity are reached before
the onset of tumors. In addition, nephrotoxicity has not been shown to be necessary for kidney
tumor induction by TCE in rodents. In particular, there is a lack of experimental support for
causal links, such as compensatory cellular proliferation or clonal expansion of initiated cells,
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between nephrotoxicity and kidney tumors induced by TCE. Furthermore, it is not clear if
nephrotoxicity is one of several key events in a MO A, if it is a marker for an "upstream" key
event (such as oxidative stress) that may contribute independently to both nephrotoxicity and
renal carcinogenesis, or if it is incidental to kidney tumor induction. Moreover, while
toxicokinetic differences in the GSH conjugation pathway along with their uncertainty are
addressed through PBPK modeling, no data suggest that any of the proposed key events for
TCE-induced kidney tumors in rats are precluded in humans. See U.S. EPA (2011,
Section 4.4.7) for additional discussion of the MOA for TCE-induced kidney tumors. Therefore,
TCE-induced rat kidney tumors provide additional support for the convincing human evidence of
TCE-induced kidney cancer, with mechanistic data supportive of a mutagenic MOA.
With respect to other tumor sites, data are insufficient to conclude that any of the other
hypothesized MO As are operant. In the liver, a mutagenic MOA mediated by CH, which has
evidence for genotoxic effects, or some other oxidative metabolite of TCE cannot be ruled out,
but data are insufficient to conclude it is operant. A second MOA hypothesis for TCE-induced
liver tumors involves activation of the peroxisome proliferator activated receptor alpha (PPARa)
receptor. Clearly, in vivo administration of TCE leads to activation of PPARa in rodents and
likely does so in humans as well. However, the evidence as a whole does not support the view
that PPARa is the sole operant MOA mediating TCE hepatocarcinogenesis. Rather, there is
evidential support for multiple TCE metabolites and multiple toxicity pathways contributing to
TCE-induced liver tumors. Furthermore, recent experiments have demonstrated that PPARa
activation and the sequence of key events in the hypothesized MOA are not sufficient to induce
hepatocarcinogenesis (Yang et al„ 2007). Moreover, the demonstration that the PPARa agonist
di(2-ethylhexyl) phthalate induces tumors in PPARa-null mice supports the view that the events
comprising the hypothesized PPARa activation MOA are not necessary for liver tumor induction
in mice by this PPARa agonist (Ito et al.. 2007). See U.S. EPA (2011, Section 4.5.7) for
additional discussion of the MOA for TCE-induced liver tumors. For mouse lung tumors, as
with the liver, a mutagenic MOA involving CH has also been hypothesized, but there are
insufficient data to conclude that it is operant. A second MOA hypothesis for mouse lung
tumors has been posited involving other effects of oxidative metabolites including cytotoxicity
and regenerative cell proliferation, but experimental support remains limited, with no data on
proposed key events in experiments of duration 2 weeks or longer. See U.S. EPA (2011,
Section 4.7.4) for additional discussion of the MOA for TCE-induced lung tumors. A MOA
subsequent to in situ oxidative metabolism, whether involving mutagenicity, cytotoxicity, or
other key events, may also be relevant to other tissues where TCE would undergo CYP
metabolism. For instance, CYP2E1, oxidative metabolites, and protein adducts have been
reported in the testes of rats exposed to TCE, and, in some rat bioassays, TCE exposure
increased the incidence of rat testicular tumors. However, inadequate data exist to adequately
define a MOA hypothesis for this tumor site (see U.S. EPA, 2011, Section 4.8.2.3 for additional
discussion of the MOA for TCE-induced testicular tumors).
II.B. QUANTITATIVE ESTIMATE OF CARCINOGENIC RISK FROM ORAL
EXPOSURE
II.B.1. SUMMARY OF RISK ESTIMATES
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II.B. 1.1. Oral Slope Factor - EPA has concluded, by a weight of evidence evaluation, that
trichloroethylene is carcinogenic by a mutagenic mode of action for induction of kidney tumors.
According to the Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure
to Carcinogens {Supplemental Guidance) (U.S. EPA. 2005c) those exposed to carcinogens with
a mutagenic mode of action are assumed to have increased early-life susceptibility. Data for
trichloroethylene are not sufficient to develop separate risk estimates for childhood exposure.
The oral slope factor of 4.6 x 10"2 per mg/kg/day, calculated from data from adult exposure, does
not reflect presumed increased early-life susceptibility to kidney tumors for this chemical.
Generally, the application of age-dependent adjustment factors (ADAFs) is recommended when
assessing cancer risks for a carcinogen with a mutagenic mode of action. However, as illustrated
in the detailed example calculation for oral drinking water exposures to TCE in Section 5.2.3.3.2
of the Toxicological Review of Trichloroethylene (U.S. EPA, 2011), because the ADAF
adjustment applies only to the kidney cancer component of the total cancer risk estimate, the
impact of the adjustment on full lifetime risk is minimal and the adjustment might reasonably be
omitted, given the greater complexity of the ADAF calculations for TCE. Nonetheless, for
exposure scenarios with increasing proportions of exposure during early life, the impact of the
ADAF adjustment becomes more pronounced and the importance of applying the ADAFs
increases.
