v/EPA
EPA/635/R-17/016a
External Review Draft
www.epa.gov/iris
Toxicological Review of Ethyl Tertiary Butyl Ether
(CASRN 637-92-3]
June 2017
NOTICE
This document is an External Review Draft. This information is distributed solely for the purpose
of pre-dissemination peer review under applicable information quality guidelines. It has not been
formally disseminated by EPA. It does not represent and should not be construed to represent any
Agency determination or policy. It is being circulated for review of its technical accuracy and
science policy implications.
Integrated Risk Information System
National Center for Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
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Toxicological Review ofETBE
1 DISCLAIMER
2 This document is a preliminary draft for review purposes only. This information is
3 distributed solely for the purpose of pre-dissemination peer review under applicable information
4 quality guidelines. It has not been formally disseminated by EPA. It does not represent and should
5 not be construed to represent any Agency determination or policy. Mention of trade names or
6 commercial products does not constitute endorsement or recommendation for use.
7
This document is a draft for review purposes only and does not constitute Agency policy.
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CONTENTS
AUTHORS | CONTRIBUTORS | REVIEWERS viii
PREFACE x
PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS xiii
EXECUTIVE SUMMARY xxi
LITERATURE SEARCH STRATEGY | STUDY SELECTION AND EVALUATION xxvii
1. HAZARD IDENTIFICATION 1-1
1.1. OVERVIEW OF CHEMICAL PROPERTIES AND TOXICOKINETICS 1-1
1.1.1. Chemical Properties 1-1
1.1.2. Toxicokinetics 1-2
1.1.3. Description of Toxicokinetic Models 1-3
1.1.4. Related Chemicals that Provide Supporting Information 1-3
1.2. PRESENTATION AND SYNTHESIS OF EVIDENCE BY ORGAN/SYSTEM 1-4
1.2.1. Kidney Effects 1-4
1.2.2. Liver Effects 1-36
1.2.3. Reproductive Effects 1-56
1.2.4. Developmental Effects 1-88
1.2.5. Carcinogenicity (Other than in the Kidney or Liver) 1-99
1.2.6. Other Toxicological Effects 1-108
1.3. INTEGRATION AND EVALUATION 1-108
1.3.1. Effects Other Than Cancer 1-108
1.3.2. Carcinogenicity 1-109
1.3.3. Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes 1-113
2. DOSE-RESPONSE ANALYSIS 2-1
2.1. ORAL REFERENCE DOSE FOR EFFECTS OTHER THAN CANCER 2-1
2.1.1. Identification of Studies and Effects for Dose-Response Analysis 2-1
2.1.2. Methods of Analysis 2-3
2.1.3. Derivation of Candidate Values 2-5
2.1.4. Derivation of Organ/System-Specific Reference Doses 2-9
2.1.5. Selection of the Overall Reference Dose 2-10
This document is a draft for review purposes only and does not constitute Agency policy.
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2.1.6. Confidence Statement 2-10
2.1.7. Previous IRIS Assessment 2-10
2.2. INHALATION REFERENCE CONCENTRATION FOR EFFECTS OTHER THAN CANCER 2-11
2.2.1. Identification of Studies and Effects for Dose-Response Analysis 2-11
2.2.2. Methods of Analysis 2-12
2.2.3. Derivation of Candidate Values 2-14
2.2.4. Derivation of Organ/System-Specific Reference Concentrations 2-18
2.2.5. Selection of the Overall Reference Concentration 2-18
2.2.6. Confidence Statement 2-19
2.2.7. Previous IRIS Assessment 2-19
2.2.8. Uncertainties in the Derivation of the Reference Dose and Reference Concentration..2-19
2.3. ORAL SLOPE FACTOR FOR CANCER 2-20
2.3.1. Analysis of Carcinogenicity Data 2-20
2.3.2. Dose-Response Analysis—Adjustments and Extrapolation Methods 2-20
2.3.3. Derivation of the Oral Slope Factor 2-22
2.3.4. Uncertainties in the Derivation of the Oral Slope Factor 2-23
2.3.5. Previous IRIS Assessment: Oral Slope Factor 2-24
2.4. INHALATION UNIT RISK FOR CANCER 2-25
2.4.1. Analysis of Carcinogenicity Data 2-25
2.4.2. Dose-Response Analysis—Adjustments and Extrapolation Methods 2-25
2.4.3. Inhalation Unit Risk Derivation 2-26
2.4.4. Uncertainties in the Derivation of the Inhalation Unit Risk 2-27
2.4.5. Previous IRIS Assessment: Inhalation Unit Risk 2-28
2.5. APPLICATION OF AGE-DEPENDENT ADJUSTMENT FACTORS 2-28
REFERENCES R-l
This document is a draft for review purposes only and does not constitute Agency policy.
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TABLES
Table ES-1. Organ-/system-specific RfDs and overall RfD for ETBE xxiii
Table ES-2. Organ-/system-specific RfCs and overall RfCfor ETBE xxiv
Table LS-1. Details of the search strategy employed for ETBE xxx
Table LS-2. Summary of additional search strategies for ETBE xxxi
Table LS-3. Inclusion-exclusion criteria xxxi
Table LS-4. Considerations for evaluation of experimental animal studies xxxiii
Table LS-5. Summary of experimental animal database xxxiv
Table 1-1. Physicochemical properties and chemical identity of ETBE 1-1
Table 1-2. Evidence pertaining to kidney histopathology effects in animals following exposure to
ETBE 1-10
Table 1-3. Evidence pertaining to kidney biochemistry and urine effects in animals following
exposure to ETBE 1-13
Table 1-4. Evidence pertaining to kidney tumor effects in animals following exposure to ETBE 1-16
Table 1-5. Comparison of nephropathy and urothelial hyperplasia in individual male rats from 2-
year oral exposure (JPEC, 2010a) 1-18
Table 1-6. Comparison of nephropathy and urothelial hyperplasia in individual male rats from 2-
year inhalation exposure (JPEC, 2010b) 1-18
Table 1-7. Additional kidney effects potentially relevant to mode of action in animals exposed to
ETBE 1-24
Table 1-8. Summary of data informing whether the a2u-globulin process is occurring in male rats
exposed to ETBE 1-26
Table 1-9. Evidence pertaining to liver weight effects in animals exposed to ETBE 1-38
Table 1-10. Evidence pertaining to liver histopathology effects in animals exposed to ETBE 1-40
Table 1-11. Evidence pertaining to liver biochemistry effects in animals exposed to ETBE 1-44
Table 1-12. Evidence pertaining to liver tumor effects in animals exposed to ETBE 1-48
Table 1-13. Positive evidence of key characteristics of cancer for ETBE 1-50
Table 1-14. Evidence pertaining to male reproductive effects in animals exposed to ETBE 1-58
Table 1-15. Evidence pertaining to female reproductive effects in animals exposed to ETBE 1-78
Table 1-16. Evidence pertaining to developmental effects in animals following exposure to ETBE 1-91
Table 1-17. Evidence pertaining to ETBE promotion of mutagen-initiated tumors in animals 1-102
Table 1-18. Evidence pertaining to carcinogenic effects (in tissues other than liver or kidney) in
animals exposed to ETBE 1-104
Table 2-1. Summary of derivation of points of departure following oral exposure for up to 2
years 2-4
Table 2-2. Effects and corresponding derivation of candidate values 2-6
Table 2-3. Organ/system-specific RfDs and overall RfD for ETBE 2-10
Table 2-4. Summary of derivation of PODs following inhalation exposure 2-13
Table 2-5. Effects and corresponding derivation of candidate values 2-16
Table 2-6. Organ-/system-specific RfCs and overall RfCfor ETBE 2-18
Table 2-7. Summary of the oral slope factor derivation 2-22
Table 2-8. Summary of uncertainties in the derivation of the oral slope factor for ETBE 2-23
Table 2-9. Summary of the inhalation unit risk derivation 2-27
Table 2-10. Summary of uncertainties in the derivation of the inhalation unit risk for ETBE 2-27
This document is a draft for review purposes only and does not constitute Agency policy.
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FIGURES
Figure LS-1. Summary of literature search and screening process for ETBE xxix
Figure 1-1. Proposed metabolism of ETBE 1-3
Figure 1-2. Comparison of absolute kidney weight change in male and female rats across oral
and inhalation exposure based on internal blood concentration 1-8
Figure 1-3. Comparison of absolute kidney weight change in male and female mice following
inhalation exposure based on administered ETBE concentration 1-9
Figure 1-4. Exposure-response array of kidney effects following oral exposure to ETBE 1-19
Figure 1-5. Exposure-response array of kidney effects following inhalation exposure to ETBE 1-20
Figure 1-6. Temporal pathogenesis of a2u-globulin-associated nephropathy in male rats 1-23
Figure 1-7. ETBE oral exposure array of a2u-globulin data in male rats 1-28
Figure 1-8. ETBE inhalation exposure array of a2u-globulin data in male rats 1-29
Figure 1-9. Exposure-response array of noncancer liver effects following oral exposure to ETBE 1-46
Figure 1-10. Exposure-response array of noncancer liver effects following inhalation exposure to
ETBE 1-47
Figure 1-11. Exposure-response array of male reproductive effects following oral exposure to
ETBE 1-74
Figure 1-12. Exposure-response array of male reproductive effects following inhalation exposure
to ETBE 1-75
Figure 1-13. Exposure-response array of female reproductive effects following oral exposure to
ETBE 1-86
Figure 1-14. Exposure-response array of female reproductive effects following inhalation
exposure to ETBE 1-87
Figure 1-15. Exposure-response array of developmental effects following oral exposure to ETBE 1-98
Figure 1-16. Exposure-response array of carcinogenic effects following oral exposure to ETBE 1-106
Figure 1-17. Exposure-response array of carcinogenic effects following inhalation exposure to
ETBE 1-107
Figure 2-1. Candidate values with corresponding POD and composite UF. Each bar corresponds
to one data set described in Table 2-1 and Table 2-2 2-8
Figure 2-2. Candidate values with corresponding POD and composite UF 2-17
This document is a draft for review purposes only and does not constitute Agency policy.
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i ABBREVIATIONS
ACGIH
American Conference of Governmental
LC50
median lethal concentration
Industrial Hygienists
LD50
median lethal dose
AIC
Akaike's information criterion
LOAEL
lowest-observed-adverse-effect level
ATSDR
Agency for Toxic Substances and
MN
micronuclei
Disease Registry
MNPCE
micronucleated polychromatic
ALP
alkaline phosphatase
erythrocyte
ALT
alanine
MTD
maximum tolerated dose
aminotransferase/transaminase
MTBE
methyl tertiary butyl ether
AST
aspartate
NCEA
National Center for Environmental
aminotransferase/transaminase
Assessment
BMD
benchmark dose
NCI
National Cancer Institute
BMDL
benchmark dose lower confidence limit
NOAEL
no-observed-adverse-effect level
BMDS
Benchmark Dose Software
NTP
National Toxicology Program
BMR
benchmark response
ORD
Office of Research and Development
BUN
blood urea nitrogen
PBPK
physiologically based pharmacokinetic
BW
body weight
PCE
polychromatic erythrocytes
CA
chromosomal aberration
PCNA
proliferating cell nuclear antigen
CASRN
Chemical Abstracts Service Registry
PND
postnatal day
Number
POD
point of departure
CUT
Chemical Industry Institute of
POD [AD J]
duration-adjusted POD
Toxicology
QSAR
quantitative structure-activity
CL
confidence limit
relationship
CNS
central nervous system
RD
relative deviation
CPN
chronic progressive nephropathy
RfC
inhalation reference concentration
CYP450
cytochrome P450
RfD
oral reference dose
DAF
dosimetric adjustment factor
RNA
ribonucleic acid
DNA
deoxyribonucleic acid
SAR
structure activity relationship
EPA
Environmental Protection Agency
SCE
sister chromatid exchange
FDA
Food and Drug Administration
SD
standard deviation
FEVi
forced expiratory volume of 1 second
SE
standard error
GD
gestation day
SGOT
glutamic oxaloacetic transaminase, also
GDH
glutamate dehydrogenase
known as AST
GGT
y-glutamyl transferase
SGPT
glutamic pyruvic transaminase, also
GLP
Good Laboratory Practices
known as ALT
GSH
glutathione
UF
uncertainty factor
GST
glutathione-S-transferase
UFa
animal-to-human uncertainty factor
Hb/g-A
animal blood:gas partition coefficient
UFh
human variation uncertainty factor
Hb/g-H
human blood:gas partition coefficient
UFl
LOAEL-to-NOAEL uncertainty factor
HEC
human equivalent concentration
UFs
subchronic-to-chronic uncertainty
HED
human equivalent dose
factor
i.p.
intraperitoneal
UFd
database deficiencies uncertainty factor
IRIS
Integrated Risk Information System
U.S.
United States
JPEC
Japan Petroleum Energy Center
WT
wild type
KO
Knockout
This document is a draft for review purposes only and does not constitute Agency policy.
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AUTHORS CONTRIBUTORS REVIEWERS
Assessment Team
Keith Salazar, Ph.D. (Chemical
Manager)
U.S. EPA
Office of Research and Development
National Center for Environmental Assessment
Christopher Brinkerhoff, Ph.D.* *Former ORISE Postdoctoral Fellow at U.S.
EPA/ORD/NCEA
Currently with U.S. EPA, Office of Chemical Safety
and Pollution Prevention, Office of Pollution
Prevention and Toxics
Contributors
Andrew Hotchkiss, Ph.D.*
Channa Keshava, Ph.D.*
Jason Fritz, Ph.D.
Janice Lee, Ph.D.*
Christine Cai, M.S.
Alan Sasso, Ph.D.
Paul Schlosser, Ph.D.*
Karen Hogan, M.S.
Vincent Cogliano, Ph.D.
Susan Makris, Ph.D.
Brandy Beverly, Ph.D.*
Erin Yost, Ph.D.*
U.S. EPA
Office of Research and Development
National Center for Environmental Assessment
Washington, DC
*Research Triangle Park, NC
Production Team
Taukecha Cunningham U.S. EPA
Maureen Johnson Office of Research and Development
Terri Konoza National Center for Environmental Assessment
Vicki Soto
Dahnish Shams
Contractor Support
Robyn Blain, Ph.D.
Pam Ross, M.S.P.H.
Ami Gordon, M.P.H.
ICF
9300 Lee Highway
Fairfax, VA
Executive Direction
Kenneth Olden, Ph.D., Sc.D., L.H.D. (former Center Director - Retired) U.S. EPA/ORD/NCEA
Michael Slimak, Ph.D. (Acting Center Director)
John Vandenberg, Ph.D. (National Program Director, HHRA)
Lynn Flowers, Ph.D., DABT (Associate Director for Health, currently with
the Office of Science Policy)
This document is a draft for review purposes only and does not constitute Agency policy.
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Vincent Cogliano, Ph.D. (IRIS Program Director)
Gina Perovich, M.S. (IRIS Program Deputy Director)
Samantha Jones, Ph.D. (IRIS Associate Director for Science)
Jamie Strong, Ph.D. (former Branch Chief, Toxic Effects Branch)
Glinda Cooper, Ph.D. (acting Branch Chief, Toxicity Pathways Branch)
formerly with U.S. EPA
Weihsueh A. Chiu, Ph.D. (Branch Chief, Toxicity Pathways Branch) formerly
with U.S. EPA
Andrew Hotchkiss, Ph.D. (former Acting Branch Chief, Toxicity Pathways
Branch)
Jason Fritz, Ph.D. (acting Branch Chief, Toxicity Pathways Branch)
Jason Lambert, Ph.D., DABT (Acting Branch Chief, Biological Risk
Assessment Branch)
Ted Berner, M.S. (Assistant Center Director)
Karen Hogan, M.S. (former Acting Branch Chief, Toxicity Effects Branch)
Internal Review Team
General Toxicology Workgroup
Inhalation Workgroup
Neurotoxicity Workgroup
PK Workgroup
Reproductive and Developmental
Toxicology Workgroup
Statistical Workgroup
Toxicity Pathways Workgroup
Executive Review Committee
Reviewers
1 This assessment was provided for review to scientists in EPA's Program and Regional Offices.
2 Comments were submitted by:
Office of Children's Health Protection, Washington, DC
Office of Policy, Washington, DC
Office of Solid Waste and Emergency Response, Washington, DC
Office of Air and Radiation, Washington, DC
Region 2, New York City
Region 8, Denver
3 This assessment was provided for review to other federal agencies and the Executive Office of the
4 President Comments were submitted by:
Department of Health and Human Services/Agency for Toxic Substances and Disease Registry,
Department of Health and Human Services/National Institute of Environmental Health
Sciences/National Toxicology Program
Executive Office of the President/Office of Management and Budget
U.S. EPA
Office of Research and Development
National Center for Environmental Assessment
This document is a draft for review purposes only and does not constitute Agency policy.
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PREFACE
This Toxicological Review critically reviews the publicly available studies on ethyl tertiary
butyl ether (ETBE) to identify its adverse health effects and to characterize exposure-response
relationships. The assessment examined all effects by oral and inhalation routes of exposure and
includes an oral noncancer reference dose (RfD), an inhalation noncancer reference concentration
(RfC), a cancer weight of evidence descriptor, and a cancer dose-response assessment It was
prepared under the auspices of the U.S. Environmental Protection Agency's (EPA's) Integrated Risk
Information System (IRIS) program.
This assessment updates a previous IRIS draft assessment of ETBE that went to peer review
in 2010. The previous draft assessment was suspended pending completion of several new studies
that were identified during the peer review and are now included in this document.
The Toxicological Reviews for ETBE and tert-butyl alcohol (tert-butanol) were developed
simultaneously because they have overlapping scientific aspects:
• tert-Butanol and acetaldehyde are the primary metabolites of ETBE, and some of the
toxicological effects ofETBE are attributed to tert- butanol. Therefore, data on tert- butanol
are considered informative for the hazard identification and dose-response assessment of
ETBE, and vice versa.
• The scientific literature for the two chemicals includes data on a2U-globulin-related
nephropathy; therefore, a common approach was used to evaluate the data as they relate to
the mode of action for kidney effects.
• A combined physiologically based pharmacokinetic (PBPK) model for ETBE and tert-
butanol in rats was applied to support the dose-response assessments for these chemicals
fBorghoff etal.. 20161.
Prior to the development of the IRIS assessment, a public meeting was held in December
2013 to obtain input on preliminary materials for ETBE, including draft literature searches and
associated search strategies, evidence tables, and exposure-response arrays. In June 2016, EPA
convened a public science meeting to discuss the public comment draft Toxicological Review of
tert-Butyl Alcohol (tert-butanol) during which time the Agency heard comments on "disentangling
mechanisms of kidney toxicity and carcinogenicity," an issue relevant to both tert-butanol and
ETBE. The complete set of public comments, including the slides presented at the June 2016 public
science meeting, is available on the docket at http://www.regulations.gov (Docket ID No. EPA-HO-
QRD-2013-1111). In October 2016, a public science meeting was held to provide the public an
opportunity to engage in early discussions on the draft IRIS Toxicological Review ofETBE and the
draft charge to the peer review panel prior to release for external peer review. The complete set of
This document is a draft for review purposes only and does not constitute Agency policy.
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public comments, including the slides is available on the docket at http://www.regulations.gov
(DocketID No. EPA-HQ-QRD-2009-02291.
Organ-/system-specific reference values are calculated based on kidney and liver toxicity
data. These reference values could be useful for cumulative risk assessments that consider the
combined effect of multiple agents acting on the same biological system.
This assessment was conducted in accordance with EPA guidance, which is cited and
summarized in the Preamble to IRIS Toxicological Reviews. Appendices for toxicokinetic
information, PBPK modeling, genotoxicity study summaries, dose-response modeling, and other
information are provided as Supplemental Information to this Toxicological Review. For additional
information about this assessment or for general questions regarding IRIS, please contact EPA's
IRIS Hotline at 202-566-1676 (phone), 202-566-1749 (fax), or hotline.iris@epa.gov.
Uses
ETBE has been used as a fuel oxygenate in the United States to improve combustion
efficiency and reduce pollutants in exhaust. From approximately 1990 to 2006, ETBE was
periodically added to gasoline at levels up to approximately 20%, but methyl tert-butyl ether
(MTBE) and other oxygenates were more commonly used. In 2006, use ofETBE and other ether fuel
additives ceased in the United States, and the use of ethanol increased dramatically (Weaver etal..