Risk Assessment Considerations: The Supplemental Guidance establishes ADAFs for three
specific age groups. The current ADAFs and their age groupings are 10 for <2 years, 3 for 2 to
<16 years, and 1 for 16 years and above (U.S. EPA, 2005c). The 10-fold and 3-fold adjustments
in slope factor are to be combined with age-specific exposure estimates when estimating kidney
cancer risks from early life (<16 years age) exposure to trichloroethylene. These ADAFs and
their age groups were derived from the 2005 Supplemental Guidance, and they may be revised
over time. The most current information on the application of ADAFs for cancer risk assessment
can be found at www.epa.gov/cancerguidelines/. In estimating risk, EPA recommends using
age-specific values for both exposure and cancer potency; for trichloroethylene, age-specific
values for cancer potency for kidney tumors are calculated using the appropriate ADAFs. A
cancer risk is derived for each age group, including adjusted kidney cancer potency values and
unadjusted potency values for liver cancer and NHL, and these are summed across age groups to
obtain the total risk for the exposure period of interest (see Section 6 of the Supplemental
Guidance and Section 5.2.3.3.2 of the Toxicological Review of Trichloroethylene).
The oral slope factor, calculated from adult exposure, is equivalent to the risk (as a
fraction, i.e., 0.01 here) divided by the LED0i, the 95% lower bound on the exposure associated
with an 1% extra cancer risk, and represents an upper bound risk estimate for continuous lifetime
exposure without consideration of increased early-life susceptibility due to trichloroethylene's
mutagenic mode of action for kidney tumors. A 1% extra risk level is used for the determination
of the point of depature (POD) for low-exposure extrapolation because the exposure-response
analysis is based on epidemiologic data, which normally demonstrate lower cancer response
rates than rodent bioassays; an LEDio is not calculated because it would involve an upward
extrapolation for these data.
Adult-based oral slope factor - 4.6 x 10" per mg/kg/day
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Adult-based LEDoi, lower 95% bound on exposure at 1% extra risk - 0.21 mg/kg/day*
Adult-based ED0i, central estimate of exposure at 1% extra risk - 0.46 mg/kg/day**
The slope of the linear extrapolation from the central estimate ED0i is
0.01/(0.46 mg/kg/day) = 0.022 per mg/kg/day.
The slope factor for trichloroethylene should not be used with exposures exceeding 10
mg/kg/d, because above this level, the route-to-route extrapolation relationship is no
longer linear. Additionally, it is recommended that the application of ADAFs to (the
kidney cancer component of) this slope factor be considered when assessing cancer risks
to individuals exposed in early life (i.e., <16 years old), as discussed above (U.S. EPA.
2005b; U.S. EPA, 2011, Section 5.2.3.3.2).
* The oral slope factor estimate for TCE is actually calculated from route-to-route extrapolation
of the inhalation unit risk estimate for kidney cancer with a factor of 5 applied to include NHL
and liver cancer risks (U.S. EPA, 2011, Section 5.2.2.3). The LED0i can be back-calculated, in
abbreviated form, as follows: total cancer LEDoi = kidney cancer LECoi in ppm / 1.70
ppm/(mg/kg/day) /5 = 1.82 ppm / 1.70 ppm/(mg/kg/day) /5 = 0.21 mg/kg/day.
** The EDoi can be back-calculated as in the above footnote but using the kidney cancer ECoi in
place of the LECoi; thus, ED0i = 3.87 ppm /1.70 ppm/(mg/kg/day) /5 = 0.46 mg/kg/day.
II.B. 1.2. Drinking Water Concentrations at Specified Risk Levels
Drinking water unit risk and concentrations at specified risk levels are not provided for
trichloroethylene. Since trichloroethylene is carcinogenic by a mutagenic mode of action for
kidney tumors and increased susceptibility to kidney tumors is assumed for early-life exposures
(<16 years of age), the unit risk and concentrations at a specified risk levels will change based on
the age of the individuals in the exposed group. Risk assessors should use the oral slope factor
and current EPA guidance to assess risk based on site-specific populations and exposure
conditions. The most current information on the application of ADAFs for cancer risk
assessment can be found at www.epa.gov/cancerguidelines/. A detailed example application of
ADAFs for oral drinking water exposures is provided in Section 5.2.3.3.2 of the Toxicological
Review of Trichloroethylene (U.S. EPA. 2011).
II.B. 1.3. Modeling Approach and Extrapolation Method
The oral slope factor for trichloroethylene cancer risk, without consideration of increased
early-life susceptibility due to trichloroethylene's mutagenic mode of action for kidney tumors,
is derived from route-to-route extrapolation of the inhalation unit risk for trichloroethylene, using
a PBPK model. As discussed in more detail below (II.C.2 and II.C.3), the inhalation unit risk for
trichloroethylene is based on three separate target tissue sites - kidney, lymphoid tissue, and
liver. Because different internal dose metrics are preferred for each target tissue site, a separate
route-to-route extrapolation was performed for each site-specific unit risk estimate. The
approach taken is to apply the human PBPK model in the low-dose range, where external and
internal doses are linearly related, to derive a conversion that is the ratio of internal dose per
mg/kg/d to internal dose per ppm. The expected value of the population mean for this
conversion factor (in ppm per mg/kg/d) was used to extrapolate each inhalation unit risk in units
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of risk per ppm to an oral slope factor in units of risk per mg/kg/d.