2010). ETBE is still registered with EPA for use as a fuel additive, but it is not used currently in the
United States. The use of ether fuel additives has been banned or limited by several states, largely in
response to groundwater contamination concerns.
The United States is a major exporter ofETBE, producing 25% of the world's ETBE in 2012.
Worldwide consumption ofETBE is concentrated in Western Europe (~70%). Use in Eastern
Europe and Japan also is relatively high. Japan's use increased dramatically in 2010 to fulfill its
2010 Kyoto Accord obligations (USDA. 2012).
Fate and Transport
ETBE is expected to be highly mobile in soil due to its high carbon-water partitioning
coefficient fHSDB. 20121. ETBE is not predicted to adsorb onto suspended particles and is unlikely
to undergo biodegradation in water fHSDB. 20121. ETBE is estimated to have a half-life of 2 days in
air fHSDB. 20121.
Occurrence in the Environment
ETBE can be released to the environment by gasoline leaks, evaporation, spills, and other
releases. ETBE degrades slowly in the environment and can move with water in soil. Monitoring
studies targeting groundwater near areas where petroleum contamination likely occurred detected
ETBE. For instance, a survey of states reported an average detection rate of 18% for ETBE in
groundwater samples associated with gasoline contamination (NEIWPCC. 2003). Nontargeted
studies, such as a 2006 U.S. Geological Survey (USGS) study (USGS. 2006) measuring volatile
This document is a draft for review purposes only and does not constitute Agency policy.
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organic compounds (VOCs) in general, have lower detection rates. The 2006 USGS study showed
detections ofETBE above 0.2 ng/L in five samples from two public drinking water wells,
corresponding to a 0.0013 rate of detection. The USGS study, which measured several VOCs, was
not targeted to sites that would be most vulnerable to ETBE contamination.
Fuel contamination cleanup is done largely by states, and information on the number of
private contaminated drinking water wells is not consistently available. The State of California
maintains an online database of measurements from contaminated sites fCal/EPA. 20161. From
2010 to 2013, ETBE has been detected in California at 607 and 73 sites in groundwater and air,
respectively. Most of the contamination is attributed to leaking underground storage tanks, and
some contamination is associated with refineries and petroleum transportation. The contamination
was noted in approximately 48 counties, with higher-population counties (e.g., Los Angeles and
Orange) having more contaminated sites.
The occurrence ofETBE in other states was found using fewer and less-standardized data.
Currently, only 13 states routinely analyze for ETBE at fuel-contaminated sites fNEIWPCC. 20031.
Monitoring data associated with leaking storage tanks in Maryland show contamination in
groundwater affecting multiple properties fMarvland Department of the Environment. 20161.
General Population Exposure
ETBE exposure can occur in many different settings. Releases from underground storage
tanks could result in exposure to individuals who obtain their drinking water from wells. Due to its
environmental mobility and resistance to biodegradation, ETBE has the potential to contaminate
and persist in groundwater and soil fHSDB. 20121: therefore, exposure through ingestion of
contaminated drinking water is possible.
Other human exposure pathways ofETBE include inhalation and, to a lesser extent, dermal
contact ETBE inhalation exposure can occur due the chemical's volatility and release from
industrial processes and contaminated sites fHSDB. 20121.
Assessments by Other National and International Health Agencies
Toxicity information on ETBE has been evaluated by the National Institute for Public Health
and the Environment (Bilthoven, The Netherlands) fTiesiema and Baars. 20091. The results of this
assessment are presented in Appendix A of the Supplemental Information to this Toxicological
Review. Of importance to recognize is that this earlier assessment could have been prepared for
different purposes and might use different methods. In addition, newer studies have been included
in the IRIS assessment.
The International Agency for Research on Cancer (IARC) may evaluate ETBE within the next
few years fStraif etal.. 20141.
This document is a draft for review purposes only and does not constitute Agency policy.
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PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS
Note: The Preamble summarizes the
objectives and scope of the IRIS program,
general principles and systematic review
procedures used in developing IRIS
assessments, and the overall development
process and document structure.
1. Objectives and Scope of the IRIS
Program
Soon after EPA was established in 1970, it
was at the forefront of developing risk
assessment as a science and applying it in
support of actions to protect human health
and the environment EPA's IRIS program1
contributes to this endeavor by reviewing
epidemiologic and experimental studies of
chemicals in the environment to identify
adverse health effects and characterize
exposure-response relationships. Health
agencies worldwide use IRIS assessments,
which are also a scientific resource for
researchers and the public.
IRIS assessments cover the hazard
identification and dose-response steps of
risk assessment Exposure assessment and
risk characterization are outside the scope of
IRIS assessments, as are political, economic,
and technical aspects of risk management An
IRIS assessment may cover one chemical, a
group of structurally or toxicologically
related chemicals, or a chemical mixture.
Exceptions outside the scope of the IRIS
program are radionuclides, chemicals used
only as pesticides, and the "criteria air
pollutants" (particulate matter, ground-level
37 ozone, carbon monoxide, sulfur oxides,
38 nitrogen oxides, and lead).
39 Enhancements to the IRIS program are
40 improving its science, transparency, and
41 productivity. To improve the science, the IRIS
42 program is adapting and implementing
43 principles of systematic review (i.e., using
44 explicit methods to identify, evaluate, and
45 synthesize study findings). To increase
46 transparency, the IRIS program discusses key
47 science issues with the scientific community
48 and the public as it begins an assessment.
49 External peer review, independently
50 managed and in public, improves both
51 science and transparency. Increased
52 productivity requires that assessments be
53 concise, focused on EPA's needs, and
54 completed without undue delay.
55 IRIS assessments follow EPA guidance2
56 and standardized practices of systematic
57 review. This Preamble summarizes and does
58 not change IRIS operating procedures or EPA
59 guidance.
60 Periodically, the IRIS program asks for
61 nomination of agents for future assessment
62 or reassessment. Selection depends on EPA's
63 priorities, relevance to public health, and
64 availability of pertinent studies. The IRIS
65 multiyear agenda3 lists upcoming
66 assessments. The IRIS program may also
67 assess other agents in anticipation of public
68 health needs.
1 IRIS program website: http: //www.epa.gov/iris/
2 EPA guidance documents: http://www.epa.gov/iris/basic-information-about-integrated-risk-information-
svstem#guidance/
3 IRIS multiyear agenda: https: //www.epa.gov/iris/iris-agenda
This document is a draft for review purposes only and does not constitute Agency policy.
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2. Planning an Assessment:
Scoping, Problem Formulation,
and Protocols
Early attention to planning ensures that
IRIS assessments meet their objectives and
properly frame science issues.
Scoping refers to the first step of
planning, where the IRIS program consults
with EPA's program and regional offices to
ascertain their needs. Scoping specifies the
agents an assessment will address, routes
and durations of exposure, susceptible
populations and lifestages, and other topics of
interest
Problem formulation refers to the
science issues an assessment will address
and includes input from the scientific
community and the public. A preliminary
literature survey, beginning with secondary
sources (e.g., assessments by national and
international health agencies and
comprehensive review articles), identifies
potential health outcomes and science issues.
It also identifies related chemicals (e.g.,
toxicologically active metabolites and
compounds that metabolize to the chemical
of interest).
Each IRIS assessment comprises multiple
systematic reviews for multiple health
outcomes. It also evaluates hypothesized
mechanistic pathways and characterizes
exposure-response relationships. An
assessment may focus on important health
outcomes and analyses rather than expand
beyond what is necessary to meet its
objectives.
Protocols refer to the systematic review
procedures planned for use in an assessment
They include strategies for literature
searches, criteria for study inclusion or
exclusion, considerations for evaluating
study methods and quality, and approaches
to extracting data. Protocols may evolve as an
Toxicological Review ofETBE
44 assessment progresses and new agent-
45 specific insights and issues emerge.
46
47 3. Identifying and Selecting
48 Pertinent Studies
49 IRIS assessments conduct systematic
50 literature searches with criteria for inclusion
51 and exclusion. The objective is to retrieve the
52 pertinent primary studies (i.e., studies with
53 original data on health outcomes or their
54 mechanisms). PECO statements (Populations,
55 Exposures, Comparisons, Outcomes) govern
56 the literature searches and screening criteria.
57 "Populations" and animal species generally
58 have no restrictions. "Exposures" refers to
59 the agent and related chemicals identified
60 during scoping and problem formulation and
61 may consider route, duration, or timing of
62 exposure. "Comparisons" means studies that
63 allow comparison of effects across different
64 levels of exposure. "Outcomes" may become
65 more specific (e.g., from "toxicity" to
66 "developmental toxicity" to "hypospadias")
67 as an assessment progresses.
68 For studies of absorption, distribution,
69 metabolism, and elimination, the first
70 objective is to create an inventory of
71 pertinent studies. Subsequent sorting and
72 analysis facilitates characterization and
73 quantification of these processes.
74 Studies on mechanistic events can be
75 numerous and diverse. Here, too, the
76 objective is to create an inventory of studies
77 for later sorting to support analyses of related
78 data. The inventory also facilitates generation
79 and evaluation of hypothesized mechanistic
80 pathways.
81 The IRIS program posts initial protocols
82 for literature searches on its website and
83 adds search results to EPA's HERO database.4
84 Then the IRIS program takes extra steps to
85 ensure identification of pertinent studies: by
4 Health and Environmental Research Online: https: //hero.epa.gov/hero/
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
encouraging the scientific community and the
public to identify additional studies and
ongoing research; by searching for data
submitted under the Toxic Substances
Control Act or the Federal Insecticide,
Fungicide, and Rodenticide Act; and by
considering late-breaking studies that would
impact the credibility of the conclusions, even
during the review process.5
4. Evaluating Study Methods and
Quality
IRIS assessments evaluate study methods
and quality, using uniform approaches for
each group of similar studies. The objective is
that subsequent syntheses can weigh study
results on their merits. Key concerns are
potential bias (factors that affect the
magnitude or direction of an effect) and
insensitivity (factors that limit the ability of a
study to detect a true effect).
For human and animal studies, the
evaluation of study methods and quality
considers study design, exposure measures,
outcome measures, data analysis, selective
reporting, and study sensitivity. For human
studies, this evaluation also considers
selection of participant and referent groups
and potential confounding. Emphasis is on
discerning bias that could substantively
change an effect estimate, considering also
the expected direction of the bias. Low
sensitivity is a bias towards the null.
Study-evaluation considerations are
specific to each study design, health effect,
and agent. Subject-matter experts evaluate
each group of studies to identify
characteristics that bear on the
informativeness of the results. For
carcinogenicity, neurotoxicity, reproductive
toxicity, and developmental toxicity, there is
EPA guidance for study evaluation (U.S. EPA.
2005a. 1998b. 1996. 1991bl. As subject-
matter experts examine a group of studies,
44 additional agent-specific knowledge or
45 methodologic concerns may emerge and a
46 second pass become necessary.
47 Assessments use evidence tables to
48 summarize the design and results of
49 pertinent studies. If tables become too
50 numerous or unwieldy, they may focus on
51 effects that are more important or studies
52 that are more informative.
53 The IRIS program posts initial protocols
54 for study evaluation on its website, then
55 considers public input as it completes this
56 step.
57 5. Integrating the Evidence of
58 Causation for Each Health
59 Outcome
60 Synthesis within lines of evidence. For
61 each health outcome, IRIS assessments
62 synthesize the human evidence and the
63 animal evidence, augmenting each with
64 informative subsets of mechanistic data. Each
65 synthesis considers aspects of an association
66 that may suggest causation: consistency,
67 exposure-response relationship, strength of
68 association, temporal relationship, biological
69 plausibility, coherence, and "natural
70 experiments" in humans (U.S. EPA. 1994)
71 fU.S. EPA. 2005al
72 Each synthesis seeks to reconcile
73 ostensible inconsistencies between studies,
74 taking into account differences in study
75 methods and quality. This leads to a
76 distinction between conflicting evidence
77 (unexplained positive and negative results in
78 similarly exposed human populations or in
79 similar animal models) and differing results
80 (mixed results attributable to differences
81 between human populations, animal models,
82 or exposure conditions) (U.S. EPA. 2005a).
83 Each synthesis of human evidence
84 explores alternative explanations (e.g.,
85 chance, bias, or confounding) and determines
86 whether they may satisfactorily explain the
5 IRIS "stopping rules": https: //www.epa.gov/sites/production/files/2014-06/documents/
iris stoppingrules.pdf
This document is a draft for review purposes only and does not constitute Agency policy.
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results. Each synthesis of animal evidence
explores the potential for analogous results in
humans. Coherent results across multiple
species increase confidence that the animal
results are relevant to humans.
Mechanistic data are useful to augment
the human or animal evidence with
information on precursor events, to evaluate
the human relevance of animal results, or to
identify susceptible populations and
lifestages. An agent may operate through
multiple mechanistic pathways, even if one
hypothesis dominates the literature fU.S.
EPA. 2005a").
Integration across lines of evidence.
For each health outcome, IRIS assessments
integrate the human, animal, and mechanistic
evidence to answer the question: What is the
nature of the association between exposure to
the agent and the health outcome?
For cancer, EPA includes a standardized
hazard descriptor in characterizing the
strength of the evidence of causation. The
objective is to promote clarity and
consistency of conclusions across
assessments fU.S. EPA. 2005al.
Carcinogenic to humans: convincing
epidemiologic evidence of a causal
association; or strong human evidence of
cancer or its key precursors, extensive animal
evidence, identification of mode-of-action
and its key precursors in animals, and strong
evidence that they are anticipated in humans.
Likely to be carcinogenic to humans:
evidence that demonstrates a potential
hazard to humans. Examples include a
plausible association in humans with
supporting experimental evidence, multiple
positive results in animals, a rare animal
response, or a positive study strengthened by
other lines of evidence.
Suggestive evidence of carcinogenic
potential: evidence that raises a concern for
humans. Examples include a positive result in
the only study, or a single positive result in an
extensive database.
Inadequate information to assess
carcinogenic potential: no other descriptors
apply. Examples include little or no pertinent
50 information, conflicting evidence, or negative
51 results not sufficiently robust for not likely.
52 Not likely to be carcinogenic to humans:
53 robust evidence to conclude that there is no
54 basis for concern. Examples include no effects
55 in well-conducted studies in both sexes of
56 multiple animal species, extensive evidence
57 showing that effects in animals arise through
58 modes-of-action that do not operate in
59 humans, or convincing evidence that effects
60 are not likely by a particular exposure route
61 or below a defined dose.
62 If there is credible evidence of
63 carcinogenicity, there is an evaluation of
64 mutagenicity, because this influences the
65 approach to dose-response assessment and
66 subsequent application of adjustment factors
67 for exposures early in life fU.S. EPA. 2005al.
68 riJ.S. EPA. 2005b").
69 6. Selecting Studies for Derivation
70 of Toxicity Values
71 The purpose of toxicity values (slope
72 factors, unit risks, reference doses, reference
73 concentrations; see section 7) is to estimate
74 exposure levels likely to be without
75 appreciable risk of adverse health effects.
76 EPA uses these values to support its actions
77 to protect human health.
78 The health outcomes considered for
79 derivation of toxicity values may depend on
80 the hazard descriptors. For example, IRIS
81 assessments generally derive cancer values
82 for agents that are carcinogenic or likely to be
83 carcinogenic, and sometimes for agents with
84 suggestive evidence fU.S. EPA. 2005al.
85 Derivation of toxicity values begins with a
86 new evaluation of studies, as some studies
87 used qualitatively for hazard identification
88 may not be useful quantitatively for
89 exposure-response assessment Quantitative
90 analyses require quantitative measures of
91 exposure and response. An assessment
92 weighs the merits of the human and animal
93 studies, of various animal models, and of
94 different routes and durations of exposure
95 fU.S. EPA. 19941. Study selection is not
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
reducible to a formula, and each assessment
explains its approach.
Other biological determinants of study
quality include appropriate measures of
exposure and response, investigation of early
effects that precede overt toxicity, and
appropriate reporting of related effects (e.g.,
combining effects that comprise a syndrome,
or benign and malignant tumors in a specific
tissue).
Statistical determinants of study quality
include multiple levels of exposure (to
characterize the shape of the exposure-
response curve) and adequate exposure
range and sample sizes (to minimize
extrapolation and maximize precision) (U.S.
EPA. 20121.
Studies of low sensitivity may be less
useful if they fail to detect a true effect or
yield toxicity values with wide confidence
limits.
7. Deriving Toxicity Values
General approach. EPA guidance
describes a two-step approach to dose-
response assessment: analysis in the range of
observation, then extrapolation to lower
levels. Each toxicity value pertains to a route
(e.g., oral, inhalation, dermal) and duration or
timing of exposure (e.g., chronic, subchronic,
gestational) fU.S. EPA. 20021.
IRIS assessments derive a candidate
value from each suitable data set
Consideration of candidate values yields a
toxicity value for each organ or system.
Consideration of the organ/system-specific
values results in the selection of an overall
toxicity value to cover all health outcomes.
The organ/system-specific values are useful
for subsequent cumulative risk assessments
that consider the combined effect of multiple
agents acting at a common anatomical site.
Analysis in the range of observation.
Within the observed range, the preferred
approach is modeling to incorporate a wide
range of data. Toxicokinetic modeling has
become increasingly common for its ability to
support target-dose estimation, cross-species
adjustment, or exposure-route conversion. If
data are too limited to support toxicokinetic
modeling, there are standardized approaches
to estimate daily exposures and scale them
from animals to humans fU.S. EPA. 19941.
("U.S. EPA. 2005al. fU.S. EPA. 2011. 20061.
For human studies, an assessment may
develop exposure-response models that
reflect the structure of the available data fU.S.
EPA. 2005a). For animal studies, EPA has
developed a set of empirical ("curve-fitting")
models6 that can fit typical data sets fU.S.
EPA. 2005a). Such modeling yields a point of
departure, defined as a dose near the lower
end of the observed range, without significant
extrapolation to lower levels (e.g., the
estimated dose associated with an extra risk
of 10% for animal data or 1% for human data,
or their 95% lower confidence limits) fU.S.
EPA. 2005al. fU.S. EPA. 20121.
When justified by the scope of the
assessment, toxicodynamic ("biologically
based") modeling is possible if data are
sufficient to ascertain the key events of a
mode-of-action and to estimate their
parameters. Analysis of model uncertainty
can determine the range of lower doses
where data support further use of the model
CU.S. EPA. 2005al.
For a group of agents that act at a
common site or through common
mechanisms, an assessment may derive
relative potency factors based on relative
toxicity, rates of absorption or metabolism,
quantitative structure-activity relationships,
or receptor-binding characteristics (U.S. EPA.
2005a).
Extrapolation: slope factors and unit
risks. An oral slope factor or an inhalation
unit risk facilitates subsequent estimation of
human cancer risks. Extrapolation proceeds
linearly (i.e., risk proportional to dose) from
the point of departure to the levels of interest
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6 Benchmark Dose Software: http: //www.epa.gov/bmds/
This document is a draft for review purposes only and does not constitute Agency policy.
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This is appropriate for agents with direct
mutagenic activity. It is also the default if
there is no established mode-of-action fU.S.
EPA. 2005a").
Differences in susceptibility may warrant
derivation of multiple slope factors or unit
risks. For early-life exposure to carcinogens
with a mutagenic mode-of-action, EPA has
developed default age-dependent adjustment
factors for agents without chemical-specific
susceptibility data fU.S. EPA. 2005al. fU.S.
EPA. 2005b").
If data are sufficient to ascertain the
mode-of-action and to conclude that it is not
linear at low levels, extrapolation may use the
reference-value approach (U.S. EPA. 2005a).
Extrapolation: reference values. An
oral reference dose or an inhalation reference
concentration is an estimate of human
exposure (including in susceptible
populations) likely to be without appreciable
risk of adverse health effects over a lifetime
fU.S. EPA. 20021. Reference values generally
cover effects other than cancer. They are also
appropriate for carcinogens with a nonlinear
mode-of-action.
Calculation of reference values involves
dividing the point of departure by a set of
uncertainty factors (each typically 1, 3, or 10,
unless there are adequate chemical-specific
data) to account for different sources of
uncertainty and variability (U.S. EPA. 2002).
riJ.S. EPA. 20141
Human variation: An uncertainty factor
covers susceptible populations and lifestages
that may respond at lower levels, unless the
data originate from a susceptible study
population.
Animal-to-human extrapolation: For
reference values based on animal results, an
uncertainty factor reflects cross-species
differences, which may cause humans to
respond at lower levels.
Subchronic-to-chronic exposure: For
chronic reference values based on subchronic
studies, an uncertainty factor reflects the
likelihood that a lower level over a longer
duration may induce a similar response. This
49 factor may not be necessary for reference
50 values of shorter duration.
51 Adverse-effect level to no-observed-
52 adverse-effect level: For reference values
53 based on a lowest-observed-adverse-effect
54 level, an uncertainty factor reflects a level
55 judged to have no observable adverse effects.
56 Database deficiencies: If there is concern
57 that future studies may identify a more
58 sensitive effect, target organ, population, or
59 lifestage, a database uncertainty factor
60 reflects the nature of the database deficiency.
61 8. Process for Developing and Peer-
62 Reviewing IRIS Assessments
63 The IRIS process (revised in 2009 and
64 enhanced in 2013) involves extensive public
65 engagement and multiple levels of scientific
66 review and comment. IRIS program scientists
67 consider all comments. Materials released,
68 comments received from outside EPA, and
69 disposition of major comments (steps 3, 4,
70 and 6 below) become part of the public
71 record.
72 Step 1: Draft development. As outlined
73 in section 2 of this Preamble, IRIS program
74 scientists specify the scope of an assessment
75 and formulate science issues for discussion
76 with the scientific community and the public.
77 Next, they release initial protocols for the
78 systematic review procedures planned for
79 use in the assessment. IRIS program
80 scientists then develop a first draft, using
81 structured approaches to identify pertinent
82 studies, evaluate study methods and quality,
83 integrate the evidence of causation for each
84 health outcome, select studies for derivation
85 of toxicity values, and derive toxicity values,
86 as outlined in Preamble sections 3-7.
87 Step 2: Agency review. Health scientists
88 across EPA review the draft assessment.
89 Step 3: Interagency science
90 consultation. Other federal agencies and the
91 Executive Office of the President review the
92 draft assessment
93 Step 4: Public comment, followed by
94 external peer review. The public reviews
95 the draft assessment. IRIS program scientists
This document is a draft for review purposes only and does not constitute Agency policy.
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release a revised draft for independent
external peer review. The peer reviewers
consider whether the draft assessment
assembled and evaluated the evidence
according to EPA guidance and whether the
evidence justifies the conclusions.
Step 5: Revise assessment. IRIS
program scientists revise the assessment to
address the comments from the peer review.
Step 6: Final agency review and
interagency science discussion. The IRIS
program discusses the revised assessment
with EPA's program and regional offices and
with other federal agencies and the Executive
Office of the President
Step 7: Post final assessment. The IRIS
program posts the completed assessment
and a summary on its website.
9. General Structure of IRIS
Assessments
Main text. IRIS assessments generally
comprise two major sections: (1) Hazard
Identification and (2) Dose-Response
Assessment Section 1.1 briefly reviews
chemical properties and toxicokinetics to
describe the disposition of the agent in the
body. This section identifies related
chemicals and summarizes their health
outcomes, citing authoritative reviews. If an
assessment covers a chemical mixture, this
section discusses environmental processes
that alter the mixtures humans encounter
and compares them to mixtures studied
experimentally.
Section 1.2 includes a subsection for each
major health outcome. Each subsection
discusses the respective literature searches
and study considerations, as outlined in
Preamble sections 3 and 4, unless covered in
the front matter. Each subsection concludes
with evidence synthesis and integration, as
outlined in Preamble section 5.
Section 1.3 links health hazard
information to dose-response analyses for
each health outcome. One subsection
identifies susceptible populations and
lifestages, as observed in human or animal
48 studies or inferred from mechanistic data.
49 These may warrant further analysis to
50 quantify differences in susceptibility.
51 Another subsection identifies biological
52 considerations for selecting health outcomes,
53 studies, or data sets, as outlined in Preamble
54 section 6.
55 Section 2 includes a subsection for each
56 toxicity value. Each subsection discusses
57 study selection, methods of analysis, and
58 derivation of a toxicity value, as outlined in
59 Preamble sections 6 and 7.
60 Front matter. The Executive Summary
61 provides information historically included in
62 IRIS summaries on the IRIS program website.
63 Its structure reflects the needs and
64 expectations of EPA's program and regional
65 offices.
66 A section on systematic review methods
67 summarizes key elements of the protocols,
68 including methods to identify and evaluate
69 pertinent studies. The final protocols appear
70 as an appendix.
71 The Preface specifies the scope of an
72 assessment and its relation to prior
73 assessments. It discusses issues that arose
74 during assessment development and
75 emerging areas of concern.
76 This Preamble summarizes general
77 procedures for assessments begun after the
78 date below. The Preface identifies
79 assessment-specific approaches that differ
80 from these general procedures.
81 August 2016
82
83 10. Preamble References
84 U.S. EPA. (1991). Guidelines for
85 developmental toxicity risk assessment (pp.
86 1-83). (EPA/600/FR-91/001). Washington,
87 DC: U.S. Environmental Protection Agency,
88 Risk Assessment Forum.
89 http://cfpub.epa.gov/ncea/cfm/recordispla
90 v. cfm?deid=2 316 2
91 U.S. EPA. (1994). Methods for derivation of
92 inhalation reference concentrations and
This document is a draft for review purposes only and does not constitute Agency policy.
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application of inhalation dosimetry [EPA
Report] (pp. 1-409). (EPA/600/8-90/066F).
Research Triangle Park, NC: U.S.
Environmental Protection Agency, Office of
Research and Development, Office of Health
and Environmental Assessment,
Environmental Criteria and Assessment
Office.
https://cfpub.epa.gov/ncea/risk/recordispl
av.cfm?deid=71993&CFID=51174829&CFTO
KEN=25006317
U.S. EPA. (1996). Guidelines for reproductive
toxicity risk assessment (pp. 1-143).
(EPA/630/R-96/009). Washington, DC: U.S.
Environmental Protection Agency, Risk
Assessment Forum.
U.S. EPA. (1998). Guidelines for neurotoxicity
risk assessment Fed Reg 63: 26926-26954.
U.S. EPA. (2002). A review of the reference
dose and reference concentration processes
(pp. 1-192). (EPA/630/P-02/002F).
Washington, DC: U.S. Environmental
Protection Agency, Risk Assessment Forum.
http://www.epa.gov/osa/review-reference-
dose-and-reference-concentration-processes
U.S. EPA. (2005a). Guidelines for carcinogen
risk assessment [EPA Report] (pp. 1-166).
(EPA/630/P-03/001F). Washington, DC: U.S.
Environmental Protection Agency, Risk
Assessment Forum.
http://www2.epa.gov/osa/guidelines-
carcinogen-risk-assessment
U.S. EPA. (2005b). Supplemental guidance for
assessing susceptibility from early-life
exposure to carcinogens (pp. 1-125).
(EPA/630/R-03/003F). Washington, DC: U.S.
Environmental Protection Agency, Risk
Assessment Forum.
U.S. EPA. (2006). Approaches for the
application of physiologically based
pharmacokinetic (PBPK) models and
supporting data in risk assessment (Final
Report) [EPA Report] (pp. 1-123).
(EPA/600/R-05/043F). Washington, DC: U.S.
Environmental Protection Agency, Office of
Research and Development, National Center
for Environmental Assessment
http://cfpub.epa.gov/ncea/cfm/recordispla
v.cfm?deid=157668
50 U.S. EPA. (2011). Recommended use of body
51 weight 3/4 as the default method in
52 derivation of the oral reference dose (pp. 1-
53 50). (EPA/100/R11/0001). Washington, DC:
54 U.S. Environmental Protection Agency, Risk
55 Assessment Forum, Office of the Science
56 Advisor.
57 https://www.epa.gov/risk/recommended-
58 use-body-weight-34-default-method-
59 derivation-oral-reference-dose
60 U.S. EPA. (2012). Benchmark dose technical
61 guidance (pp. 1-99). (EPA/100/R-12/001).
62 Washington, DC: U.S. Environmental
63 Protection Agency, Risk Assessment Forum.
64 U.S. EPA. (2014). Guidance for applying
65 quantitative data to develop data-derived
66 extrapolation factors for interspecies and
67 intraspecies extrapolation. (EPA/100/R-
68 14/002F). Washington, DC: Risk Assessment
69 Forum, Office of the Science Advisor.
70 http://www.epa.gov/raf/DDEF/pdf/ddef-
71 final.pdf
72
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EXECUTIVE SUMMARY
Summation of Occurrence and Health Effects
Ethyl tert-butyl ether (ETBE) does not occur naturally; it is an ether oxygenate
produced by humans and primarily used as a gasoline additive. It was used until 2006
in the United States, and is still used in Japan and the European Union. ETBE is
released into the environment because of gasoline leaks, evaporation, and spills.
Exposure to ETBE can occur by drinking contaminated groundwater or by inhaling
off-gases containing ETBE. Dermal exposure is possible in occupational settings
where the manufacture ofETBE occurs. The magnitude of human exposure to ETBE
depends on factors such as the distribution ofETBE in groundwater and the extent of
the contamination.
Animal studies demonstrate that exposure to ETBE is associated with noncancer
kidney effects. Evidence is suggestive of carcinogenic potential for ETBE based on
liver tumors in rats. Studies in animals indicate that deficient clearance of
acetaldehyde, a metabolite of ETBE, could increase susceptibility to ETBE toxicity or
carcinogenicity.
Effects Other Than Cancer Observed Following Oral Exposure
No human studies are available to evaluate the effects of oral exposure. Kidney effects were
identified as a potential human hazard ofETBE exposure, with increased kidney weight in male and
female rats accompanied by increased chronic progressive nephropathy (CPN), urothelial
hyperplasia (in males), and increased blood concentrations of total cholesterol, blood urea nitrogen
(BUN), and creatinine. Overall, there was consistency across multiple measures of potential kidney
toxicity, including organ weight increases, exacerbated CPN, urothelial hyperplasia, and increases in
serum markers of kidney function. Additionally, effects were consistently observed across routes of
exposure, species, and sex; however, male rats appeared more sensitive to exposure than female
rats, and rats seemed to be more sensitive to exposure than mice. A mode of action (MOA) analysis
determined that the data were insufficient to conclude that kidney effects in male rats were
mediated by a2U-globulin-associated nephropathy alone. CPN and the exacerbation of CPN play a
role in renal tubule nephropathy, although CPN is unlikely to be associated with urothelial
hyperplasia. Changes in absolute kidney weights, urothelial hyperplasia, and increased blood
biomarkers are considered to result from ETBE exposure and are appropriate for identifying a
hazard to the kidney.
Evidence is suggestive that liver toxicity follows ETBE exposure. The strongest supporting
evidence is the increased liver weights and centrilobular hypertrophy in exposed male and female
rats consistently reported across studies evaluating both oral and inhalation exposures. No
additional histopathological findings were observed, however, and only one serum marker of liver
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toxicity [gamma-glutamyl transferase (GGT)] was elevated, while other markers [aspartate
aminotransferase (AST), alanine aminotransferase (ALT), and alkaline phosphatase (ALP)] were
unchanged. The magnitude of change for these noncancer effects was mild to moderate and, except
for organ weight data, did not exhibit consistent dose-response relationships. Mechanistic data
suggest that ETBE exposure leads to activation of several nuclear receptors, but inadequate
evidence exists to establish a relationship between receptor activation and liver toxicity resulting
from ETBE exposure. In addition, mechanistic data suggest possibly greater susceptibility of toxic
effects related to reduced clearance of acetaldehyde, a metabolite ofETBE. Thus, even with the
consistently observed increases in rat liver weight and centrilobular hypertrophy, the evidence
remains suggestive that liver toxicity follows ETBE exposure.
Inadequate information exists to draw conclusions regarding male reproductive effects,
female reproductive effects developmental effects, changes in body weight, adrenal function,
immune status or mortality.
Oral Reference Dose (RfD) for Effects Other Than Cancer
Kidney toxicity, represented by urothelial hyperplasia, was chosen as the basis for the
overall oral reference dose (RfD) (See Table ES-1). The chronic study by (TPEC. 2010a) [selected
data published as Suzuki etal. (2012)] and the observed kidney effects were used to derive the RfD.
The endpoint of urothelial hyperplasia was selected as the critical effect because it is a specific and
sensitive indicator of kidney toxicity and was induced in a dose-responsive manner. Benchmark
dose (BMD) modeling was used to derive the benchmark dose lower confidence limit (BMDLioo/0) of
60.5 mg/kg-day. The BMDL was converted to a human equivalent dose (HED) of 14.5 mg/kg-day
using body weight3/4 scaling, and this value was used as the point of departure (POD) for RfD
derivation (U.S. EPA. 2011).
The overall RfD was calculated by dividing the POD for increased urothelial hyperplasia by a
composite uncertainty factor (UF) of 30 to account for extrapolation from animals to humans (3)
and interindividual differences in human susceptibility (10).
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Table ES-1. Organ-/system-specific RfDs and overall RfD for ETBE
Hazard
Basis
Point of
departure*
(mg/kg-day)
UF
Chronic RfD
(mg/kg-day)
Study
exposure
description
Confidence
Kidney
Urothelial
hyperplasia
14.5
30
5 x 10 1
Chronic
High
Overall
RfD
Kidney
14.5
30
5 x 101
Chronic
High
* Human equivalent dose (HED) PODs were calculated using body weight to the % power (BW3/4) scaling (U.S. EPA,
2011).
Effects Other Than Cancer Observed Following Inhalation Exposure
No human studies are available to evaluate the effects of inhalation exposure. Kidney effects
are a potential human hazard of inhalation exposure to ETBE. Increases in kidney weight,
nephropathy, mineralization, urothelial hyperplasia, and blood concentration of cholesterol, BUN,
and creatinine were observed in male or female rats following 13 weeks of inhalation exposure or
longer. In these studies, changes in serum biomarkers lacked consistency and strength of
association. Changes in rat kidney weight and urothelial hyperplasia, however, were consistent
findings across multiple studies, and are considered a result of ETBE exposure and appropriate for
identifying a hazard to the kidney.
Inhalation Reference Concentration (RfC) for Effects Other Than Cancer
Kidney toxicity, represented by urothelial hyperplasia, was chosen as the basis for the
overall inhalation reference concentration (RfC) (See Table ES-2). The chronic study by 1PEC
f2010bl [selected data published as Saito etal. f20131] and the observed kidney effects were used
to derive the RfC. The endpoint, urothelial hyperplasia, was selected as the critical effect because it
is a specific and sensitive indicator of kidney toxicity and was induced in a dose-responsive manner.
Benchmark dose (BMD) modeling was used to derive the BMCLioo/0 of 1,498 mg/m3. The BMCL was
adjusted to a continuous exposure and converted to a human equivalent concentration (HEC) of
265 mg/m3.
The overall RfC was calculated by dividing the POD by a composite UF of 30 to account for
toxicodynamic differences between animals and humans (3) and interindividual differences in
human susceptibility (10).
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Table ES-2. Organ-/system-specific RfCs and overall RfC for ETBE
Hazard
Basis
Point of
departure*
(mg/m3)
UF
Chronic RfC
(mg/m3)
Study exposure
description
Confidence
Kidney
Urothelial
hyperplasia
265
30
9x 10°
Chronic
High
Overall RfC
Kidney
265
30
9x10°
Chronic
High
*Continuous inhalation HEC was adjusted for continuous daily exposure and calculated by adjusting the duration-
adjusted POD (PODadj) by the dosimetric adjustment factor (DAF = 0.992) for a Category 3 gas.
Evidence of Human Carcinogenicity
Under EPA's cancer guidelines (U.S. EPA. 2005a). there is suggestive evidence of carcinogenic
potential for ETBE. ETBE induced liver tumors in male (but not female) rats in a 2-year inhalation
exposure study, and increased mutagen-initiated liver, thyroid, colon, urinary bladder, and kidney
tumor incidence in 2-stage oral carcinogenesis bioassays. The potential for carcinogenicity applies
to all routes of human exposure.
Quantitative Estimate of Carcinogenic Risk from Oral Exposure
A quantitative estimate of carcinogenic potential from oral exposure to ETBE was based on
the increased incidence of hepatocellular adenomas and carcinomas in male F344 rats following
2-year inhalation exposure fSaito etal.. 2013: TPEC. 2010b). The study included histological
examinations for tumors in many different tissues, contained three exposure levels and controls,
contained adequate numbers of animals per dose group (~50/sex/group), treated animals for up to
2 years, and included detailed reporting of methods and results.
Although ETBE was considered to have "suggestive evidence of carcinogenic potential," the
main study fSaito etal.. 2013: TPEC. 2010b) was well conducted and suitable for quantitative
analyses. A PBPK model in rats for ETBE and its metabolite, tert-butanol, was used for route-to-
route extrapolation of the inhalation BMCLio (described below) to an oral equivalent BMDLio,
which was adjusted to a human equivalent BMDLio based on body weight3/4 fU.S. EPA. 2011.
2005a). Using linear extrapolation from the BMDLio, a human equivalent oral slope factor was
derived (slope factor = 0.1/BMDLio). The resulting oral slope factor is 1 x 10"3 per mg/kg-day.
Quantitative Estimate of Carcinogenic Risk from Inhalation Exposure
A quantitative estimate of carcinogenic potential from inhalation exposure to ETBE was
derived from the same inhalation study used for the estimate of oral carcinogenic risk (Saito etal..
2013: TPEC. 2010bl. A unit risk factor was derived for liver tumors in male F344 rats. The modeled
ETBE POD was scaled to an HEC according to EPA guidance based on inhalation dosimetry for a
Category 3 gas fU.S. EPA. 19941. Using linear extrapolation from the BMCLio, a human equivalent
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inhalation unit risk was derived (inhalation unit risk = 0.1 /BMCLio). The inhalation unit risk is
8 x 10"5 per mg/m3.
Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes
ETBE is metabolized to tert-butanol and acetaldehyde. Evidence is suggestive that genetic
polymorphism of aldehyde dehydrogenase (ALDH)—the enzyme that oxidizes acetaldehyde to
acetic acid—could affect ETBE toxicity. The virtually inactive form, ALDH2*2, is found in about one-
half of all East Asians (and by extension people of East Asian ancestry) fBrennan et al.. 20041.
Evidence is strong in humans that heterozygous ALDH2 increases the internal dose and the cancer
risks from acetaldehyde, especially in the development of alcohol-related cancer in the esophagus
and upper aerodigestive tract, but relevance of this finding on liver tumorigenesis is less clear
(IARC. 20101. Several in vivo and in vitro genotoxicity assays in Aldh2 knockout (KO) mice reported
that genotoxicity was significantly increased compared with wild-type controls following ETBE
exposure to similar doses associated with cancer and noncancer effects in rodents fWengetal..
2014: Wengetal.. 2013: Wengetal.. 2012: Weng etal.. 20111. Inhalation ETBE exposure increased
blood concentrations of acetaldehyde in AIdh2 KO mice compared with wild type. Thus, exposure to
ETBE in individuals with the ALDH2*2 variant would increase the internal dose of acetaldehyde and
potentially increase risks associated with acetaldehyde produced by ETBE metabolism.
Collectively, these data present evidence that people with diminished ALDH2 activity could
be considered a susceptible population that could experience more severe health outcomes.
Key Issues Addressed in Assessment
An evaluation of whether ETBE caused a2U-globulin-associated nephropathy was
performed. ETBE induced an increase in hyaline droplet accumulation and increased a2U-globulin
deposition in male rats; however, with the exception of granular casts and linear mineralization,
most of the subsequent steps in the pathological sequence were not observed despite identical
study conditions and doses in several experiments over a 2-year exposure period. Although CPN
also plays a role in renal tubule nephropathy in both male and female rats, several effects in the
kidney cannot be explained by either the a2U-globulin or CPN processes, including absolute kidney
weight, urothelial hyperplasia, and increased blood biomarkers fSaito etal.. 2013: Suzuki etal..
2012: TPEC. 2010a. b). These specific effects are considered the result of ETBE exposure and
therefore relevant to humans.
In addition, an increase in the incidence of hepatocellular adenomas or carcinomas was
observed in male rats in a 2-year inhalation exposure study fSaito etal.. 2013: TPEC. 2010bl. The
available database for the nuclear hormone receptor MOAs (i.e., PPARa, PXR, and CAR) was
inadequate to determine the role these pathways play, if any, in ETBE-induced liver carcinogenesis.