II.B.2. DOSE-RESPONSE DATA
See II.C.2, below.
II.B.3. ADDITIONAL COMMENTS
As discussed above, the weight of evidence supports a mutagenic mode of action for
trichloroethylene kidney carcinogenicity. Generally, in the absence of chemical-specific data to
evaluate differences in susceptibility, increased early-life susceptibility is assumed for
carcinogens with a mutagenic mode of action and application of the ADAFs to the adult-based
unit risk estimate, in accordance with the Supplemental Guidance (U.S. EPA. 2005c). is
recommended. However, as illustrated in the example calculation in Section 5.2.3.3.2 of the
Toxicological Review of Trichloroethylene (U.S. EPA. 2011). because the ADAF adjustment
applies only to the kidney cancer component of the total cancer risk estimate, the impact of the
adjustment on full lifetime risk is minimal and the adjustment might reasonably be omitted,
given the greater complexity of the ADAF calculations for TCE. Nonetheless, for exposure
scenarios with increasing proportions of exposure during early life, the impact of the ADAF
adjustment becomes more pronounced and the importance of applying the ADAFs increases.
Please consult the example in Section 5.2.3.3.3 (U.S. EPA. 2011) when applying the ADAFs for
oral TCE exposures.
The adult-based oral slope factor estimate presented above (4.6x10" per mg/kg/d) is for
total cancer incidence, reflecting the incidence risks for kidney cancer (renal cell carcinoma,
RCC), NHL, and liver cancer. The adult-based oral slope factor estimates for the separate cancer
types were 9.33 xlO"3 per mg/kg/d for RCC, 2.16><10"2 per mg/kg/d for NHL, and 1.55xl0"2 per
mg/kg/d for liver cancer.
II.B.4. DISCUSSION OF CONFIDENCE
The oral slope factor estimate is based on good-quality human data, thus avoiding
uncertainties inherent in interspecies extrapolation. Uncertainties with respect to the inhalation
unit risk, from which the oral slope factor was derived via route-to-route extrapolation, are
discussed in II.C.4, below. In general, uncertainty in PBPK model-based route-to-route
extrapolation is relatively low (Chiu. 2006; Chiu and White. 2006). In this particular case,
extrapolation using different dose metrics yielded expected population mean risks within about a
2-fold range, and, for any particular dose metric, the 95% confidence interval for the
extrapolated population mean risks for each site spanned a range of no more than about 3-fold.
This oral slope factor estimate is further supported by estimates from multiple rodent
2 i
bioassays, the most sensitive of which range from 3x10 to 3 x 10 per mg/kg/d. From the
oral bioassays selected for analysis (U.S. EPA, 2011, Section 5.2.1.1), and using the preferred
PBPK model-based dose metrics, the oral unit risk estimate for the most sensitive sex/species is
3 x 10_1 per mg/kg/d, based on kidney tumors in male Osborne-Mendel rats (NTP, 1988). The
oral unit risk estimate for testicular tumors in male Marshall rats (NTP. 1988) is somewhat lower
_2
at 7 x 10 per mg/kg/d. The next most sensitive sex/species result from the oral studies is for
_2
male mouse liver tumors (NCI, 1976), with an oral unit risk estimate of 3 x 10 per mg/kg/d. In
addition, the 90% confidence intervals for male Osborne-Mendel rat kidney tumors (NTP, 1988),
29
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male F344 rat kidney tumors (NTP, 1990a). and male Marshall rat testicular tumors (NTP,
1988), derived from the quantitative analysis of PBPK model uncertainty, all included the
_2
estimate based on human data of 5 x 10 per mg/kg/d, while the upper 95% confidence bound
_2
for male mouse liver tumors from NCI (1976) was slightly below this value at 4 x 10 per
mg/kg/d. Furthermore, PBPK model-based route-to-route extrapolation of the most sensitive
endpoint from the inhalation bioassays, male rat kidney tumors from Maltoni et al. (1986). leads
to an oral unit risk estimate of 1 x 10 1 per mg/kg/d, with the preferred estimate based on human
data falling within the route-to-route extrapolation of the 90% confidence interval. Finally, for
all these estimates, the ratios of BMDs to the BMDLs did not exceed a value of 3, indicating that
the uncertainties in the dose-response modeling for determining the POD in the observable range
are small.
Therefore, although there are uncertainties in these various estimates (U.S. EPA, 2011,
Sections 5.2.1.4, 5.2.2.1.3, 5.2.2.2, and 5.2.2.3), confidence in the oral slope factor estimate of
_2
5x10 per mg/kg/d, resulting from PBPK model-based route-to-route extrapolation of the
inhalation unit risk estimate based on the human kidney cancer risks reported in
Charbotel et al. (2006) and adjusted for potential risk for tumors at multiple sites (U.S. EPA,
2011, Section 5.2.2.2), is further increased by the similarity of this estimate to estimates based on
multiple rodent data sets.
_II.C. QUANTITATIVE ESTIMATE OF CARCINOGENIC RISK FROM
INHALATION EXPOSURE
II.C.1. SUMMARY OF RISK ESTIMATES
II.C. 1.1. Inhalation Unit Risk - EPA has concluded, by a weight of evidence evaluation,
that trichloroethylene is carcinogenic by a mutagenic mode of action for induction of kidney
tumors. According to the Supplemental Guidance for Assessing Susceptibility from Early-Life
Exposure to Carcinogens {Supplemental Guidance) (U.S. EPA, 2005c) those exposed to
carcinogens with a mutagenic mode of action are assumed to have increased early-life
susceptibility. Data for trichloroethylene are not sufficient to develop separate risk estimates for
childhood exposure. The inhalation unit risk of 4.1 x 10"6 per (J,g/m3, calculated from data from
adult exposure, does not reflect presumed increased early-life susceptibility to kidney tumors for
this chemical. Generally, the application of age-dependent adjustment factors (ADAFs) is
recommended when assessing cancer risks for carcinogens with a mutagenic mode of action.
However, as illustrated in the detailed example calculation for inhalation exposures to TCE in
Section 5.2.3.3.1 of the Toxicological Review of Trichloroethylene (U.S. EPA, 2011), because
the ADAF adjustment applies only to the kidney cancer component of the total cancer risk
estimate, the impact of the adjustment on full lifetime risk is minimal and the adjustment might
reasonably be omitted, given the greater complexity of the ADAF calculations for TCE.
Nonetheless, for exposure scenarios with increasing proportions of exposure during early life, the
impact of the ADAF adjustment becomes more pronounced and the importance of applying the
ADAFs increases.
Risk Assessment Considerations: The Supplemental Guidance establishes ADAFs for three
specific age groups. The current ADAFs and their age groupings are 10 for <2 years, 3 for 2 to
30
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<16 years, and 1 for 16 years and above (U.S. EPA. 2005c). The 10-fold and 3-fold adjustments
in slope factor are to be combined with age-specific exposure estimates when estimating kidney
cancer risks from early life (<16 years age) exposure to trichloroethylene. These ADAFs and
their age groups were derived from the 2005 Supplemental Guidance, and they may be revised
over time. The most current information on the application of ADAFs for cancer risk assessment
can be found at www.epa.gov/cancerguidelines/. In estimating risk, EPA recommends using
age-specific values for both exposure and cancer potency; for trichloroethylene, age-specific
values for cancer potency for kidney tumors are calculated using the appropriate ADAFs. A
cancer risk is derived for each age group, including adjusted kidney cancer potency values and
unadjusted potency values for liver cancer and NHL, and these are summed across age groups to
obtain the total risk for the exposure period of interest (see Section 6 of the Supplemental
Guidance and Section 5.2.3.3.1 of the Toxicological Review of Trichloroethylene).
The inhalation unit risk, calculated from adult exposure, is equivalent to the risk (as a
fraction, i.e., 0.01 here) divided by the LECoi, the 95% lower bound on the exposure associated
with an 1% extra cancer risk, and represents an upper bound risk estimate for continuous lifetime
exposure without consideration of increased early-life susceptibility due to trichloroethylene's
mutagenic mode of action for kidney tumors. A 1% extra risk level is used for the determination
of the point of depature (POD) for low-exposure extrapolation because the exposure-response
analysis is based on epidemiologic data, which normally demonstrate lower cancer response
rates than rodent bioassays; an LECio is not calculated because it would involve an upward
extrapolation for these data.
Adult-based unit risk estimate -4.1 x 10"6 per (J,g/m3
"3
Adult-based LECoi, lower 95% bound on exposure at 1% extra risk - 2.4 mg/m *
"3
Adult-based ECoi, central estimate of exposure at 1% extra risk - 5.2 mg/m **
The slope of the linear extrapolation from the central estimate ECoi is
0.01/(5.2 mg/m3) = 1.9 x 10"6 per (J,g/m3
Additionally, it is recommended that the application of ADAFs to (the kidney cancer
component of) this unit risk estimate be considered when assessing cancer risks to
individuals exposed in early life (i.e., <16 years old), as discussed above (U.S. EPA,
2005; U.S. EPA, 2011, Section 5.2.3.3.10J.S. EPA. 2005aY
*The inhalation unit risk estimate for TCE is calculated from the inhalation unit risk estimate for
kidney cancer with a factor of 4 applied to include NHL and liver cancer risks (U.S. EPA, 2011,
Section 5.2.2.2). The LECoi can be back-calculated, in abbreviated form, as follows: total
cancer LECoi = kidney cancer LECoi/4 = 1.82 ppm 14 = 0.455 ppm x (5.374 mg/m3)/ppm = 2.4
mg/m3.
** The ECoi can be back-calculated as in the above footnote but using the kidney cancer ECoi in
place of the LECoi; thus, ECoi = 3.87 ppm 14 = 0.968 ppm x (5.374 mg/m3)/ppm = 5.2 mg/m3.
Air Concentrations at Specified Risk Levels
31
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Air concentrations at specified risk levels are not provided for trichloroethylene. Since
trichloroethylene is carcinogenic by a mutagenic mode of action for kidney tumors and
increased susceptibility to kidney tumors is assumed for early-life exposures (<16 years of age),
the concentrations at specified risk levels will change based on the age of the individuals in the
exposed group. Risk assessors should use the unit risk and current EPA guidance to assess risk
based on site-specific populations and exposure conditions. The most current information on the
application of ADAFs for cancer risk assessment can be found at
www.epa.gov/cancerguidelines/. A detailed example application of ADAFs for TCE inhalation
exposures is provided in Section 5.2.3.3.1 of the Toxicological Review of Trichloroethylene (U.S.