Acetaldehyde-mediated genotoxicity also was evaluated as a possible MOA, and although evidence
suggests that ALDH2 deficiency enhanced ETBE-induced genotoxicity in exposed mice, the available
database was inadequate to establish acetaldehyde-mediated mutagenicity as an MOA for ETBE-
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induced liver tumors. No other MOAs for liver carcinogenesis were identified, and the rat liver
tumors are considered relevant to humans fU.S. EPA. 2005al
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LITERATURE SEARCH STRATEGY | STUDY
SELECTION AND EVALUATION
A literature search and screening strategy consisted of a broad search of online scientific
databases and other sources to identify all potentially pertinent studies. In subsequent steps,
references were screened to exclude papers not pertinent to an assessment of the health effects of
ETBE, and remaining references were sorted into categories for further evaluation.
The chemical-specific search was conducted in four online scientific databases, PubMed,
Toxline, Web of Science, andTSCATS, through December 2016, using the keywords and limits
described in Table LS-1. The overall literature search approach is shown graphically in Figure LS-1.
Another 114 citations were obtained using additional search strategies described in Table LS-2.
After electronically eliminating duplicates from the citations retrieved through these databases,
847 unique citations were identified.
The resulting 847 citations were screened for pertinence and separated into categories as
presented in Figure LS-1 using the title and either abstract or full text, or both, to examine the
health effects ofETBE exposure. The inclusion and exclusion criteria used to screen the references
and identify sources of health effects data are provided in Table LS-3.
• 33 references were identified as potential "Sources of Health Effects Data" and were
considered for data extraction to evidence tables and exposure-response arrays.
• 70 references were identified as "Supporting Studies." These included 31 studies describing
physiologically based pharmacokinetic (PBPK) models and other toxicokinetic information;
25 studies providing genotoxicity and other mechanistic information; 9 acute, short-term,
or preliminary toxicity studies; and 5 direct administration (e.g., dermal) studies ofETBE.
Although still considered sources of health effects information, studies investigating the
effects of acute and direct chemical exposures are generally less pertinent for characterizing
health hazards associated with chronic oral and inhalation exposures. Therefore,
information from these studies was not considered for extraction into evidence tables.
Nevertheless, these studies were still evaluated as possible sources of supplementary health
effects information.
• 29 references were identified as "Secondary Literature and Sources of Contextual
Information" (e.g., reviews and other agency assessments); these references were retained
as additional resources for development of the Toxicological Review.
• 715 references were identified as being not pertinent (not on topic) to an evaluation of
health effects for ETBE and were excluded from further consideration (see Figure LS-1 for
exclusion categories and Table LS-3 for exclusion criteria). For example, health effect
studies of gasoline and ETBE mixtures were not considered pertinent to the assessment
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because the separate effects of gasoline components could not be determined. Retrieving
numerous references that are not on topic is a consequence of applying an initial search
strategy designed to cast a wide net and to minimize the possibility of missing potentially
relevant health effects data.
The complete list of references as sorted above can be found on the ETBE project page of
the HERO website at https://hero.epa.gov/hero/index.cfm/proiect/page/project id/1376.
Selection of Studies for Inclusion in Evidence Tables
To summarize the important information systematically from the primary health effects
studies in the ETBE database, evidence tables were constructed in a standardized tabular format as
recommended by NRC f2011I Studies were arranged in evidence tables by route of exposure and
then alphabetized by author. Of the studies retained after the literature search and screen, 31 were
identified as "Sources of Health Effects Data" and considered for extraction into evidence tables for
the hazard identification in Section 1. Initial review of studies examining neurotoxic endpoints did
not find consistent effects to warrant a comprehensive hazard evaluation; thus, the one subchronic
study fDorman etal.. 19971 that examined neurotoxic endpoints only was not included in evidence
tables. Data from the remaining 30 studies were extracted into evidence tables.
Supplementary studies that contain pertinent information for the Toxicological Review and
augment hazard identification conclusions, such as genotoxic and mechanistic studies, studies
describing the kinetics and disposition ofETBE absorption and metabolism, and pilot studies, were
not included in the evidence tables. One controlled human exposure toxicokinetic study was
identified, which is discussed in Appendix B.2 (Toxicokinetics). Short-term and acute studies did
not differ qualitatively from the results of the longer-term studies (i.e., >90-day exposure studies).
These were grouped as supplementary studies, however, because the database of chronic and
subchronic rodent studies was considered sufficient for evaluating chronic health effects of ETBE
exposure. Additionally, studies of effects from chronic exposure are most pertinent to lifetime
human exposure (i.e., the primary characterization provided by IRIS assessments) and are the focus
of this assessment Such supplementary studies can be discussed in the narrative sections of
Section 1 and are described in sections such as Mode of action analysis to augment the discussion or
presented in appendices, if they provide additional information.
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Database Searches
(see Table LS-1 for keywords and limits)
f
N
PubMed
n = 148
v.
Web of Science
n = 518
Toxline
(inci. TSCATS)
n = 105
c
\
TSCATS 2
rH
II
c
v.
y
Additional Search Strategies
(see Table LS-2 for methods and
results)
n = 114
Combined Dataset
(After duplicates removed electronically)
n = 847
Manual Screening For Pertinence
(Title/Abstract/Full Text)
Excluded/Not on Topic {n = 715)
50
Abstract only/comment/society
abstracts
82
Biodegradation/environmental fate
385
Chemical analysis/fuel chemistry
180
Other chemical/non ETBE
7
Exposure and biological monitoring
11
Methodology
I
Secondary Literature and Sources of
Contextual Information (n = 29)
1 QSAR
7 Mixtures
14 Reviews/editorials
5 Other agency assessments
2 Odor threshold
J
Supporting Studies
Sources of Health Effects Data (n = 33)
0 Human health effects studies
33 Animal studies
Sources of Supporting Health Effects Data
(n = 14)
5 Not relevant exposure paradigms (e.g.,
dermal, eye irritation)
9 Preliminary/acute data
Sources of Mechanistic and Toxicokinetic
Data (n = 56)
31 PBPK/ADME
13 Genotoxicity
12 Other mechanistic studies
1 Figure LS-1. Summary of literature search and screening process for ETBE.
2
3
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1
2 Table LS-1. Details of the search strategy employed for ETBE
Database
(Search date)
Keywords
Limits
PubMed
(03/31/2014)
Updated
(11/2015)
"ETBE" OR "Ethyl tert-butyl ether"
OR "2-ethoxy-2-methyl-propane" OR
"ethyl tertiary butyl ether" OR "ethyl
tert-butyl oxide" OR "tert-butyl ethyl
ether" OR "ethyl t-butyl ether" OR
"637-92-3"
None
Web of Science
(03/31/2014)
Updated
(11/2015)
"ETBE" OR "ethyl tert-butyl ether"
OR "2-ethoxy-2-methyl-propane" OR
"ethyl tertiary butyl ether" OR "ethyl
tert-butyl oxide" OR "tert-butyl ethyl
ether" OR "ethyl t-butyl ether" OR
"637-92-3"
Lemmatization on
Toxline
(includes
TSCATS)
(03/31/2014)
Updated
(11/2015)
"ETBE" OR "Ethyl tert-butyl ether"
OR "2-Ethoxy-2-methyl-propane" OR
"ethyl tertiary butyl ether" OR "ethyl
tert-butyl oxide" OR "tert-butyl ethyl
ether" OR "ethyl t-butyl ether" OR
"637-92-3"
Not PubMed
TSCATS2
(3/31/2014)
Updated
(12/2016)
637-92-3
01/01/2004 to 12/01/2016
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Table LS-2. Summary of additional search strategies for ETBE
Approach used
Source(s)
Date
performed
Number of additional
references identified
Electronic
backward search
through Web of
Science
Review article: McGregor (2007).
"Ethyl tertiary-butyl ether: a
toxicological review." Critical
Reviews in Toxicology 37(4): 287
-312
3/2014
68 references
Review article: de Pevster (2010).
"Ethyl t-butyl ether: Review of
reproductive and developmental
toxicity." Birth Defects Research,
Part B: Developmental and
Reproductive Toxicology 89(3):
239-263
3/2014
26 references
Personal
communication
Japan Petroleum Energy Center
3/2014
Updated
(12/2016)
21 references
Table LS-3. Inclusion-exclusion criteria
Inclusion criteria
Exclusion criteria
Population
• Humans
• Standard mammalian animal models,
including rat, mouse, rabbit, guinea
pig, monkey, dog
• Ecological species*
• Nonmammalian species*
Exposure
• Exposure is to ETBE
• Exposure is measured in an
environmental medium (e.g., air,
water, diet)
• Exposure via oral or inhalation routes;
for supporting health effect studies,
exposure via oral or inhalation routes
• Study population is not exposed to ETBE
• Exposure to a mixture only (e.g., gasoline
containing ETBE)
• Exposure via injection (e.g., intravenous)
• Exposure paradigm not relevant (e.g., acute,
dermal, or ocular)
Outcome
• Study includes a measure of one or
more health effect endpoints,
including effects on the nervous,
kidney/urogenital, musculoskeletal,
cardiovascular, immune, and
gastrointestinal systems;
reproduction; development; liver;
eyes; and cancer
• Odor threshold studies
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Inclusion criteria
Exclusion criteria
Other
Not on topic, including:
• Abstract only, editorial comments, policy
papers, were not considered further because
study was not potentially relevant
• Bioremediation, biodegradation, or
environmental fate of ETBE, including
evaluation of wastewater treatment
technologies and methods for remediation of
contaminated water and soil
• Chemical, physical, or fuel chemistry studies
• Analytical methods for measuring/detecting/
remotely sensing ETBE
• Not chemical specific: Studies that do not
involve testing of ETBE
• Quantitative structure activity relationship
studies
• Exposure studies without health effect
evaluation
*Studies that met this exclusion criterion were not considered a source of health effects or supplementary health
effects data/mechanistic and toxicokinetic data, but were considered as sources of contextual information.
1 Database Evaluation
2 For this draft assessment, 30 experimental animal studies comprised the primary sources of
3 health effects data; no studies were identified that evaluated humans exposed to ETBE (e.g., cohort
4 studies, case reports, ecological studies). The animal studies were evaluated considering aspects of
5 design, conduct, or reporting that could affect the interpretation of results, overall contribution to
6 the synthesis of evidence, and determination of hazard potential as noted in various EPA guidance
7 documents (U.S. EPA. 2005a. 1998b. 1996.1991b). The objective was to identify the stronger, more
8 informative studies based on a uniform evaluation of quality characteristics across studies of
9 similar design. Studies were evaluated to identify their suitability based on:
• Study design
• Nature of the assay and validity for its intended purpose
• Characterization of the nature and extent of impurities and contaminants of ETBE
administered, if applicable
• Characterization of dose and dosing regimen (including age at exposure) and their
adequacy to elicit adverse effects, including latent effects
• Sample sizes to detect dose-related differences or trends
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• Ascertainment of survival, vital signs, disease or effects, and cause of death
• Control of other variables that could influence the occurrence of effects
Additionally, several general considerations, presented in Table LS-4, were used in evaluating the
animal studies (Table LS-5). Much of the key information for conducting this evaluation can be
determined based on study methods and how the study results were reported. Importantly, the
evaluation at this stage does not consider the direction or magnitude of any reported effects.
EPA considered statistical tests to evaluate whether the observations might be due to
chance. The standard for determining statistical significance of a response is a trend test or
comparison of outcomes in the exposed groups against those of concurrent controls. Studies that
did not report statistical testing were identified and, when appropriate, statistical tests were
conducted by EPA.
Information on study features related to this evaluation is reported in evidence tables and
documented in the synthesis of evidence. Discussions of study strengths and limitations are
included in the text where relevant If EPA's interpretation of a study differs from that of the study
authors, the draft assessment discusses the basis for the difference.
Experimental Animal Studies
The 30 experimental animal studies, all of which were performed on rats, mice, and rabbits,
were associated with drinking water, oral gavage, or inhalation exposures to ETBE. A large
proportion of these studies was conducted according to Organisation for Economic Co-operation
and Development Good Laboratory Practice (GLP) guidelines, presented extensive
histopathological data, or clearly presented their methodology; thus, they are considered high
quality. For the remaining studies, a more detailed discussion of methodological concerns that were
identified precedes each endpoint evaluated in the hazard identification section. Overall, the
experimental animal studies ofETBE involving repeated oral or inhalation exposure were
considered acceptable quality, and whether yielding positive, negative, or null results, were
considered in assessing the evidence for health effects associated with chronic exposure to ETBE.
Table LS-4. Considerations for evaluation of experimental animal studies
Methodological
feature
Considerations
(relevant information extracted into evidence tables)
Test animal
Suitability of species, strain, sex, and source of test animals
Experimental design
Suitability of animal age/lifestage at exposure and endpoint testing; periodicity and
duration of exposure (e.g., hr/day, day/week); timing of endpoint evaluations; and
sample size and experimental unit (e.g., animals, dams, litters)
Exposure
Characterization of test article source, composition, purity, and stability; suitability of
control (e.g., vehicle control); documentation of exposure techniques (e.g., route,
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Toxicological Review ofETBE
chamber type, gavage volume); verification of exposure levels (e.g., consideration of
homogeneity, stability, analytical methods)
Endpoint evaluation
Suitability of specific methods for assessing endpoint(s) of interest
Results presentation
Data presentation for endpoint(s) of interest (including measures of variability) and for
other relevant endpoints needed for results interpretation (e.g., maternal toxicity,
decrements in body weight relative to organ weight)
Table LS-5. Summary of experimental animal database
Study Category
Study duration, species/strain, and administration method
Chronic
2-vear studv in F344 rats (drinking water) JPEC (2010a);Suzuki et al. (2012)
2-vear studv in F344 rats (inhalation) JPEC (2010b), Saito et al. (2013)
2-vear studv in Sprague-Dawlev rats (gavage) Maltoni et al. (1999)
2-vear studv in F344 rats (drinking water) JPEC (2010a)*
2-vear studv in F344 rats (inhalation) JPEC (2010b)*
Subchronic
13-week studv in F344 rats (inhalation) Medinskv et al. (1999); Bond et al. (1996b)
26-week studv in Sprague-Dawlev rats (gavage) JPEC (2008c); Mivata et al. (2013)
Fuiiietal. (2010); JPEC (2008e)
13-week studv in Sprague-Dawlev rats (inhalation) JPEC (2008b)
23-week studv in F344 rats (gavage) Hagiwara et al. (2011); JPEC (2008d)
13-week studv in CD-I mice (inhalation) Medinskv et al. (1999); Bond et al. (1996a)
23-week studv in Wistar rats (gavage) Hagiwara et al. (2015)
31-week studv in F344/DuCrlCrli rats (drinking water) Hagiwara et al. (2013)
13-week studv in C57BL/6 mice (inhalation) Weng et al. (2012)
26-week studv in Sprague-Dawlev rats (gavage) JPEC (2008c)*
13-week studv in Sprague-Dawlev rats (inhalation) JPEC (2008b)*
Reproductive
Two-generation reproductive toxicity studv on Sprague-Dawlev rats (gavage) Gaoua
(2004b)*
One-generation reproductive toxicity studv on Sprague-Dawlev rats (gavage) Fuiii et al.
(2010); JPEC (2008e)
2-week studv on Simonson albino rats (drinking water) Berger and Horner (2003)
9-week studv on C57BL/6 mice (inhalation) Weng et al. (2014)
14-dav studv on F344 rats (gavage) de Pevster et al. (2009)
Two-generation reproductive toxicity studv in Sprague-Dawlev rats (gavage) Gaoua
(2004b)*
Developmental
Developmental studv (GD6-27) on New Zealand rabbits (gavage) Asano et al. (2011);
JPEC (2008i)
Developmental studv (GD5-19) on Sprague-Dawlev rats (gavage) Aso et al. (2014); JPEC
(2008h)
Developmental studv (GD5-19) on Sprague-Dawlev rats (gavage) Gaoua (2004b)*
Developmental studv (GD5-19) on Sprague-Dawlev rats (gavage) Gaoua (2004a)*
Pharmacokinetic
Single-dose studv on Sprague-Dawlev rats (gavage) JPEC (2008g)
14-dav studv on Sprague-Dawlev rats (gavage) JPEC (2008f)
Single-dose studv on Sprague-Dawlev rats (gavage) JPEC (2008g)*
14-dav studv on Sprague-Dawlev rats (gavage) JPEC (2008f)*
2 *The IRIS program had this study peer reviewed.
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i 1.HAZARD IDENTIFICATION
2 1.1. OVERVIEW OF CHEMICAL PROPERTIES AND TOXICOKINETICS
3 1.1.1. Chemical Properties
4 ETBE is a liquid at a temperature range of -94 to 72.6°C. It is soluble in ethanol, ethyl ether,
5 and water fDrogos and Diaz. 20011. ETBE has a strong, highly objectionable odor and taste at
6 relatively low concentrations. The chemical is highly flammable and reacts with strong oxidizing
7 agents. ETBE is stable when stored at room temperature in tightly closed containers fDrogos and
8 Diaz. 20011. Selected chemical and physical properties ofETBE are presented in Table 1-1.
9 Table 1-1. Physicochemical properties and chemical identity of ETBE
Characteristic or property
Value
Reference
Chemical name
2-ethoxy-2-methylpropane
2-methyl-2-ethoxypropane
NLM (2016)
Synonyms
ethyl te/t-butyl ether
ethyl te/t-butyl oxide
methyl-2-ethoxypropane
tert-butyl ethyl ether
ETBE
NLM (2016)
Chemical formula
C6H14O
NLM (2016)
CASRN (Chemical Abstracts Service
Registry Number)
637-92-3
NLM (2016)
Molecular weight
102.17
NLM (2016)
Melting point
-94°C
Drogos and Diaz (2001)
Boiling point
73.1°C
ECHA (2016)
Density at 20°C
0.74 g/cm3 @ 20°C
ECHA (2016)
Water solubility
2.37 g/L
Drogos and Diaz (2001)
Partition coefficients:
Log oil/water at 25°C
Log Kow
1.48
1.74
Montgomery (1994)
Drogos and Diaz (2001)
Viscosity at 40°C
0.47 mm2/s
ECHA (2016)
Vapor pressure
17kPa@ 25°C
NLM (2016)
Henry's Law Constant
1.39 x 10"3 atm-m3/mol @ 25°C
NLM (2016)
Odor
0.013 ppm (0.054 mg/m3)
Vetrano (1993)
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Characteristic or property
Value
Reference
Detection threshold
Recognition threshold
0.024 ppm (0.1 mg/m3)
Taste detection threshold (in water)
0.047 ppm (47 Hg/L)
Vetrano (1993)
Odor detection threshold (in water)
0.049 ppm (49 Hg/L)
Vetrano (1993)
Odor detection threshold (in water)
0.005 ppm (5 Hg/L)
Vetrano (1993)
Conversion factors
1 ppm = 4.18 mg/m3
1 mg/m3 = 0.24 ppm
1 mg/m3 = 102,180 mmol/L
Chemical structure
ch3
0 ch3
H3C—^ CH3
HSDB (2012)
1.1.2. Toxicokinetics
ETBE is rapidly absorbed following exposure by oral and inhalation routes (see Appendix
B.l.l). Studies in experimental animals indicate that >90% of the compound was absorbed after
oral administration within 6-10 hours flPEC. 2008d. e). No data are available for oral absorption in
humans. ETBE is moderately absorbed following inhalation exposure in both rats and humans;
human blood levels of ETBE approached—but did not reach—steady-state concentrations within
2 hours, and a net respiratory uptake ofETBE was estimated to be 26% fNihlen etal.. 1998bl.
ETBE and its metabolite, tert-butanol, are distributed throughout the body following oral,
inhalation, and i.v. exposures flPEC. 2008d. e; Poet etal.. 1997: Faulkner etal.. 1989: ARCO. 19831.
Following exposure to ETBE in rats, ETBE was found in kidney, liver, and blood. Comparison of
ETBE distribution in rats and mice demonstrated that concentrations ofETBE in the rat kidney and
mouse liver are proportional to the blood concentration.
A general metabolic scheme for ETBE, illustrating the biotransformation in rats and
humans, is shown in Figure 1-1 (see Appendix B.1.3).