EPA. 2011).
II.C. 1.2. Exposure-Response Model and Extrapolation Method
A weighted linear regression model was used to model the exposure-response data on
kidney cancer (renal cell carcinoma, RCC) incidence to obtain a slope estimate (regression
coefficient) for the relative risk of RCC versus cumulative exposure. The regression coefficient
was used in a lifetable analysis to estimate the LECoi, which was used as the POD for linear
extrapolation to generate the unit risk estimate. Because there is evidence from human (and
rodent) studies for increased risks of NHL and liver cancer, the inhalation unit risk estimate
derived from human data for RCC incidence was adjusted to account for potential increased risk
of those tumor types. To make this adjustment, a factor accounting for the relative contributions
to the extra risk for cancer incidence from TCE exposure for these three tumor types combined
versus the extra risk for RCC alone was estimated, and this factor was applied to the unit risk
estimate for RCC to obtain a unit risk estimate for the three tumor types combined (i.e., lifetime
extra risk for developing any of the 3 types of tumors). This factor was based on human
surveillance data on the background risk of these tumors and human epidemiologic data on the
relative risk of these tumors associated with TCE exposure.
II.C.2. EXPOSURE-RESPONSE DATA
For the unit risk of kidney cancer (renal cell carcinoma): Conditional logistic regression results
for renal cell carcinoma incidence, matching on sex and age, adjusted for tobacco smoking and
body mass index; data from the Charbotel et al. (2006) study in the Arve Valley of France (U.S.
EPA, 2011, Sections 4.4, 5.2.2.1.1, and Appendix B):
Cumulative exposure
category
Mean Cumulative exposure
(ppm x years)
Adjusted OR
(95% CI)
Nonexposed
1
Low
62.4
1.62 (0.75, 3.47)
Medium
253.2
1.15 (0.47, 2.77)
High
925.0
2.16(1.02,4.60)
CI = confidence interval.
For adjustment of the inhalation unit risk for multiple sites: The relative contributions to the
extra risk for cancer from TCE exposure for multiple tumor types (NHL and liver cancer in
32
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addition to RCC) was estimated based on two different data sets. The first calculation was based
on the results of the meta-analysis of human epidemiologic data for the three tumor types (U.S.
EPA, 2011, Appendix C); the second calculation was based on the results of the Raaschou-
Nielsen et al. (2003) study, the larget single human epidemiologic study by far with relative risk
estimates for all three tumor types:
RR
Ro
Rx
Extra risk
Ratio to
kidney value
Calculation #1: using RR estimates from the meta-analyses
Kidney (RCC)
1.27
0.0107
0.01359
0.002920
1
NHL
1.23
0.0202
0.02485
0.004742
1.62
Liver (& biliary) cancer
1.29
0.0066
0.008514
0.001927
0.66
sum
0.009589
3.28
Kidney + NHL only
sum
0.007662
2.62
Calculation #2: using RR estimates from Rasschou-Nielsen et al. (2003)
Kidney (RCC)
1.20
0.0107
0.01284
0.002163
1
NHL
1.24
0.0202
0.02505
0.004948
2.29
Liver (& biliary) cancer
1.35
0.0066
0.008910
0.002325
1.07
sum
0.009436
4.36
Kidney + NHL only
sum
0.007111
3.29
RR = relative risk.
Ro = lifetime risk in an unexposed population (from SEER statistics)
Rx = lifetime risk in the exposed population = RR x Ro
Both of these calculations suggest that a factor of 4 (within 25% of either value; and equal to the
arithmetic or geometric mean, rounded to 1 significant figure) is reasonable for adjusting the unit
risk estimate based on RCC alone to include the combined risk of RCC, NHL, and liver cancer.
II.C.3. ADDITIONAL COMMENTS
As discussed above, the weight of evidence supports a mutagenic mode of action for
trichloroethylene kidney carcinogenicity. Generally, in the absence of chemical-specific data to
evaluate differences in susceptibility, increased early-life susceptibility is assumed for
carcinogens with a mutagenic mode of action and application of the ADAFs to the adult-based
unit risk estimate, in accordance with the Supplemental Guidance (U.S. EPA. 2005c). is
recommended. However, as illustrated in the example calculation in Section 5.2.3.3.1 of the
Toxicological Review of Trichloroethylene (U.S. EPA. 2011). because the ADAF adjustment
applies only to the kidney cancer component of the total cancer risk estimate, the impact of the
adjustment on full lifetime risk is minimal and the adjustment might reasonably be omitted,
given the greater complexity of the ADAF calculations for TCE. Nonetheless, for exposure
scenarios with increasing proportions of exposure during early life, the impact of the ADAF
33
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adjustment becomes more pronounced and the importance of applying the ADAFs increases.
Please consult the example in Section 5.2.3.3.1 (U.S. EPA, 2011) when applying the ADAFs for
inhalation TCE exposures.