Human data on the excretion ofETBE was measured in several studies fNihlen et al.. 1998a.
c). The half-life ofETBE in urine was biphasic with half-lives of 8 minutes and 8.6 hours flohanson
etal.. 19951. These studies showed urinary excretion ofETBE to be less than 0.2% of the uptake or
absorption ofETBE fNihlen et al.. 1998a. c). Amberg etal. (20001 observed a similar half-life of 1-6
hours after human exposure to ETBE of 170 mg/m3; however, the elimination for ETBE in rat urine
was considerably faster than in humans, and ETBE itself was undetectable in rat urine.
A more detailed summary of ETBE toxicokinetics is provided in Appendix B.l.
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CYP2A6
CYP3A4
glucuronide-0 ——ch3
CH3
t-butyl glucuronide
rats, humans
ETBE
h3c—\ CH3
OH
ETBE-hemi-acetal
Cl-h OH
CYP450
ch3 H^C-
rats,
CH3 humans CH3 oh
t-butanol 2-methyl-1,2-propanediol
h3c—|—oh
H3
CH-j
H3C ^ CH^
\
acetaldehyde °~
ch3
t-butyl sulfate
ho_o
[O]
2-hydroxyisobutyric acid
h2c=o
formaldehyde
acetone
Source: Adapted from Dekant et al. (2001), NSF International (2003), ATSDR (1996), Bernauer et al.
(1998), Ambers et al. (1999), and Cederbaum and Cohen (1980).
Figure 1-1. Proposed metabolism of ETBE.
1.1.3. Description of Toxicokinetic Models
Two physiologically based pharmacokinetic (PBPK) models have been developed
specifically for administration of ETBE in rats (Borghoffetal.. 2016: Salazar etal.. 20151. The
previously available models have studied tert-butanol as the primary metabolite after oral or
inhalation exposure to MTBE in rats and humans or ETBE in humans. Models for MTBE oral and
inhalation exposure include a component for the binding of tert-butanol to a2U-globulin fBorghoff et
al.. 2010: Leavens and Borghoff. 20091. A PBPK model for inhalation exposure of humans to ETBE
has also been reported fNihlen and Tohanson. 19991. A more detailed summaiy of the toxicokinetic
models is provided in Appendix B.1.5 fU.S. EPA. 20171.
1.1.4. Related Chemicals that Provide Supporting Information
ETBE is metabolized to acetaldehyde and tert-butanol, and effects induced by these
metabolites can provide support for ETBE-induced effects. Some of the toxicological effects
observed in ETBE are attributed to tert- butanol f Salazar etal.. 20151. Animal studies demonstrate
that chronic exposure to tert-butanol is associated with noncancer kidney effects, including
increased kidney weights in male and female rats accompanied by increased chronic progressive
nephropathy (CPN), urothelial hyperplasia (in males and females), and increased suppurative
inflammation in females (NTP. 1997.1995b).
Inhalation exposures to acetaldehyde were concluded to cause carcinomas of the nasal
mucosa in rats and carcinomas of the larynx in hamsters flARC. 1999bl. In addition, acetaldehyde
was concluded to be the key metabolite in cancer of the esophagus and aerodigestive tract
associated with ethanol consumption flARC. 20101.
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MTBE is a structurally related compound that is metabolized to formaldehyde and
tert-butanol. In 1996, the U.S. Agency for Toxic Substances and Disease Registry's (ATSDR)
Toxicological Profile for MTBE fATSDR. 1996] identified cancer effect levels of MTBE based on
carcinogenicity data in animals. ATSDR reported that inhalation exposure resulted in kidney cancer
in rats and liver cancer in mice. ATSDR concluded that oral exposure to MTBE might cause liver and
kidney damage and nervous system effects in rats and mice. The chronic inhalation minimal risk
level was derived based on incidence and severity of chronic progressive nephropathy in female
rats fATSDR. 19961. In 1997, EPA's Office of Water concluded that MTBE is carcinogenic to animals
and poses a carcinogenic potential to humans based on an increased incidence of Leydig cell
adenomas of the testes, kidney tumors, lymphomas, and leukemia in exposed rats (U.S. EPA. 1997).
In 1998, the International Agency for Research on Cancer (IARC) found "limited" evidence of MTBE
carcinogenicity in animals and classified MTBE in Group 3 (i.e., not classifiable as to carcinogenicity
in humans) flARC. 1999dl. IARC reported that oral exposure in rats resulted in testicular tumors in
males and lymphomas and leukemias (combined) in females; inhalation exposure in male rats
resulted in renal tubule adenomas; and inhalation exposure in female mice resulted in
hepatocellular adenomas flARC. 1999d).
1.2. PRESENTATION AND SYNTHESIS OF EVIDENCE BY ORGAN/SYSTEM
1.2.1. Kidney Effects
Synthesis of Effects in Kidney
This section reviews the studies that investigated whether subchronic or chronic exposure
to ETBE can cause kidney toxicity or cancer in humans or animals. The database examining kidney
effects following ETBE exposure contains no human data and 10 animal studies, predominantly in
rats. Exposures ranged from 13 weeks to 2 years and both inhalation and oral exposure routes are
well represented. Studies using short-term and acute exposures that examined kidney effects are
not included in the evidence tables; however, they are discussed in the text if they provided data to
inform mode of action (MOA) or hazard identification. Four unpublished technical reports relevant
to the kidney were externally peer reviewed at the request of EPA in August 2012 (Table LS-5):
TPEC (2010a). TPEC (2010b). TPEC (2008c). TPEC (2008b). some of which were subsequently
published. These are TPEC (2010a) [published as Suzuki etal. (2012)]. TPEC (2010b) [published as
Saito etal. (2013)]. and TPEC (2008c) [published as Mivata etal. (2013)]. Gaoua (2004b) was
externally peer reviewed at the request of EPA in November 2008. Studies are arranged in evidence
tables by effect and alphabetical order by author.
The unpublished report by Cohen etal. f20111 was not peer reviewed externally. In Cohen
etal. (2011). a pathology working group reexamined kidney histopathology from the TPEC (2010a)
[subsequently published as Suzuki etal. (2012)] and TPEC (2007a) studies. Cohen etal. (2011) did
not report incidences of carcinomas that differed from those in the original study (Suzuki etal..
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Toxicological Review ofETBE
2012: IPEC. 2010a]; thus, these data have been presented only once. Histopathological results from
both Cohen etal. f20111 and TPEC f2007bl are considered for hazard identification.
The design, conduct, and reporting of each study were reviewed, and each study was
considered adequate to provide information pertinent to this assessment Interpretation of non-
neoplastic kidney endpoints in rats, however, is complicated by the common occurrence of age-
related spontaneous lesions characteristic of CPN (NTP. 2015: Hard etal.. 2013: Melnick etal..
2012: U.S. EPA. 1991al: http://ntp.niehs.nih.gov/nnl/urinary/kidnev/necp/index.html. CPN is
more severe in male rats than in females and is particularly common in the Sprague-Dawley and
Fischer 344 strains. Dietary and hormonal factors play a role in modifying CPN, although the
etiology is largely unknown (see further discussion below).
Kidney weight In most of the studies with data available for relative and absolute organ
weight comparisons, both relative and absolute kidney weights are increased fMivata etal.. 2013:
Saito etal.. 2013: Suzuki etal.. 2012: TPEC. 2010b. 2008b. c; Gaoua. 2004bl. Measures of relative, as
opposed to absolute, organ weight are sometimes preferred because they account for changes in
body weight that might influence changes in organ weight fBailev etal.. 20041. although potential
impact of body weight changes should be evaluated. For ETBE, body weight in exposed animals was
consistently decreased at several doses relative to controls in the oral and inhalation studies. In this
case, the decreased body weight of the animals affects the relative kidney weight measures,
resulting in an artificial exaggeration of changes. Additionally, a recent analysis indicates that
absolute, but not relative, subchronic kidney weights are significantly correlated with chemically
induced histopathological findings in the kidney in chronic and subchronic studies fCraig etal..
20141. Therefore, absolute weight was determined the more reliable measure of kidney weight
change for determining ETBE hazard potential. Numerical absolute and relative kidney weight data
are presented in Appendix B of the Supplemental Information.
Absolute kidney weights (see Figure 1-2) exhibited strong dose-related increases in male
rats following oral exposures (Spearman's rank coefficient = 0.86, p < 0.01) of 16 weeks or longer
fMivata etal.. 2013: Suzuki etal.. 2012: Fuiii etal.. 2010: TPEC. 2010a. 2008c: Gaoua. 2004bl. and
following inhalation exposures (Spearman's rank coefficient = 0.71, p = 0.05) of 13 weeks or longer
fSaito etal.. 2013: TPEC. 2010b. 2008b: Medinskv etal.. 19991. Changes in female rats also had
strong dose-related increases following inhalation exposure (Spearman's rank coefficient = 0.82,
p = 0.01) and moderate dose-related increases following oral exposure (Spearman's rank coefficient
= 0.42, p = 0.2). Short-term studies in rats also observed increased kidney weight (TPEC. 2008a). In
utero ETBE exposure induced greater increases in absolute kidney weights in F1 male and female
rats compared to parental exposure in one unpublished study f Gaoua. 2004bl. but the magnitude of
increases were comparable to those observed in other adult oral studies. The single mouse
inhalation study observed weak increases in kidney weight in both sexes (Figure 1-3).
Available 2-year kidney weight data were not considered appropriate for hazard
identification due to the prevalence of age-associated confounders such as CPN and mortality that
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affect organ weight analysis fSaito etal.. 2013: Suzuki etal.. 2012: TPEC. 2010a. b). CPN is an age-
associated disease characterized by cell proliferation and chronic inflammation that results in
increased kidney weight (Melnick et al.. 2 012: Travlos etal.. 20111. Most (64-100%) male and
female rats in the 2-year oral and inhalation studies were observed to have CPN regardless ofETBE
administration fSaito etal.. 2013: Suzuki etal.. 2012: TPEC. 2010a. b). Although mortality in the
2-year studies was significantly increased in ETBE-treated male and female rats compared with
controls following oral and inhalation exposure (see Appendix B.1.5), causes of death were the
result of age-associated diseases, such as CPN. Because using kidney weight data from these 2-year
studies would impart bias by selecting animals that survive to the end of the study for organ weight
analysis (e.g., deceased animals with CPN could have enlarged kidneys), the 2-year organ weight
data are not appropriate for hazard identification and are not discussed further.
Kidney histopathology. Kidney lesions also were observed in several studies. The incidence
of nephropathy, which was characterized as CPN due to sclerosis of glomeruli, thickening of the
renal tubular basement membranes, inflammatory cell infiltration, and interstitial fibrosis, was not
increased in any chronic study because ofETBE exposure. The severity of CPN, however, was
exacerbated by ETBE in male and female rats in a 2-year inhalation study, and the number of CPN
foci was increased in male rats in a 13-week drinking water study (see Table 1-2) (Cohen etal..
2011: TPEC. 2010b. 2007a). The effects characterized as CPN are related to age and not considered
histopathological manifestations of chemically induced toxicity [see U.S. EPA (1991a). p. 35 for
further details and a list of the typical observable histopathological features of CPN], CPN is a
common and well-established constellation of age-related lesions in the kidney of rats, and there is
no known counterpart to CPN in aging humans. CPN is not a specific diagnosis on its own but an
aggregate term describing a spectrum of effects. These individual lesions or processes may well
occur in a human kidney, and the fact they happen to occur as a group in the aged rat kidney does
not assure that the individual lesions are rat-specific if there is a treatment effect for one or more of
them. In addition, exacerbation of one of more of these processes likely reflects some type of cell
injury, which is relevant to the human kidney. Increases in CPN graded as marked or severe were
dose-related when compared on an internal dose basis across routes of exposure in male and
female rats fSalazar etal.. 20151.
Increased incidence of urothelial hyperplasia (graded as slight or minimal) was observed in
male rats in 2-year studies by both inhalation and oral exposure (Saito etal.. 2013: Suzuki etal..
2012: TPEC. 2010a. b). The increase in urothelial hyperplasia incidence appeared to be dose related
on an internal dose basis across routes of exposure (Salazar etal.. 2015). Cohen etal. (2011).
however, attributed this effect to CPN rather than the "direct" result ofETBE treatment To
determine if the severity of the hyperplasia was positively associated with the severity of CPN,
contingency tables comparing the occurrence of urothelial hyperplasia with CPN in individual rats
were arranged by severity and analyzed with Spearman's rank correlation tests to determine
strength of associations for each comparison (Table 1-25,1-6). Urothelial hyperplasia and CPN
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were weakly correlated (Spearman's rank coefficient =0.36) in males following oral and inhalation
exposure to ETBE. The biological significance of urothelial hyperplasia and any relationship with
CPN is discussed in Mode of action analysis (see below).
The number and size of hyaline droplets were increased in the proximal tubules of male
rats, but not in females, and the hyaline droplets tested positive for the presence of a2U-globulin
(Mivata etal.. 2013: TPEC. 2008c. e, f; Medinskv etal.. 1999). The significance of this effect, along
with other potentially related histopathological effects, such as necrosis, linear tubule
mineralization, and tubular hyperplasia, are discussed in Mode of action analysis (see below).
Serum and urinary biomarkers. The increased kidney weight and CPN in male rats is
associated with several changes in urinary and serum biomarkers of renal function (see Table 1-2,
Table 1-3). CPN is proposed to be associated with several changes in urinary and serum measures
such as proteinuria, blood urea nitrogen (BUN), creatinine, and hypercholesterolemia (Hard etal..
20091. ETBE exposure, however, increased serum measures at lower doses and in more studies
than were associated with increased CPN severity. Considering male rat blood concentrations in
both chronic and subchronic studies, total cholesterol was elevated in 3 of 4 studies, BUN was
elevated in 2 of 4 studies, and creatinine was elevated 1 of 4 studies fMivata etal.. 2013: Saito etal..
2013: Suzuki etal.. 2012: TPEC. 2010a. b, 2008c). In F344 female rats, cholesterol and BUN were
elevated at the highest dose in one chronic inhalation study, which corresponded with an elevated
CPN response in females (Saito etal.. 2013: TPEC. 2010b). The single reported instance of elevated
proteinuria occurred in female rats following chronic inhalation exposure; thus, no correlation of
elevated proteinuria with CPN in males was observed f Saito etal.. 2013: IPEC. 2010b],
Kidney tumors. No increase in kidney tumor incidence was observed following 2 years of
oral or inhalation exposure in either male or female F344 rats f Saito etal.. 2013: Suzuki etal.. 2012:
TPEC. 2010a. b)(see Table 1-4). In two-stage ("initiation, promotion") cancer bioassays, 23 weeks of
daily gavage ETBE exposure did not increase kidney tumor incidence following 4 weeks of
treatment with a 5-mutagens mixture (DMBDD) in male F344 rats (Hagiwara etal.. 2011: TPEC.
2008d); however, a dose-dependent increase in renal tubular adenoma or carcinoma incidence was
observed with 19 weeks of daily gavage ETBE exposure following 2 weeks of N-ethyl-N-
hydroxyethylnitrosamine (EHEN) administration in male Wistar rats fHagiwara etal.. 20151. In
Hagiwara etal. f20111. kidney tumors were not observed following 23 weeks ofETBE exposure
without mutagen exposure, although such an ETBE-only exposure group was not evaluated in the
later study in Wistar rats (Hagiwara etal.. 2015).
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Toxicological Review ofETBE
Male rats Female rats
rho= 0.75 (all)
rho= 0.86 (oral) #
rho= 0.71 (inh.)
•
°8
0
rho= 0.46 (all)
rho= 0.42 (oral)
rho= 0.82 (inh.)
o
•
0
0#
o
o
•
o
. 0 •
•
•
o
•
o®
o
•
% . •
o#
•
0 20 40 60 80 100 0 20 40 60 80 100
tert-butanol blood concentration (mg/l) tert-butanol blood concentration (mg/l)
• Oral exposure
O Inhalation exposure
Figure 1-2. Comparison of absolute kidney weight change in male and female
rats across oral and inhalation exposure based on internal blood
concentration. Spearman rank coefficient (rho) was calculated to evaluate the
direction of a monotonic association (e.g., positive value = positive association) and
the strength of association.
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Toxicological Review ofETBE
D)
£=
TO
gj
a>
<:
>>
£=
¦g
o
(/)
JD
<
Mouse inhalation exposure
5000
10000
15000
20000
25000
Administered concentration (mg/m
•
Male mice
o
Female mice
1
2
3
4
Figure 1-3. Comparison of absolute kidney weight change in male and female
mice following inhalation exposure based on administered ETBE
concentration. No significant relationships were calculated.
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Toxicological Review ofETBE
Table 1-2. Evidence pertaining to kidney histopathology effects in animals
following exposure to ETBE
Reference and study
Results
design
Cohen et al. (2011)
Male
Female
rat, F344/DuCrlCrlj
oral - water
male (50/group): 0, 625,
2,500,10,000 ppm (0, 28,
Dose
(mg/kg-d)
Average Average
severity of Incidence of Dose severity of Incidence of
CPN CPN (mg/kg-d) CPN CPN
121, 542 mg/kg-d)a; female
0
2.08 49/50
0
1.14 45/50
(50/group): 0, 625, 2,500,
28
46
0.98 41/50
10,000 ppm (0, 46, 171,
560 mg/kg-d)a
121
-
171
1.2 46/50
reanalysis of histopathology
data from JPEC (2010a)
542
2.72* 50/50
560
1.36 46/50
study, for which animals
were dosed daily for 104 wk
Cohen et al. (2011)
Male
rat, F344/DuCrlCrlj
oral - water
male (10/group): 0, 250,
Dose
(mg/kg-d)
Number of CPN foci/rat Number of granular casts/rat
1,600, 4,000, 10,000 ppm
0
1.2
0
(0, 17, 40,101, 259,
17
626 mg/kg-d)a
reanalysis of histopathology
40
-
-
data from JPEC 2007 (study
No. 0665) study, for which
101
-
-
animals were dosed daily for
259
-
-
13 wk
626
27.2
8.2
Mivata etal. (2013); JPEC
Male
Female
(2008c)
rat, CRL:CD(SD)
oral -gavage
male (15/group): 0, 5, 25,
Incidence of
Incidence of
Dose
(mg/kg-d)
papillarv
mineralization
Dose
mg/kg-d)
Daoillarv
mineralization
100, 400 mg/kg-d; female
0
0/15
0
0/15
(15/group): 0, 5, 25,100,
400 mg/kg-d
5
0/15
5
-
daily for 180 d
25
0/15
25
-
100
1/15
100
-
400
0/15
400
0/15
This document is a draft for review purposes only and does not constitute Agency policy.
1-10 DRAFT—DO NOT CITE OR QUOTE
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Toxicological Review ofETBE
Reference and study
Results
design
Saito etal. (2013); JPEC
Male
Average
Incidence of
(2010b)
severity of CPN
Incidence of
urothelial
rat, Fischer 344
as calculated
Incidence of
Daoillarv
hvoerolasia of
inhalation - vapor
Dose (mg/m
3) bv EPAC
CPN
mineralization
the renal oelvis
male (50/group): 0, 500,
1,500, 5,000 ppm (0, 2,090,
0
2.4
49/50
0/50
2/50
6,270, 20,900 mg/m3)b;
2,090
2.6
50/50
0/50
5/50
female (50/group): 0, 500,
1,500, 5,000 ppm (0, 2,090,
6,270
2.7
49/49
1/49
16/49*
6,270, 20,900 mg/m3)b
20,900
3.1*
50/50
6/50*
41/50*
dynamic whole body
inhalation; 6 hr/d, 5 d/wk
Female
Average
for 104 wk; generation
severity of CPN
method, analytical
as calculated
Incidence of
concentration reported
Dose (mg/m
0
2,090
6,270
20,900
3) bv EPAC
0.9
1.3
1.3
1.6*
CPN
32/50
38/50
41/50
40/50
Atypical tubule hyperplasia not observed in males or females.
Papillary mineralization and urothelial hyperplasia of the renal pelvis not observed
in females.
Suzuki etal. (2012); JPEC
Male
Average
(2010a)
severity of CPN
Incidence of
rat, Fischer 344
Dose
Average
as calculated bv
atvDical tubule
Incidence of
oral - water
(mg/kg-d)
severity of CPN
EPAC
hvoerolasia
CPN
male (50/group): 0, 625,
2,500,10,000 ppm (0, 28,
0
2.1
2.1
0/50
49/50
121, 542 mg/kg-d)a; female
28
2.0
1.7
0/50
43/50
(50/group): 0, 625, 2,500,
10,000 ppm (0, 46, 171,
121
2.0
1.8
0/50
45/50
560 mg/kg-d)a
542
2.4*
2.3
1/50
48/50
daily for 104 wk
Incidence of
Incidence of
Incidence of
urothelial
Dose
oaoillarv
Daoillarv
hvoerolasia of
(mg/kg-d)
necrosis
mineralization
the renal oelvis
0
0/50
0/50
0/50
28
1/50
0/50
0/50
121
0/50
16/50*
10/50*
542
2/50
42/50*
25/50*
This document is a draft for review purposes only and does not constitute Agency policy.