The adult-based unit risk estimate presented above (4.1 x 10"6 per (J,g/m3) is for total cancer
incidence, reflecting the incidence risks for kidney cancer (RCC), NHL, and liver cancer. The
adult-based unit risk estimates for the separate cancer types were 1.02x 10"6 per (j,g/m3 for RCC,
2.05xl0"6 per (J,g/m3 for NHL, and 1.02xl0"6 per (J,g/m3 for liver cancer.
II.C.4. DISCUSSION OF CONFIDENCE
Some primary sources of uncertainty in the inhalation unit risk estimates are briefly
discussed below. The two major sources of uncertainty in quantitative cancer risk estimates are
generally interspecies extrapolation and high-dose to low-dose extrapolation. The unit risk
estimate for RCC incidence derived from the Charbotel et al. (2006) results is not subject to
interspecies uncertainty because it is based on human data. A major uncertainty remains in the
extrapolation from occupational exposures to lower environmental exposures. There was some
evidence of a contribution to increased RCC risk from peak exposures; however, there remained
an apparent dose-response relationship for RCC risk with increasing cumulative exposure
without peaks, and the OR for exposure with peaks compared to exposure without peaks was not
significantly elevated (Charbotel et al.. 2006). Although the actual exposure-response
relationship at low exposure levels is unknown, the conclusion that a mutagenic MOA is
operative for TCE-induced kidney tumors supports the linear low-dose extrapolation that was
used (U.S. EPA. 2005b).
Another source of uncertainty in the cancer unit risk estimate is the dose-response model
used to model the study data to estimate the POD. A weighted linear regression across the
categorical ORs was used to obtain a slope estimate; use of a linear model in the observable
range of the data is often a good general approach for human data because epidemiological data
are frequently too limited (i.e., imprecise) to clearly identify an alternate model (U.S. EPA.
2005b). The Charbotel et al. study is a relatively small case-control study, with only 86 RCC
cases, 37 of which had TCE exposure; thus, the dose-response data upon which to specify a
model are indeed limited. In accordance with U.S. EPA's Guidelines for Carcinogen Risk
Assessment, the lower bound on the ECoi is used as the POD; this acknowledges some of the
uncertainty in estimating the POD from the available dose-response data. In this case, the
statistical uncertainty associated with the ECoi is relatively small, as the ratio between the ECoi
and the LECoi for RCC incidence is about 2-fold.
An important source of uncertainty in the underlying Charbotel et al. (2006) study is the
retrospective estimation of TCE exposures in the study subjects. This case-control study was
conducted in the Arve Valley in France, a region with a high concentration of workshops
devoted to screw cutting, which involves the use of TCE and other degreasing agents. Since the
1960s, occupational physicians of the region have collected a large quantity of well-documented
measurements, including TCE air concentrations and urinary metabolite levels (Fevotte et al.,
2006). The study investigators conducted a comprehensive exposure assessment to estimate
cumulative TCE exposures for the individual study subjects, using a detailed occupational
questionnaire with a customized task-exposure matrix for the screw-cutting workers and a more
general occupational questionnaire for workers exposed to TCE in other industries (Fevotte et
al., 2006). The exposure assessment even attempted to take dermal exposure from hand-dipping
34
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practices into account by equating it with an equivalent airborne concentration based on
biological monitoring data. Despite the appreciable effort of the investigators, considerable
uncertainty associated with any retrospective exposure assessment is inevitable, and some
exposure misclassification is unavoidable. Such exposure misclassification was most likely for
the 19 deceased cases and their matched controls, for which proxy respondents were used, and
for exposures outside the screw-cutting industry (295 of 1,486 identified job periods involved
TCE exposure; 120 of these were not in the screw-cutting industry).
Although the exposure estimates from Moore et al. (2010) were not considered to be as
quantitatively accurate as those of Charbotel et al. (2006). as discussed in U.S. EPA (2011,
Section 5.2.2), it is worth noting, in the context of uncertainty in the exposure assessment, that
the exposure estimates in Moore et al. (2010) are substantially lower than those of Charbotel et
al. (2006) for comparable OR estimates. For example, for all subjects and high-confidence
assessments only, respectively, Moore et al. (2010) report OR estimates of 1.19 and 1.77 for
cumulative exposures < 1.58 ppm x years and 2.02 and 2.23 for cumulative exposures > 1.58
ppm x years. Charbotel et al. (2006). on the other hand, report OR estimates for all subjects of
1.62, 1.15, and 2.16 for mean cumulative exposures of 62.4, 253.2, and 925.0 ppm x years,
respectively. If the exposure estimates for Charbotel et al. (2006) are overestimated, as
suggested by the exposure estimates from Moore et al. (2010), the slope of the linear regression
model, and hence the unit risk estimate, would be correspondingly underestimated.