1-11 DRAFT—DO NOT CITE OR QUOTE
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Toxicological Review ofETBE
Reference and study
design
Results
Female
Average
severity of CPN
Incidence of
Dose
Average
as calculated bv
atvoical tubule
Incidence of
(mg/kg-d)
severity of CPN
EPAC
hyperplasia
CPN
0
1.2
1.0
0/50
41/50
46
1.2
0.9
0/50
37/50
171
1.5
1.1
0/50
37/50
560
1.5*
1.2
2/50
39/50
Incidence of
Incidence of
Incidence of
urothelial
Dose
papillary
Daoillarv
hyperplasia of
(mg/kg-d)
necrosis
mineralization
the renal pelvis
0
0/50
0/50
0/50
46
1/50
0/50
0/50
171
1/50
1/50
0/50
560
2/50
3/50
0/50
Conversion performed by study authors.
b4.18 mg/m3 = 1 ppm.
cAverage severity calculated as (grade x number of affected animals) -f total number of animals exposed.
*: result is statistically significant (p < 0.05) based on analysis of data by study authors.
for controls, no response relevant; for other doses, no quantitative response reported.
Percent change compared to controls calculated as 100 x [(treated value - control value) 4 control value].
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
1 Table 1-3. Evidence pertaining to kidney biochemistry and urine effects in
2 animals following exposure to ETBE
Reference and study
Results
design
JPEC (2008b)
Male
rat, CRL:CD(SD)
inhalation - vapor
male (10/group): 0,150,
Dose (mg/m3)
Blood urea nitrogen
(BUN)
Cholesterol
Creatinine
500, 1,500, 5,000 ppm
0
-
-
-
(0, 627, 2,090, 6,270,
20,900 mg/m3)a; female
627
-9%
8%
-13%
(10/group): 0,150, 500,
2,090
-5%
9%
-6%
1,500, 5,000 ppm (0, 627,
2,090, 6,270,
6,270
4%
26%
-6%
20,900 mg/m3)a
20,900
4%
15%
-3%
dynamic whole body
chamber; 6 hr/d, 5 d/wk for
Dose (mg/m3)
Proteinuria severitvb
Proteinuria incidence
Urinarv casts
13 wk; generation method,
0
0.5
3/6
0/6
analytical concentration,
and method reported
627
1.2
5/6
0/6
2,090
1.2
5/6
0/6
6,270
1.3
6/6
0/6
20,900
1.0
4/6
0/6
Female
Blood urea nitrogen
Dose (mg/m3)
0
627
(BUN)
Cholesterol
Creatinine
-5%
7%
0%
2,090
3%
9%
3%
6,270
-8%
11%
-9%
20,900
-4%
21%
-9%
Dose (mg/m3)
Proteinuria severitvb
Proteinuria incidence
Urinarv casts
0
0.2
1/6
0/6
627
0.3
1/6
0/6
2,090
0.2
1/6
0/6
6,270
0.5
2/6
0/6
20,900
0.3
2/6
0/6
This document is a draft for review purposes only and does not constitute Agency policy.
1-13 DRAFT—DO NOT CITE OR QUOTE
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Toxicological Review ofETBE
Reference and study
Results
design
Mivata etal. (2013); JPEC
Male
(2008c)
rat, CRL:CD(SD)
oral -gavage
Dose
(mg/kg-d)
Blood urea nitrogen
(BUN)
Cholesterol
Creatinine
male (15/group): 0, 5, 25,
0
-
-
-
100, 400 mg/kg-d; female
(15/group): 0, 5, 25,100,
5
12%
-5%
0%
400 mg/kg-d
25
1%
21%
-10%
daily for approximately
26 wk
100
4%
12%
-3%
400
8%
53%*
0%
Dose
(mg/kg-d)
Proteinuria incidence
Proteinuria severitvb
Urinarv casts
0
10/10
1.5
0/10
5
10/10
1.6
-
25
10/10
1.6
-
100
10/10
1.3
-
400
10/10
1.5
0/10
Female
Dose
Blood urea nitrogen
(mg/kg-d)
0
5
(BUN)
Cholesterol
Creatinine
-5%
-7%
-19%
25
-7%
-7%
-12%
100
-1%
-2%
-16%
400
4%
3%
-16%
Dose
(mg/kg-d)
Proteinuria incidence
Proteinuria severitvb
Urinarv casts
0
8/10
1.2
0/10
5
9/10
1.3
-
25
7/10
1.0
-
100
9/10
1.3
-
400
7/10
1.0
0/10
This document is a draft for review purposes only and does not constitute Agency policy.
1-14 DRAFT—DO NOT CITE OR QUOTE
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Toxicological Review ofETBE
Reference and study
Results
design
Saito etal. (2013); JPEC
Response relative to control
(2010b)
Male
rat, Fischer 344
inhalation - vapor
male (50/group): 0, 500,
1,500, 5,000 ppm (0, 2,090,
Dose
(mg/m3)
Blood urea
nitrogen
(BUN)
Cholesterol
Creatinine
Proteinuria
incidence
Proteinuria
severitvb
6,270, 20,900 mg/m3)a;
0
-
-
-
44/44
3.7
female (50/group): 0, 500,
1,500, 5,000 ppm (0, 2,090,
2,090
41%*
10%
14%*
38/38
3.5
6,270, 20,900 mg/m3)a
6,270
45%*
29%*
29%*
40/40
3.6
dynamic whole body
inhalation; 6 hr/d, 5 d/wk
20,900
179%*
52%*
71%*
31/31
3.6
for 104 wk; generation
Female
method, analytical
concentration, and method
reported
Dose
Blood urea
nitrogen
Proteinuria
Proteinuria
(mg/m3)
0
(BUN)
Cholesterol
Creatinine
incidence
33/38
severitvb
2.8
2,090
10%
-3%
0%
39/39
3.1
6,270
4%
-4%
0%
30/30
3.3
20,900
30%*
53%*
0%
30/30
3.4*
This document is a draft for review purposes only and does not constitute Agency policy.
1-15 DRAFT—DO NOT CITE OR QUOTE
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Toxicological Review ofETBE
Reference and study
Results
design
Suzuki etal. (2012); JPEC
Response relative to control
(2010a)
Male
rat, Fischer 344
oral - water
male (50/group): 0, 625,
2,500, 10,000 ppm (0, 28,
Blood urea
Dose
(mg/kg-d)
nitrogen
(BUN)
Cholesterol
Creatinine
Proteinuria
incidence
Proteinuria
severitvb
121, 542 mg/kg-d)c; female
0
-
-
-
39/39
3.0
(50/group): 0, 625, 2,500,
10,000 ppm (0, 46, 171,
28
3%
-11%
0%
37/37
3.1
560 mg/kg-d)c
121
20%*
10%
17%
34/34
3.1
daily for 104 wk
542
Female
43%*
Blood urea
31%*
17%
35/35
3.1
Dose
nitrogen
Proteinuria
Proteinuria
(mg/kg-d)
(BUN)
Cholesterol
Creatinine
incidence
severitvb
0
-
-
-
37/37
2.8
46
-8%
-2%
0%
37/37
3.0
171
-5%
12%
-17%
38/38
3.0
560
-5%
8%
0%
38/38
3.1
a4.18 mg/m3 = 1 ppm.
Severity of proteinuria = (1 x number of animals with "1+") + (2 x number of animals with "2+") + (3 x number of
animals with "3+") + (4 x number of animals with "4+") -f total number of animals in group.
Conversion performed by study authors.
*: result is statistically significant (p < 0.05) based on analysis of data by study authors.
for controls, no response relevant; for other doses, no quantitative response reported.
Percent change compared to controls calculated as 100 x [(treated value - control value) -f control value],
1 Table 1-4. Evidence pertaining to kidney tumor effects in animals following
2 exposure to ETBE
Reference and study design
Results
Hagiwara et al. (2011); JPEC (2008d)
rat, Fischer 344
oral -gavage
male (12/group): 0,1,000 mg/kg-d
daily for 23 wk
Male
Renal tubular
Dose Renal transitional adenoma or
(mg/kg-d) cell carcinoma carcinoma
0 0/12 0/12
1,000 0/12 0/12
This document is a draft for review purposes only and does not constitute Agency policy.
1-16 DRAFT—DO NOT CITE OR QUOTE
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Toxicological Review ofETBE
Reference and study design
Results
Hagiwara et al. (2011); JPEC (2008d)
Male
rat, Fischer 344
Renal tubular
oral -gavage
Dose
adenoma or
Renal transitional
male (30/group): 0, 300,1,000 mg/kg-d
(mg/kg-d)
carcinoma
cell carcinoma
daily for 23 wk following a 4-wk tumor
initiation by DMBDDa
0
11/30
1/30
300
6/30
0/30
1,000
13/30
2/30
Hagiwara et al. (2015)
Male
rat, Wistar
Renal tubular
oral - gavage
Dose
adenoma or
male (30/group): 0,100, 300, 500,
(mg/kg-d)
carcinomab
1,000 mg/kg-d
daily for 19 wk following a 2-wk tumor
0
18/30
initiation by N-ethyl-N-
100
23/30
hydroxyethylnitrosamine (EHEN)
300
500
1,000
25/30
26/30
26/30
Saito et al. (2013); JPEC (2010b)
Male
Female
rat, Fischer 344
inhalation - vapor
male (50/group): 0, 500,1,500,
Dose
(mg/m3)
Renal cell
carcinoma
Dose
(mg/m3)
Renal cell
carcinoma
5,000 ppm (0, 2,090, 6,270,
0
0/50
0
0/50
20,900 mg/m3)c; female (50/group): 0,
500, 1,500, 5,000 ppm (0, 2,090, 6,270,
2,090
1/50
2,090
0/50
20,900 mg/m3)c
6,270
0/49
6,270
0/50
20,900
0/50
20,900
0/50
Suzuki et al. (2012); JPEC (2010a)
Male
Female
rat, Fischer 344
oral - water
male (50/group): 0, 625, 2,500,
Dose
Renal cell
Dose
Renal cell
(mg/kg-d)
carcinoma
(mg/kg-d)
carcinoma
10,000 ppm (0, 28, 121, 542 mg/kg-d)d;
0
0/50
0
0/50
female (50/group): 0, 625, 2,500,
10,000 ppm (0, 46,171, 560 mg/kg-d)d
28
0/50
46
0/50
daily for 104 wk
121
0/50
171
0/50
542
1/50
560
1/50
aDiethylnitrosamine (DEN), N-butyl-N-(4-hydroxybutyl)nitrosamine (BBN), N-methyl-N-nitrosourea (MNU),
1,2-dimethylhydrazine dihydrochloride (DMH), and N-bis(2-hydroxypropyl)nitrosamine (DHPN).
bAuthors report significant trend.
c4.18 mg/m3 = 1 ppm.
dConversion performed by study authors.
This document is a draft for review purposes only and does not constitute Agency policy.
1-17 DRAFT—DO NOT CITE OR QUOTE
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Toxicological Review ofETBE
2 Table 1-5. Comparison of nephropathy and urothelial hyperplasia in
3 individual male rats from 2-year oral exposure flPEC. 2010al
CPN
Urothelial
hyperplasia
None
Minimal
Mild
Moderate
Marked
None
15
21
105
23
1
Minimal
0
0
17
16
2
Mild
0
0
0
0
0
Moderate
0
0
0
0
0
Marked
0
0
0
0
0
Spearman's rank correlation test (1-sided), p < 0.0001, rs = 0.36
10
11
12
Table 1-6. Comparison of nephropathy and urothelial hyperplasia in
individual male rats from 2-year inhalation exposure flPEC. 2010bl
CPN
Urothelial
hyperplasia
None
Minimal
Mild
Moderate
Marked
None
1
3
59
68
4
Minimal
0
0
14
29
21
Mild
0
0
0
0
0
Moderate
0
0
0
0
0
Marked
0
0
0
0
0
Spearman's rank correlation test (1-sided), p < 0.0001, rs = 0.36
This document is a draft for review purposes only and does not constitute Agency policy.
1-18 DRAFT—DO NOT CITE OR QUOTE
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Toxicological Review ofETBE
¦ = exposures at which the midpoint was reported statistically significant by studv authors
~ = exposures at whirl) the pndptiinl w;ts reported not statistical!) sigmhnml by si tidy juthoas
I iicrcuscd Absolute
Kidney Woi;"M
PO Male rat;16wks (B)
PO Female rat:;16wks (B)
PO Male r.rt; re productive (C)
PO Female rat;reproductive (C)
F1 Male ratjreprodurtive (C)
F1 Female rat;reproductive [C]
Male rat;23wks (0)
Male rat;2Swks [E)
Female rat;26wks (E)
O B-
Q S=
~—B—0
D¦ ¦ ¦
B—B-
-S -B-
_q b_
Male rat;104wks (A)
Incidence olTlironlc Feraa)e rat;104wks [A)
rrofjivssive
Nephropathy
Male 1 j|,10-iwks (K)
Female rat, 10 twks (F)
O——B~
Q-
-B-
i i n ™
~
-0
-E)
-E3
Male rat;104wks (A)
Average Severity of
Chronic Progressive Female rat;104wks (A)
Nephropathy
Male rat;104wks (F)
Ft male ratlU-1wks fF)
~ B-
-Q
HB-
Q 0~
Urothelial ilyperpla->ia
ol iIh- Renal I'elvh
Male r.it; J04vvk<; (]¦'
Female rat;104wks [F]
~ ¦-
O ~
-a
10 100
Dose (mg/fcg-day)
1,000 10,000
Sources (A) Cohen et .41, 2011 nvtitalystsot )I'FC 201flj, (B) Fuiilet J , 2010, jPRf, 2O01U', (0) G.iou.i, 2004b;
(Dj lUgm'ai a et Jl, 2011; (El MiyaUttal -11113, [PEC, 2008c, (F) Suzuki el .J , 2012; |PFf, 2(1 10j
Figure 1-4. Exposure-response array of kidney effects following oral exposure
to ETBE.
This document is a draft for review purposes only and does not constitute Agency policy.
1-19 DRAFT—DO NOT CITE OR QUOTE
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Toxicological Review ofETBE
¦ = exposures at which the endpoint was reported statistically significant by study authors
~ = exposures at which the endpoint was reported not statistically significant by study authors
Increased Absolute
Kidney Weight
Male rat;13wks (A]
Female rat;13wks (A)
Male rat;13wks, 28d recovery (A)
Female rat;13wks, 28d recovery (A)
Male rat; 13 wks (C)
Female rat; 13 wks (C)
Male mouse;13 wks (B)
Female mouse;13 wis (B)
-B-
Q-
-e-i a
a bh ~
Male rat;lG4wks (D)
Incidence of Chronic
Progressive Female rat;104wks (D)
Nephropathy
~ O-
O ID-
Male rat;104wks (D)
Average Severity of
Chronic Progressive Fema|e 104wks (D)
Nephropathy
a id-
id B-
Male rat;104wks (D)
Urothelial Hyperplasia
of the Renal Pelvis
Female rat;104wks (D)
~ a—i ~
100 1,000 10,000
Exposure Concentration (mg/m3)
100,000
Sources: (A) JPEG, 2008b; (B) Medinskyet al, 1999; Bond et al, 1996a (€) Medinsky et al,, 1999; Bond et al,
1996b CD) Saito et al, 2013; JPEC, 2010b
Figure 1-5. Exposure-response array of kidney effects following inhalation
exposure to ETBE.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Mode of Action Analysis—Kidney Effects
a) Toxicokinetic Considerations Relevant to Kidney Toxicity
ETBE is metabolized by cytochrome P450 (CYP) enzymes to an unstable hemiacetal that
decomposes spontaneously into tert-butanol and acetaldehyde (Bernauer etal.. 19981.
Acetaldehyde is metabolized further in the liver and is not thought to play a role in extrahepatic
toxicity. The main circulating breakdown product ofETBE metabolism is tert- butanol, which is
filtered from the blood by the kidneys and excreted in urine. Thus, following ETBE exposure, the
kidney is exposed to significant concentrations of tert-butanol, and kidney effects caused by tert-
butanol (described in the more detail in the draft IRIS assessment of tert-butanol) also are relevant
to evaluating the kidney effects observed after ETBE exposure. In particular, similar to ETBE, tert-
butanol has been reported to cause nephrotoxicity in rats, including effects associated with
a2u-globulin nephropathy. Unlike ETBE, however, increased renal tumors were reported following
chronic drinking water exposure to tert-butanol.
b) a?,,-Globulin-Associated Renal Tubule Nephropathy
One disease process to consider when interpreting kidney effects in rats is related to the
accumulation of a2U-globulin protein. a2U-Globulin, a member of a large superfamily of low-
molecular-weight proteins, was first characterized in male rat urine. Such proteins have been
detected in various tissues and fluids of most mammals (including humans), but the particular
isoform of a2U-globulin commonly detected in male rat urine is considered specific to that sex and
species. Exposure to chemicals that induce a2U-globulin accumulation can initiate a sequence of
histopathological events leading to kidney tumorigenesis. Because a2U-globulin-related renal tubule
nephropathy and carcinogenicity occurring in male rats are presumed not relevant for assessing
human health hazards (U.S. EPA. 1991a). evaluating the data to determine whether a2U-globulin
plays a role is important. The role of a2U-globulin accumulation in the development of renal tubule
nephropathy and carcinogenicity observed following ETBE exposure was evaluated using the U.S.
EPA fl991bl Risk Assessment Forum Technical panel report, AIphct2u-GIobuIin: Association with
Chemically Induced Renal Toxicity and Neoplasia in the Male Rat. This report provides specific
guidance for evaluating renal tubule tumors that are related to chemical exposure for the purpose
of risk assessment, based on an examination of the potential involvement of a2U-globulin
accumulation.
The hypothesized sequence of a2U-globulin renal tubule nephropathy, as described by U.S.
EPA fl991al. is as follows. Chemicals that induce a2U-globulin accumulation do so rapidly.
a2u-Globulin accumulating in hyaline droplets is deposited in the S2 (P2) segment of the proximal
tubule within 24 hours of exposure. Hyaline droplets are a normal constitutive feature of the
mature male rat kidney; they are particularly evident in the S2 (P2) segment of the proximal tubule
and contain a2U-globulin (U.S. EPA. 1991a). Abnormal increases in hyaline droplets have more than
one etiology and can be associated with the accumulation of different proteins. As hyaline droplet
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
deposition continues, single-cell necrosis occurs in the S2 (P2) segment, which leads to exfoliation
of these cells into the tubule lumen within 5 days of chemical exposure. In response to the cell loss,
cell proliferation occurs in the S2 (P2) segment after 3 weeks and continues for the duration of the
exposure. After 2 or 3 weeks of exposure, the cell debris accumulates in the S3 (P3) segment of the
proximal tubule to form granular casts. Continued chemical exposure for 3 to 12 months leads to
the formation of calcium hydroxyapatite in the papilla, which results in linear mineralization. After
1 or more years of chemical exposure, these lesions can result in the induction of renal tubule
adenomas and carcinomas (Figure 1-6).
U.S. EPA fl991al identified two questions that must be addressed to determine the extent
to which a2U-globulin-mediated processes induce renal tubule nephropathy and carcinogenicity.
First, whether the a2U-globulin process occurs in male rats and influences renal tubule tumor
development must be determined. Second, whether the renal effects in male rats exposed to ETBE
are due solely to the a2U-globulin process must be determined.
U.S. EPA fl991al stated that the criteria for answering the first question in the affirmative
are as follows:
1) hyaline droplets are larger and more numerous in treated male rats,
2) the protein in the hyaline droplets in treated male rats is a2U-globulin (i.e.,
immunohistochemical evidence), and
3) several (but not necessarily all) additional steps in the pathological sequence appear in
treated male rats as a function of time, dose, and progressively increasing severity
consistent with the understanding of the underlying biology, as described above, and
illustrated in Figure 1-6.
The available data relevant to this first question are summarized in Table 1-7, Table 1-8,
Figure 1-7, and Table 1-10, and are evaluated below.
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Male rat liver
Male rat: kidney
TBA binding
Sesorption- of poorly digestible
protein-chemical complex
£ 1 days
5 days - 48 weeks
Linear
S months
> 12 months
112 months
Renal adenoma,
carcinoma
Sustained cell
Synthesis of a^-globulin
Cell death and exfoliation
Hyaline droplet accumulation
within lysosomes
1 Source: Adapted from Swenbergand Lehman-McKeeman (1999); U.S. EPA (1991a).
2 Figure 1-6. Temporal pathogenesis of a2U-globulin-associated nephropathy
3 in male rats. a2U-Globulin synthesized in the livers of male rats is delivered to the
4 kidney, where it can accumulate in hyaline droplets and be retained by epithelial
5 cells lining the S2 (P2) segment of the proximal tubules. Renal pathogenesis
6 following continued exposure and increasing droplet accumulation can progress
7 stepwise from increasing epithelial cell damage, death, and dysfunction, leading to
8 the formation of granular casts in the corticomedullary junction, and linear
9 mineralization of the renal papilla, in parallel with carcinogenesis of the renal
10 tubular epithelium.
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Table 1-7. Additional kidney effects potentially relevant to mode of action in
animals exposed to ETBE
Reference and study design
Results
JPEC (2008b)
rat, CRL:CD(SD)
inhalation - vapor
male (10/group): 0,150, 500,1,500,
5,000 ppm (0, 627, 2,090, 6,270,
20,900 mg/m3)a; female (10/group): 0,150,
500, 1,500, 5,000 ppm (0, 627, 2,090, 6,270,
20,900 mg/m3)a
dynamic whole body chamber; 6 hr/d,
5 d/wk for 13 wk; generation method,
analytical concentration, and method
reported
Male
Incidence of
hyaline
droplets in
the proximal
Dose
tube
(mg/m3)
epithelium
0
0/10
627
3/10
2,090
8/10*
6,270
8/10*
20,900
8/10*
Unspecified representative samples reported as "weakly positive"
for ci2u-globulin in males; no hyaline droplets observed in proximal
tubule of females; hyaline droplets positive for a2u-globulin not
examined in females.
JPEC (2008c); Mivata et al. (2013)
rat, CRL:CD(SD)
oral -gavage
male (15/group): 0, 5, 25,100, 400 mg/kg-d;
female (15/group): 0, 5, 25,100,
400 mg/kg-d
daily for 180 d
Male
Dose
(mg/kg-d)
0
5
25
100
400
Incidence of
hyaline
droplets
0/15
0/15
0/15
4/15
10/15*
Female
Incidence of
hyaline
droplets
positive for
Qbu-globulin (mg/kg-d)
Dose
0/1
2/2
1/1
0
5
25
100
400
Incidence
of hyaline
droplets
0/15
0/15
Medinskv et al. (1999); Bond et al. (1996b)
rat, Fischer 344
inhalation - vapor
male (48/group): 0, 500,1,750, 5,000 ppm
(0, 2,090, 7,320, 20,900 mg/m3)a; female
(48/group): 0, 500, 1,750, 5,000 ppm
(0, 2,090, 7,320, 20,900 mg/m3)a
dynamic whole body chamber; 6 hr/d,
5 d/wk for 13 wk; generation method,
analytical concentration, and method
reported
Male
Dose
(mg/m3
Proximal tubule proliferation
Hyaline droplet
severity
1.8
1 week
4 weeks 13 weeks
2,090
3.0
39%
24%
137%*
7,320
3.2
23%
-14%
274%*
20,900
3.8
102%*
175%*
171%*
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Reference and study design
Results
Female
Dose
(mg/m3)
0
2,090
7,320
20,900
1 week
60%*
88%*
49%*
Proximal tubule proliferation
4 weeks
3%
15%
31%*
13 weeks
73%
64%
47%
Saito et al. (2013); JPEC (2010b)
rat, Fischer 344
inhalation - vapor
male (50/group): 0, 500,1,500, 5,000 ppm
(0, 2,090, 6,270, 20,900 mg/m3)a; female
(50/group): 0, 500,1,500, 5,000 ppm
(0, 2,090, 6,270, 20,900 mg/m3)a
dynamic whole body inhalation; 6 hr/d,
5 d/wk for 104 wk; generation method,
analytical concentration, and method
reported
Male
No hyaline droplets observed.
Female
No hyaline droplets observed.
Suzuki et al. (2012); JPEC (2010a)
rat, Fischer 344
oral - water
male (50/group): 0, 625, 2,500,10,000 ppm
(0, 28,121, 542 mg/kg-d)b; female
(50/group): 0, 625, 2,500, 10,000 ppm (0, 46,
171, 560 mg/kg-d)b
daily for 104 wk
Male
No hyaline droplets observed.
Female
No hyaline droplets observed.
a4.18 mg/m3 = 1 ppm.
Conversion performed by study authors.
*: result is statistically significant (p < 0.05) based on analysis of data by study authors.
-: for controls, no response relevant; for other doses, no quantitative response reported.
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1 Table 1-8. Summary of data informing whether the a2u-globulin process is
2 occurring in male rats exposed to ETBE
Criterion
Duration
Results
Reference
(1) hyaline droplets are increased
in size and number
1 wk
(+)a
Medinskv et al. (1999)
4 wk
(+)a
Medinskv et al. (1999)
13 wk
(+)a
Medinskv et al. (1999)
13 wk
+
JPEC (2008b)
26 wk
+
Mivata et al. (2013); JPEC (2008c)
104 wk
-
Suzuki et al. (2012)
104 wk
-
Saito et al. (2013); JPEC (2010b)
(2) the protein in the hyaline
droplets is a2u-globulin
1 wk
(+)b
JPEC (2008b)
4 wk
(+)b
Medinskv et al. (1999)
13 wk
(+)b
Medinskv et al. (1999)
13 wk
(+)b
JPEC (2008b)
26 wk
(+)c
Mivata et al. (2013); JPEC (2008c)
(3) Several (but not necessarily all) additional steps in the pathological sequence are present in male rats, such as:
(a) single-cell necrosis
13 wk
-
JPEC (2008b)
13 wk
-
Medinskv et al. (1999)
26 wk
-
Mivata et al. (2013); JPEC (2008c)
104 wk
-
Suzuki et al. (2012); JPEC (2010a)
104 wk
-
Saito et al. (2013); JPEC (2010b)
(b) exfoliation of epithelial cells
into the tubular lumen
13 wk
-
JPEC (2008b)
13 wk
-
Medinskv et al. (1999)
26 wk
-
Mivata et al. (2013); JPEC (2008c)
104 wk
-
Suzuki et al. (2012); JPEC (2010a)
104 wk
-
Saito et al. (2013); JPEC (2010b)
(c) granular casts
13 wk
-
JPEC (2008b)
13 wk
(+)
Cohen et al. (2011); JPEC 2007a
13 wk
-
Medinskv et al. (1999)
26 wk
-
Mivata et al. (2013); JPEC (2008c)
104 wk
-
Suzuki et al. (2012); JPEC (2010a)
104 wk
-
Saito et al. (2013); JPEC (2010b)
(d) linear mineralization of tubules
in the renal papilla
13 wk
-
JPEC (2008b)
13 wk
-
Medinskv et al. (1999)
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Criterion
Duration
Results
Reference
26 wk
-
Mivata et al. (2013); JPEC (2008c)
104 wk
+
Suzuki et al. (2012); JPEC (2010a), Cohen et al. (2011)
104 wk
+
Saito et al. (2013); JPEC (2010b)
(e) Proliferation and foci of tubular
hyperplasia
13 wk
-
JPEC (2008b)
13 wk
+/-d
Medinskv et al. (1999)
26 wk
-
Mivata et al. (2013); JPEC (2008c)
104 wk
-
Suzuki et al. (2012); JPEC (2010a)
104 wk
-
Saito et al. (2013); JPEC (2010b)
1 + = Statistically significant change reported in one or more treated groups.
2 (+) = Effect reported in one or more treated groups, but statistics not reported.
3 - = No statistically significant change reported in any of the treated groups.
4 aDroplet severity.
5 bUnspecified "representative samples" examined.
6 Three samples from highest two dose groups examined.
7 dLabeling index statistically significantly increased, but no hyperplasia reported.
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¦ = exposures at which the endpoint was reported statistically significant by study authors
~ = exposures at which the endpoint was reported not statistically significant by study authors
% = effect was observed but statistics not reported
+ = unspecified representative samples reported positive for a2u-globulin
Miyata et a I, 2013; JPEC, 2008c - 26wks
Accumulation of
hyaline droplets
Suzutd et al, 2012; JPEC, 2010a - 104wks
In
hyaline Miyata et al, 2013; JPEC, 2008c - 26wks
droplets
Granular
casts/dilation
Cohen et al, 2011 -13 wks
Miyata et at, 2013; JPEC, 2008c - 26wks
Suzuki ct al, 2012; JPEC, 2010a - 104wb
Miyata et al, 2013; JPEC, 2008c - 26wks
Linear
papillary
mineralization
Suzuki et al, 2012; JPEC, 2010a - 104wte ¦
Tubular
hyperplasia
Suzuki et al, 2012; JPEC, 2010a - MMwfc
Renal
adenoma
or
carcinoma
Suzuki el al, 2012; JPEC, 2010a • KMwfc
Q B S3 El
-b a
-e b
Q-
n i i
0 B—— G
1 10 100 1,000
Dose (mg/kg-day)
Figure 1-7. ETBE oral exposure array of a2u-globulin data in male rats.
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¦ = exposures at which the endpoint was reported statistically significant by study authors
~ = exposures at which the endpoint was reported not statistically significant by study authors
• = effect was observed but statistics not reported
+ = unspecified representative samples reported positive for ct2u-globulin
Medinsky et al., 1999; Bond et a], 1996 - lwk
Medinsky et al., 1999; Bond ot al, 1996 - 4wk -
Accumulation
of hyaline Medinsky et al., '1999; Bond et al, 1996 ¦ 13wk
droplets
JPEG, 2008b -13wk ¦
Saito et al., 2013; JPEC, 2010b -104wk
B—
• •-
• •-
• •-
¦ ¦
B B—
•
•
•
¦
B
Medimdy et al., 1999; Bond et al, "1996 - lwk
a globulin in Medinsky et al, 1999; Bond et al, 1996 - 4wk
hyaline
droplets Medinsky et at, 1999; Bond et al, 1996 - 13wk
J PEC, 2008b - *13 wk
4—
it 11:
T T IT
Medinsky et al., 1999; Bond et al, 1996 - 13wk
Granular
casts/dilation ) PEC, 2008b-13 wk
Saito et al., 2013; JPEC, 2010b - 104wk
B—
B B-
B B—
B B—
—~
B
B
Medinsky et al., 1999; Bond et al* 1996 - 13wk
Linear
papillary jpec, 2008b - I3wk •
mineralization
Saito et al., 2013; JPEC, 2010b - 104wk
B—
B B-
B B—
B B—
B
B
¦
Tubular Saito et al, 20*13; [PEC, 2010b - 104wk
hyperplasia
B B—
~
-
Renal adenoma Saito et al, 2013; JPEC, 2010b - 104wk -
or carcinoma
B B—
B
100 1,000 10,000 *100,000
Exposure Concentration (mg/m3)
1 Figure 1-8. ETBE inhalation exposure array of a2u-globulin data in male rats.
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Question One: Is the
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incidence of papillary mineralization was increased statistically significantly in both 2-year studies.
Papillary mineralization increased in a dose-related manner following oral ETBE exposure in male
rats at concentrations of 0, 28,121, and 542 mg/kg-day, respectively (Suzuki etal.. 2012: TPEC.
2010a), and in males at ETBE inhalation concentrations of 0, 2,090, 6,270, and 20,900 mg/m3 fSaito
etal.. 2013: TPEC. 2010b). Hyaline droplet deposition was observed at a similar frequency as
mineralization following oral ETBE exposure (Mivata etal.. 2013: Suzuki etal.. 2012: TPEC. 2010a.
2008c); however, hyaline droplet deposition was observed in 80% of animals at all three inhalation
exposure concentrations flPEC. 2008bl compared with mineralization rates of 0, 2, and 12%
(lowest to highest exposure concentration) fSaito etal.. 2013: TPEC. 2010bl. A detailed evaluation
and analysis of all the evidence relevant to this criterion follows.
Detailed evaluation of the available evidence supporting the third criterion
a) Single cell death, exfoliation into the renal tubules, and necrosis were not observed in
any study flPEC. 2008b. c; Medinskv etal.. 19991. This observation might not be
inconsistent with the hypothesized MOA because cell death and exfoliation could occur
as early as 5 days post exposure, peak at 3 weeks, and then decline to near background
levels by 4-5 weeks (Kanerva etal.. 1987): this endpoint was not examined in any study
evaluating ETBE exposures less than 13 weeks. Thus, the lack of exfoliation
observations could be the result of both weak induction of a2u-globulin and a lack of
appropriately timed examinations.
b) Granular cast formation was observed in one study. The TPEC f2007al study reported
that, at 13 weeks, granular casts were observed in high-dose males, while none were
observed in controls (no statistical tests performed). Other studies at similar time
points did not report the presence of granular casts (TPEC. 2008b. c; Medinskv etal..
1999) despite using similar exposure concentrations. Granular cast formation, however,
might not occur with weak inducers of a2u-globulin (Short etal.. 1986). which is
consistent with the weak staining of a2u-globulin, as discussed above (TPEC. 2008b).
c) Linear mineralization of tubules within the renal papilla was consistently observed in
male rats after 2 years (Saito etal.. 2013: Suzuki etal.. 2012). This lesion typically
appears at chronic time points, occurring after exposures of 3 months up to 2 years (U.S.
EPA. 1991a").
d) Cellular proliferation was increased after 1, 4, and 13 weeks in males and females;
however, the magnitude of effect was reduced in females compared to males.
Observation of proliferation in both sexes suggests that this effect is not male specific,
and thus not a2U-globulin specific. Furthermore, renal tubule hyperplasia was not
observed in any 2-year study, suggesting that ETBE does not induce sustained
proliferation (Saito etal.. 2013: Suzuki etal.. 2012). Renal tubule hyperplasia is the
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preneoplastic lesion associated with a2U-globulin nephropathy in chronic exposures that
leads to renal tubule tumors fU.S. EPA. 1991a 1.
The progression of histopathological lesions for a2U-globulin nephropathy is predicated on
the initial response of excessive hyaline droplet accumulation (containing a2u-globulin) leading to
cell necrosis and cytotoxicity, which in turn cause the accumulation of granular casts, linear
mineralization, and tubular hyperplasia resulting from sustained cellular proliferation. Therefore,
observations of temporal and dose-response concordance for these effects are informative for
drawing conclusions on causation.
As mentioned above (see Table 1-8), some steps in the sequence of a2U-globulin
nephropathy are observed at the expected time points following exposure to ETBE. Accumulation of
hyaline droplet severity was observed early, at 1 week following inhalation exposure (Medinskv et
al.. 19991. and increased incidence was subsequently observed at 90 days flPEC. 2008b] or 26
weeks flPEC. 2008cl: a2u-globulin was identified as the protein in these droplets fBorghoffetal..
2001: Williams and Borghoff. 20011. Lack of necrosis and exfoliation might be due to the weak
induction of a2U-globulin and a lack of appropriately timed examinations. Granular cast formation
was reported in one oral study fCohen etal.. 20111. while three other oral and inhalation studies
reported none (TPEC. 2008b. c; Medinskv etal.. 19991. which also could indicate weak a2U-globulin
induction. Observations of the subsequent linear mineralization of tubules fall within the expected
timeframe of the appearance of these lesions. Neither a2U-globulin-mediated regenerative cell
proliferation nor atypical renal tubule hyperplasia were observed. Overall, no explicit
inconsistencies are present in the temporal appearance of the histopathological lesions associated
with the a2u-globulin nephropathy induced following ETBE exposure; however, the data set would
be bolstered by measurements at additional time points to lend strength to the MOA evaluation.
Hyaline droplets were weakly induced in all male rats in the 13-week inhalation studies
(TPEC. 2008b: Medinskv etal.. 19991. which did not result in increased linear mineralization at the
corresponding doses. The lack of increased linear mineralization at low doses also is consistent
with weak induction of hyaline droplets.
Overall, the histopathological sequence has numerous data gaps, such as the lack of
observable necrosis, cytotoxicity, and tubule hyperplasia at stages plausibly within the timeframe
of detectability. Therefore, the number of histopathological steps observed was insufficient to fulfill
the third criterion.
Summary and conclusions for question one
The evidence suggests that ETBE causes hyaline droplets to increase in size and number.
The documentation of a2U-globulin staining is poor and provides weak evidence of a2U-globulin in
the hyaline droplets. Only one of the additional steps in the pathological sequence was consistently
observed (linear papillary mineralization), and the ETBE database lacks evidence of renal tubule
hyperplasia and adenomas or carcinomas, despite multiple studies, exposure routes, and durations
This document is a draft for review purposes only and does not constitute Agency policy.
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ranging from 13 weeks to 2 years. Overall, the available data were insufficient to conclude that the
a2u-globulin process is operative.
Comparison ofETBE and tert-butanol a2U-gIobuIin data
Both EPA and IARC have accepted the biological plausibility of the a2U-globulin-mediated
hypothesis for inducing nephropathy and cancer in male rats fSwenberg and Lehman-McKeeman.
1999: U.S. EPA. 1991al. and those rationales will not be repeated here. A more recent retrospective
analysis indicating that several steps in the sequence of pathological events are not required for
tumor development has demonstrated this by evaluating several a2U-globulin-inducing chemicals
which fail to induce many of the pathological sequences in the a2U-globulin pathway fDoi et al..
20071. For instance, dose-response concordance was not observed for several endpoints such as
linear mineralization, tubular hyperplasia, granular casts, and hyaline droplets following exposure
to chemicals that induce the a2U-globulin process such as d-limonene, decalin, propylene glycol
mono-t-butyl ether, and Stoddard Solvent IICA (SS IICA). Although some of these chemicals induced
dose-response effects for a few endpoints, all failed to induce a dose-response for at all of the
endpoints in the sequence. Furthermore, no endpoint in the pathological sequence was predictive
for tumor incidence when considering either the dose responsiveness or the severity. Tumor
incidence was not affected in a dose-related manner following either d-limonene or decalin
exposure. Tumor incidence was not correlated with the severity of any one effect in the a2U-globulin
sequence as demonstrated by SS IICA, which induced some of the most severe nephropathy relative
to the other chemicals, but did not significantly increase kidney tumors fDoi etal.. 20071. Thus, this
analysis suggests that another MOA could be operative for inducing kidney tumors in male rats.
As described above, ETBE is metabolized to tert-butanol, so kidney data following
tert-butanol exposure also are potentially relevant to evaluating the MOA of ETBE. In particular, the
effects of tert-butanol on the a2U-globulin process are relevant for evaluating the coherence of the
available data on ETBE-induced nephropathy.
Hyaline droplet deposition and linear mineralization were both observed following similar
exposure durations to tert-butanol and ETBE. After 13 weeks of exposure to tert-butanol or ETBE,
hyaline droplets were dose-responsively increased. ETBE exposure increased hyaline droplets at
lower internal concentrations of tert-butanol than did direct tert-butanol administration.
Tubule hyperplasia and renal tumors were both observed following 2-year exposure to
tert-butanol but not to ETBE, despite similar internal concentrations of tert-butanol following ETBE
exposure fSaito etal.. 2013: Suzuki etal.. 2012: TPEC. 2010bl. Similarly, the incidence of renal
tumors was increased at internal concentrations of tert-butanol that were achieved in two separate
ETBE studies. The failure ofETBE to induce several histopathological lesions in the a2U-globulin
pathological sequence at similar internal tert-butanol concentrations as those that induced
hyperplasia and tumorigenesis following exposure to tert-butanol directly suggests a lack of
coherence across the two data sets.
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c) Chronic Progressive Nephropathy
Exacerbation of CPN has been proposed as another rat-specific mechanism of
nephrotoxicity that is not relevant to humans (Hard etal.. 20091. CPN is an age-related renal
disease that occurs in rats of both sexes (NTP. 2015. 2014: Hard etal.. 2013: Melnick etal.. 2012:
U.S. EPA. 1991al. CPN is more severe in males than in females and is particularly common in the
Sprague-Dawley and Fischer 344 strains. Dietary and hormonal factors play a role in modifying
CPN, though its etiology is largely unknown.