Another source of uncertainty in the Charbotel et al. (2006) study is the possible
influence of potential confounding or modifying factors. This study population, with a high
prevalence of metal-working, also had relatively high prevalences of exposure to petroleum oils,
cadmium, petroleum solvents, welding fumes, and asbestos (Fevotte et al.. 2006). Other
exposures assessed included other solvents (including other chlorinated solvents), lead, and
ionizing radiation. None of these exposures was found to be significantly associated with RCC
at ap = 0.05 significance level. Cutting fluids and other petroleum oils were associated with
RCC at ap = 0.1 significance level; however, further modeling suggested no association with
RCC when other significant factors were taken into account (Charbotel et al„ 2006). Moreover,
a review of other studies suggested that potential confounding from cutting fluids and other
petroleum oils is of minimal concern (U.S. EPA, 2011, Section 4.4.2.3). Nonetheless, a
sensitivity analysis was conducted using the OR estimates further adjusted for cutting fluids and
other petroleum oils from the unpublished report by Charbotel et al. (2005), and an essentially
"3
identical unit risk estimate of 5.46 x 10" per ppm was obtained. In addition, the medical
questionnaire included familial kidney disease and medical history, such as kidney stones,
infection, chronic dialysis, hypertension, and use of anti-hypertensive drugs, diuretics, and
analgesics. Body mass index (BMI) was also calculated, and lifestyle information such as
smoking habits and coffee consumption was collected. Univariate analyses found high levels of
smoking and BMI to be associated with increased odds of RCC, and these two variables were
included in the conditional logistic regressions. Thus, although impacts of other factors are
possible, this study took great pains to attempt to account for potential confounding or modifying
factors.
Some other sources of uncertainty associated with the epidemiological data are the dose
metric and lag period. As discussed above, there was some evidence of a contribution to
increased RCC risk from peak TCE exposures; however, there appeared to be an independent
effect of cumulative exposure without peaks. Cumulative exposure is considered a good
measure of total exposure because it integrates exposure (levels) over time. If there is a
contributing effect of peak exposures, not already taken into account in the cumulative exposure
35
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metric, the linear slope may be overestimated to some extent. Sometimes cancer data are
modeled with the inclusion of a lag period to discount more recent exposures not likely to have
contributed to the onset of cancer. In an unpublished report, Charbotel et al. (2005) also present
the results of a conditional logistic regression with a 10-year lag period, and these results are
very similar to the untagged results reported in their published paper, suggesting that the lag
period might not be an important factor in this study.
Some additional sources of uncertainty are not so much inherent in the exposure-response
modeling or in the epidemiologic data themselves but, rather, arise in the process of obtaining
more general Agency risk estimates from the epidemiologic results. U.S. EPA cancer risk
estimates are typically derived to represent an upper bound on increased risk of cancer incidence
for all sites affected by an agent for the general population. From experimental animal studies,
this is accomplished by using tumor incidence data and summing across all the tumor sites that
demonstrate significantly increased incidences, customarily for the most sensitive sex and
species, to attempt to be protective of the general human population. However, in estimating
comparable risks from the Charbotel et al. (2006) epidemiologic data, certain limitations are
encountered. For one thing, these epidemiology data represent a geographically limited (Arve
Valley, France) and likely not very diverse population of working adults. Thus, there is
uncertainty about the applicability of the results to a more diverse general population.
Additionally, the Charbotel et al. (2006) study was a study of RCC only, and so the risk estimate
derived from it does not represent all the tumor sites that may be affected by TCE.
To attempt to account for the potential risk for other cancers associated with TCE
exposure, in particular NHL and liver cancer, for which there were no exposure-response data
available, an adjustment factor reflecting the relative potency of TCE across tumor sites was
derived, using two different approaches. In both approaches, an underlying assumption in
deriving the relative potencies is that the relative values of the age-specific background incidence
risks for the person-years from the epidemiologic studies for each tumor type approximate the
relative values of the lifetime background incidence risks for those tumor types. In other words,
at least on a proportional basis, the lifetime background incidence risks (for the United States
population) for each site approximate the age-specific background incidence risks for the study
populations. A further assumption is that the lifetime risk of RCC up to 85 years is an adequate
approximation to the full lifetime risk, which is what was used for the other two tumor types.
The first calculation, based on the results of the meta-analyses for the three tumor types, has the
advantage of being based on a large data set, incorporating data from many different studies.
However, this calculation relies on a number of additional assumptions. First, it is assumed that
the summary relative risk estimates (RRm's) from the meta-analyses, which are based on
different groups of studies, reflect similar overall TCE exposures, i.e., that the overall TCE
exposures are similar across the different groups of studies that went into the different meta-
analyses for the three tumor types. Second, it is assumed that the RRm's, which incorporate RR
estimates for both mortality and incidence, represent good estimates for cancer incidence risk
from TCE exposure. In addition, it is assumed that the RRm for kidney cancer, for which RCC
estimates from individual studies were used when available, is a good estimate for the overall RR
for RCC and that the RRm estimate for NHL, for which different studies used different
classification schemes, is a good estimate for the overall RR for NHL. The second calculation,
based on the results of the Raaschou-Nielsen et al. (2003) study, the largest single study with RR
estimates for all three tumor types, has the advantage of having RR estimates that are directly
comparable. In addition, the Raaschou-Nielsen et al. study provided data for the precise tumor
types of interest for the calculation, i.e., RCC, NHL, and liver (and biliary) cancer.
36
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The fact that the calculations based on two different data sets yielded comparable values
for the adjustment factor (both within 25% of the selected factor of 4) provides more robust
support for the use of the factor of 4. Additional uncertainties pertain to the weight of evidence
supporting the association of TCE exposure with increased risk of cancer for the three cancer
types. As discussed above, it was found that the weight of evidence for kidney cancer was
sufficient to classify TCE as "carcinogenic to humans." It was also concluded that there was
strong evidence that TCE causes NHL as well, although the evidence for liver cancer was more
limited. In addition, the rodent studies demonstrate clear evidence of multisite carcinogenicity,
with tumor types including those for which associations with TCE exposure are observed in
human studies, i.e., liver and kidney cancers and NHLs. Overall, the evidence was found to be
sufficiently persuasive to support the use of the adjustment factor of 4 based on these three
cancer types. Alternatively, if one were to use the factor based only on the two cancer types with
the strongest human evidence (a factor of 3 for kidney cancer + NHL is suggested by the two
calculations in the table above), the cancer inhalation unit risk estimate would be only slightly
reduced (25%).
Finally, there are uncertainties in the application of ADAFs to adjust for potential
increased early-life susceptibility. The adjustment is made only for the kidney-cancer
component of total cancer risk because that is the tumor type for which the weight of evidence
was sufficient to conclude that TCE-induced carcinogenesis operates through a mutagenic MOA.
However, it may be that TCE operates through a mutagenic MOA for other tumor types as well
or that it operates through other MO As that might also convey increased early-life susceptibility.
Additionally, the ADAFs from the 2005 Supplemental Guidance are not specific to TCE, and it
is uncertain to what extent they reflect increased early-life susceptibility to kidney cancer from
exposure to TCE, if increased early-life susceptibility occurs.
II.D. EPA DOCUMENTATION, REVIEW, AND CONTACTS (CARCINOGENICITY
ASSESSMENT)
II.D.1. EPA DOCUMENTATION
Source Document - U.S. EPA (2011)
This document has been reviewed by EPA scientists, interagency reviewers from other
federal agencies and White House offices, and the public, and peer reviewed by independent
scientists external to EPA. A summary and EPA's disposition of the comments received from
the independent external peer reviewers and from the public is included in Appendix I of the
ToxicologicalReview of Trichloroethylene (U.S. EPA. 2011).
II.D.2. EPA REVIEW
Agency Completion Date — _/_/
II.D.3. EPA CONTACTS
Please contact the IRIS Hotline for all questions concerning this assessment or IRIS, in
37
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general, at (202) 566-1676 (phone), (202) 566-1749 (fax), or hotline.iris@epa.gov (email
address).
III. [reserved]
IV. [reserved]
V. [reserved]
VI. BIBLIOGRAPHY
Substance Name - Trichloroethylene
CASRN-79-01-6
Section VI. Last Revised — 00/00/0000
VI.A. ORAL RfD REFERENCES
Johnson, P.; Goldberg, S.; Mays, M.; Dawson, B. (2003). Threshold of trichloroethylene
contamination in maternal drinking waters affecting fetal heart development in the rat. Environ
HealthPerspect, 111, 289-292. http://www.ncbi.nlm.nih.gov/pubmed/12611656
Keil, D. E.; Peden-Adams, M. M.; Wallace, S.; Ruiz, P.; Gilkeson, G. S. (2009). Assessment of
trichloroethylene (TCE) exposure in murine strains genetically-prone and non-prone to develop
autoimmune disease. J Environ Sci Health A Tox Hazard Subst Environ Eng, 44, 443-453.
http://dx.doi.org/10.1080/109345209Q2719738
NTP. (National Toxicology Program). (1988). Toxicology and carcinogenesis studies of
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VII. REVISION HISTORY
Substance Name - Trichloroethylene
CASRN-79-01-6
File First On-Line 00/00/00
Date Section Description
/ /
_VIII. SYNONYMS
Substance Name - Trichloroethylene
CASRN-79-01-6
Section VIII. Last Revised — 00/00/0000
ACETYLENE TRICHLORIDE
AI3-00052
ALGYLEN
ANAMENTH
BENZINOL
Caswell No 876
CECOLENE
CHLORILEN
1 -CHLORO-2,2-DICHLOROETHYLENE
Chlorylea, Chorylen, CirCosolv, Crawhaspol, Dow-Tri, Dukeron, Per-A-Clor, Triad, Trial, TRI-
Plus M, Vitran
DENSINFLUAT
1,1 -Dichloro-2-chloroethylene
Pesticide Code: 081202
44
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EPA Pesticide Chemical Code 081202
ETHENE, TRICHLORO-
ETHINYL TRICHLORIDE
ETHYLENE TRICHLORIDE
ETHYLENE, TRICHLORO-
FLECK-FLIP
FLOCK FLIP
FLUATE
GERMALGENE
LANADIN
LETHURIN
NARCOGEN
NARKOSOID
NCI-C04546
NIALK
NSC 389
PERM-A-CHLOR
PETZINOL
PHILEX
THRETHYLEN
THRETHYLENE
TRETHYLENE
TRI
TRIASOL
Trichloraethen (German)
Trichloraethylen, tri (German)
TRICHLORAN
TRICHLOREN
Trichlorethene (French)
TRICHLORETHYLENE
Tri chl or ethylene, tri (French)
TRICHLOROETHENE
1,1,2-TRICHLOROETHYLENE
TRICLENE
Tricloretene (Italian)
Tricloroetilene (Italian)
Trielin
Trielina (Italian)
TRIKLONE
TRILENE
TRIMAR
TRI-PLUS
VESTROL
45
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