CPN has been suggested as a key event in the onset of renal tubule tumors, and a sequence
of key events in the MOA is as follows: (1) metabolic activation, (2) chemically exacerbated CPN, (3)
increased tubule cell proliferation, (4) tubule hyperplasia, and (5) adenomas (Hard etal.. 20131.
Arguments against this MOA also have been proposed (Melnick et al.. 20121. ETBE exposure
increased CPN severity following 2-year inhalation and 13-week oral exposure, but did not affect
tubule hyperplasia or increase renal tubule tumor incidence. Thus, the CPN-mediated cancer MOA
proposed by Hard et al. (2013.; 20091 is not operative for ETBE.
Additional markers associated with CPN include elevated proteinuria and albumin in the
urine and increased BUN, creatinine, and cholesterol in the serum, of which proteinuria is the major
urinary effect and a very sensitive measure of CPN (Hard etal.. 20091. In the case of ETBE exposure,
however, increased severity or incidence of proteinuria was not correlated with increased severity
of CPN in male rats possibly due to high background severity of CPN. In female rats, background
severity of CPN was much milder, thus increased proteinuria was observable only when CPN was
increased as in the 2-year inhalation exposure study fSaito etal.. 20131. Elevated BUN and
creatinine typically are not observed until very late in CPN progression. This was true for ETBE, as
most of these markers were elevated only after 2-year exposures.
Several of the CPN pathological effects are similar to—and can obscure the lesions
characteristic of—a2U-globulin-related hyaline droplet nephropathy (Webb etal.. 19901.
Additionally, renal effects of a2U-globulin accumulation can exacerbate the effects associated with
CPN fU.S. EPA. 1991al.
CPN often is more severe in males than in females, which was observed to be the case with
ETBE. Increased severity of CPN was reported in both male and female rats due to ETBE exposure,
but these increases were statistically significant only in the highest exposure groups of both sexes
following chronic inhalation. Some of the observed renal lesions in male rats following exposure to
ETBE are effects commonly associated with CPN. A strong, statistically significant, treatment-
related relationship was observed between chronic ETBE exposure and increased incidence of
urothelial hyperplasia in male rats in both the inhalation and oral studies fSaito etal.. 2013: Suzuki
etal.. 2012: TPEC. 2010a. b). Urothelial hyperplasia is both increased by dose and weakly correlated
with CPN, which is also dose-related (Table 1-5 andTable 1-6). Thus, disentangling the
contributions of dose and nephropathy in the development of urothelial hyperplasia is not possible
with the currently available information. Moreover, no evidence is available to support that
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Toxicological Review ofETBE
urothelial hyperplasia is independent ofETBE treatment, given the robust dose-response
relationships. Therefore, the data are insufficient to dismiss urothelial hyperplasia as causally
related to ETBE exposure.
Finally, because tert- butanol is a major metabolite ofETBE and both chemicals induce
similar noncancer kidney effects, tert-butanol could be the active toxic moiety responsible for these
effects. The three noncancer kidney endpoints (kidney weights, urothelial hyperplasia, CPN) were
evaluated on an internal dose basis to compare these data from ETBE and tert-butanol studies
fSalazar etal.. 20151. The results demonstrate that noncancer kidney effects, including kidney
weight changes, urothelial hyperplasia, and exacerbated CPN, yielded consistent dose-response
relationships across routes of exposure and across ETBE and tert-butanol studies using tert-butanol
blood concentration as the dose metric. These results are consistent with the hypothesis that tert-
butanol mediates the noncancer kidney effects following ETBE administration.
Overall Conclusion on MOA for Kidney Effects
ETBE increases a2U-globulin deposition and hyaline droplet accumulation in male rat
kidneys, but only one of the five additional steps in the pathological sequence (linear
mineralization) was consistently observed (see Table 1-8). These data are insufficient to conclude
that ETBE induces a2U-globulin nephropathy. CPN and the exacerbation of CPN could play a role in
renal tubule nephropathy, although several endpoints indicate that urothelial hyperplasia and
increased kidney weights related to ETBE exposure cannot be entirely explained by the a2U-globulin
or CPN processes. Collectively, the evidence indicates other, unknown processes contribute to renal
nephrotoxicity.
Integration of Kidney Effects
Kidney effects (increases in severity of nephropathy, blood biomarkers, hyaline droplets,
linear mineralization, urothelial hyperplasia, and kidney weight) were observed across multiple
studies, predominantly in male and female rats; chronic bioassays found no treatment-related
increases in renal tumors. The available evidence indicates that multiple processes induce the
noncancer kidney effects. CPN is a common and well-established constellation of age-related lesions
in the kidney of rats, and there is no known counterpart to CPN in aging humans. However, CPN is
not a specific diagnosis on its own but an aggregate term describing a spectrum of effects, employed
to reduce the time and effort required to grade each component of the disease. The individual
lesions associated with CPN (tubular degeneration, glomerular sclerosis, etc.) also occur in the
human kidney. Thus, exacerbation of one or more of these lesions may reflect a type of injury
relevant to the human kidney.
Some endpoints in male rats (hyaline droplets, linear mineralization) are components of the
a2u-globulin process. U.S. EPA f!991al states that, if the a2U-globulin process were occurring in male
rats, the renal tubule nephropathy associated with this process in male rats would not be relevant
to humans for purposes of hazard identification. In the case ofETBE exposure, for which the
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available data were insufficient to conclude that the a2U-globulin process is operative, the
characterization of human health hazard for noncancer kidney toxicity relied on effects weakly
associated with CPN or typically observed with the a2U-globulin-process in male rats.
Several noncancer endpoints that were concluded to result from ETBE exposure
independent of a2U-globulin are appropriate for consideration of a kidney hazard. These effects are
change in absolute kidney weights, urothelial hyperplasia, and increased blood biomarkers in male
and female rats, with the effects in males tending to be stronger than in females. Noncancer kidney
effects yielded consistent dose-response relationships using tert-butanol blood concentration as the
dose metric, consistent with the hypothesis that tert-butanol mediates the noncancer kidney effects
following ETBE administration. Based on dose-related increases in these noncancer endpoints in
rats, kidney effects are a potential human hazard of tert-butanol exposure. The hazard and dose-
response conclusions regarding these noncancer endpoints associated with ETBE exposure are
discussed further in Section 1.3.1.
1.2.2. Liver Effects
Synthesis of Effects in Liver
This section reviews the studies that investigated whether exposure to ETBE can cause liver
noncancer or cancer effects in humans or animals. The database for ETBE-induced liver effects
includes nine studies conducted in animals, all but two of which were performed in rats. A
description of the studies comprising the database is provided in Section 1.2.1. Briefly, exposures
ranged from 13 weeks to 2 years and both inhalation and oral exposure routes are represented.
Studies using short-term and acute exposures that examined liver effects are not included in the
evidence tables; however, they are discussed in the text if they provide data informative of MOA or
hazard identification. Studies are arranged in evidence tables first by effect and then in alphabetical
order by author. The design, conduct, and reporting of each study were reviewed, and each study
was considered adequate to provide information pertinent to this assessment
Liver weight. Several factors associated with the 2-year organ weight data confound
consideration for hazard identification. As mentioned previously in the discussion of kidney effects,
mortality was a confounding factor in 2-year studies. In addition, proliferative lesions (altered
hepatocellular foci) were observed in rat livers, especially males, in both 2-year oral and inhalation
studies, which further complicates interpretation of changes in organ weight. Furthermore,
inhalation exposure significantly increased liver adenomas and carcinomas in male rats at the
highest dose, corresponding to increased liver weights in those dose groups fSaito etal.. 2013:
IPEC. 2010b], Collectively, these observations preclude including 2-year liver weight data for
hazard identification. Organ weight data obtained from studies of shorter duration, however, are
not confounded by these age-associated factors (e.g., tumors, mortality) and therefore could be
appropriate for hazard identification.
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Toxicological Review ofETBE
Chronic and subchronic studies by both oral and inhalation routes reported consistent,
statistically significant, dose-related increases in liver weights (see Figure 1-9, Figure 1-10, Table
1-9). Liver weight and body weight have been demonstrated to be proportional, and liver weight
normalized to body weight was concluded to be optimal for data analysis (Bailey etal.. 20041: thus,
only relative liver weight is considered in the determination of hazard. Relative liver weights were
consistently increased at similar exposure concentrations in four of five studies for males and three
of four studies for females; however, statistically significant increases often occurred only at the
highest tested concentration with increases in relative liver weight ranging from 17 to 27% in
males and 8 to 18% in females. Relative liver weights in rats were increased at only the highest
dose following oral exposures of 16 weeks or longer (Mivata etal.. 2013: Fuiii etal.. 2010: TPEC.
2008c: Gaoua. 2004b). In utero exposure yielded similar effects on F1 liver weights, in terms of the
magnitude of percent change, from adult exposure (Gaoua. 2004b). Inhalation exposure increased
liver weight at the highest dose in female rats, but not in males, following 13-week exposure flPEC.
2008b). Following a 28-day recovery period, male but not female liver weights were increased
fTPEC. 2008bl. Short-term studies observed similar effects on liver weight fTPEC. 2008a: White et
al.. 19951
Liver histopathology. Centrilobular hypertrophy and acidophilic and basophilic focal
lesions were the only dose-related types of pathological lesions observed in the liver. Centrilobular
hypertrophy was inconsistently increased throughout the database, but also was observed at the
same concentrations that induced liver weight changes in rats of both sexes after 13-week
inhalation and 26-week oral exposures (see Table 1-10; Figure 1-9, Figure 1-10). A 26-week oral
gavage study f Mivata etal.. 2013: TPEC. 2008cl in rats and three 13-week inhalation studies in mice
and rats fWeng etal.. 2012: TPEC. 2008b: Medinskv etal.. 19991 demonstrated a statistically
significant increase in centrilobular hypertrophy at the highest dose, but 2-year oral or inhalation
studies in rats reported no changes in centrilobular hypertrophy following ETBE exposure,
suggesting a transient effect
Acidophilic and basophilic preneoplastic lesions were increased in male rats, but not
female, at the highest tested dose following a 2-year inhalation exposure to ETBE fSaito etal.. 2013:
TPEC. 2010bl. Following 2-year drinking water exposure to ETBE, an increasing, but not statistically
significant, trend in basophilic preneoplastic lesions was observed in the liver of male rats, while
incidence of these lesions decreased in female rats (Suzuki etal.. 2012: TPEC. 2010a).
Serum liver enzymes. Serum liver enzymes were inconsistently affected across exposure
routes (see Table 1-11; Figure 1-9, Figure 1-10). No enzyme levels were affected in studies of
exposure durations less than 2 years fMivata etal.. 2013: TPEC. 2008bl. Gamma-glutamyl
transpeptidase (GGT) was significantly increased in male rats at one intermediate dose following
oral exposure and the two highest doses following inhalation exposure in 2-year studies fTPEC.
2010a. b). GGT was not significantly affected in female rats in any study. No consistent dose-related
changes were observed in aspartate aminotransferase (AST), alanine aminotransferase (ALT), or
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Toxicological Review ofETBE
alkaline phosphatase (ALP) liver enzymes following either oral or inhalation exposure of any
duration. Serum liver enzyme levels were not temporally consistent with hypertrophy or liver
weight effects, and changes were observed only following 2-year exposure. With the exception of a
dose-related increase in serum GGT in male rats and an increase in AST at the highest dose in
females, no other dose-related changes in liver enzyme levels were observed that were
directionally consistent with the liver weight and hypertrophy effects.
Liver tumors. Data on liver tumor induction by ETBE are presented in Table 1-12. Liver
adenomas or carcinomas (combined) were increased in male F344 rats, but not in females,
following 2-year inhalation exposure fSaito etal.. 2013: TPEC. 2010bl No significant increase in
tumors was observed following 2-year oral exposure (Suzuki etal.. 2012: TPEC. 2010a: Maltoni et
al.. 1999). Acidophilic and basophilic focal lesions increased following a similar exposure duration,
route, and concentration as were used for the increased tumors. Two-stage "initiation, promotion"
studies in male F344 and Wistar rats administered mutagens for 2-4 weeks reported statistically
significant increases in liver adenomas, carcinomas, or total neoplasms after 19-23 weeks ofETBE
exposure via oral gavage f Hagiwara etal.. 2015: Hagiwara etal.. 20111. Liver tumors were not
observed in male F344 rats exposed to ETBE for 23 weeks without prior mutagen exposure
(Hagiwara etal.. 2011). while liver tumorigenesis without prior mutagen exposure was not
evaluated in Wistar rats (Hagiwara etal.. 2015).
Table 1-9. Evidence pertaining to liver weight effects in animals exposed to
ETBE
Reference and study design
Results
Fuiiietal. (2010): JPEC (2008e)
Response relative to control
rat, Sprague-Dawley
P0, Male
P0, Female
oral - gavage
PO, male (24/group): 0,100, 300,1,000 mg/kg-d
daily for 16 wk beginning 10 wk prior to mating
Dose
(mg/kg-d)
Relative
weight
Dose
(mg/kg-d)
Relative
weight
P0, female (24/group): 0,100, 300,1,000 mg/kg-d
0
-
0
-
daily for 17 wk beginning 10 wk prior to mating to
lactation day (LD) 21
100
1%
100
-1%
300
2%
300
3%
1,000
21%*
1,000
9%*
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Toxicological Review ofETBE
Reference and study design
Results
Gaoua(2004b)
Response relative to control
rat, Sprague-Dawley
oral -gavage
P0, Male
P0, Female
PO, male (25/group): 0, 250, 500,1,000 mg/kg-d
Dose
(mg/kg-d)
Relative
weight
Dose
(mg/kg-d)
Relative
weight
daily for a total of 18 wk beginning 10 wk before
mating until after weaning of the pups
P0, female (25/group): 0, 250, 500,1,000 mg/kg-d
0
-
0
-
daily for a total of 18 wk beginning 10 wk before
mating until PND 21
250
3%
250
10%
Fl, male (25/group): 0, 250, 500,1,000 mg/kg-d
500
6%
500
8%
P0 dams dosed daily through gestation and
lactation, then Fl doses beginning PND 22 until
1,000
24%*
1,000
4%
weaning of the F2 pups
Fl, Male
Fl, Female
Fl, female (24-25/group): 0, 250, 500,
1,000 mg/kg-d
P0 dams dosed daily through gestation and
Dose
(mg/kg-d)
Relative
weight
Dose
(mg/kg-d)
Relative
weight
lactation, then Fl dosed beginning PND 22 until
0
-
0
-
weaning of F2 pups
250
0%
250
3%
500
11%*
500
6%
1,000
25%*
1,000
9%*
Hagiwara et al. (2011); JPEC (2008d)
Response relative to control
rat, Fischer 344
Male
oral - gavage
male (12/group): 0,1,000 mg/kg-d
daily for 23 wk
Dose
(mg/kg-d)
0
1,000
Relative
weight
27%*
JPEC (2008b)
Response relative to control
rat, CRL:CD(SD)
Male
Female
inhalation - vapor
Dose
(mg/m3)
Relative
weight
Dose
(mg/m3)
Relative
weight
male (NR): 0, 150, 500, 1,500, 5,000 ppm (0, 627,
2,090, 6,270, 20,900 mg/m3)b; female (NR): 0,150,
500, 1,500, 5,000 ppm (0, 627, 2,090, 6,270,
0
-
0
-
20,900 mg/m3)
dynamic whole body chamber; 6 hr/d, 5 d/wk for
627
5%
627
4%
13 wk; generation method, analytical
2,090
5%
2,090
-1%
concentration, and method reported
6,270
5%
6,270
6%
20,900
10%
20,900
18%*
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Toxicological Review ofETBE
Reference and study design
Results
JPEC (2008b)
Response relative to control
rat, CRL:CD(SD)
Male
Female
inhalation - vapor
Dose
(mg/m3)
Relative
weight
Dose
(mg/m3)
Relative
weight
male (6/group): 0, 5,000 ppm (0, 20,900 mg/m3)b;
female (6/group): 0, 5,000 ppm (0,
20,900 mg/m3)b
0
-
0
-
dynamic whole body chamber; 6 hr/d, 5 d/wk for
13 wk followed by a 28-d recovery period;
20,900
9%*
20,900
7%
generation method, analytical concentration, and
method reported
Mivata et al. (2013); JPEC (2008c)
Response relative to control
rat, CRL:CD(SD)
Male
Female
oral -gavage
male (15/group): 0, 5, 25,100, 400 mg/kg-d;
female (15/group): 0, 5, 25,100, 400 mg/kg-d
Dose
(mg/kg-d)
Relative
weight
Dose
(mg/kg-d)
Relative
weight
daily for 26 wk
0
-
0
-
5
5%
5
1%
25
7%
25
1%
100
9%
100
4%
400
17%*
400
12%*
1 Conversion performed by study authors.
2 b4.18 mg/m3 = 1 ppm.
3 NR: not reported; *: result is statistically significant (p < 0.05) based on analysis of data by study authors.
4 for controls, no response relevant; for other doses, no quantitative response reported.
5 Percent change compared to controls calculated as 100 x [(treated value - control value) -f control value],
6 Table 1-10. Evidence pertaining to liver histopathology effects in animals
7 exposed to ETBE
Reference and study design
Results
Gaoua(2004b)
P0, Male
P0, Female
rat, Sprague-Dawley
oral - gavage
P0, male (25/group): 0, 250, 500,1,000 mg/kg-d
daily for a total of 18 wk beginning 10 wk before
Dose
(mg/kg-d)
Incidence of
centrilobular
hvoertrophv
Dose
(mg/kg-d)
Incidence of
centrilobular
hvoertrophv
mating until after weaning of the pups
0
0/25
0
0/25
P0, female (25/group): 0, 250, 500,1,000 mg/kg-d
250
0/25
250
0/25
daily for a total of 18 wk beginning 10 wk before
mating until PND 21
500
0/25
500
0/25
1,000
3/25
1,000
0/25
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Toxicological Review ofETBE
Reference and study design
Results
JPEC (2008b)
Male
Female
rat, CRL:CD(SD)
inhalation - vapor
male (NR): 0, 150, 500, 1,500, 5,000 ppm (0, 627,
2,090, 6,270, 20,900 mg/m3)b; female (NR): 0,150,
Incidence of
Incidence of
Dose
centrilobular
Dose
centrilobular
(mg/m3)
hvoertrophv
(mg/m3)
hvpertrophv
500, 1,500, 5,000 ppm (0, 627, 2,090, 6,270,
0
0/10
0
0/10
20,900 mg/m3)
dynamic whole body chamber; 6 hr/d, 5 d/wk for
627
0/10
627
0/10
13 wk; generation method, analytical
2,090
0/10
2,090
0/10
concentration, and method reported
6,270
0/10
6,270
0/10
20,900
4/10*
20,900
6/10*
JPEC (2008b)
Male
Female
rat, CRL:CD(SD)
inhalation - vapor
male (6/group): 0, 5,000 ppm (0, 20,900 mg/m3)b;
female (6/group): 0, 5,000 ppm (0,
Incidence of
Incidence of
Dose
(mg/m3)
centrilobular
hvoertrophv
Dose
(mg/m3)
centrilobular
hvpertrophv
20,900 mg/m3)b
0
0/6
0
0/6
dynamic whole body chamber; 6 hr/d, 5 d/wk for
13 wk followed by a 28-d recovery period;
20,900
0/6
20,900
0/6
generation method, analytical concentration, and
method reported
Medinskv et al. (1999); Bond et al. (1996b)
Male
Female
rat, Fischer 344
inhalation - vapor
male (48/group): 0, 500,1,750, 5,000 ppm (0,
2,090, 7,320, 20,900 mg/m3)b; female (48/group):
Incidence of
Incidence of
Dose
(mg/m3)
centrilobular
hvoertrophv
Dose
(mg/m3)
centrilobular
hvpertrophv
0, 500, 1,750, 5,000 ppm (0, 2,090, 7,320,
0
0/11
0
0/10
20,900 mg/m3)b;
dynamic whole body chamber; 6 hr/d, 5 d/wk for
2,090
0/11
2,090
0/11
13 wk; generation method, analytical
7,320
0/11
7,320
0/11
concentration, and method reported
20,900
0/11
20,900
0/11
Medinskv et al. (1999); Bond et al. (1996a)
Male
Female
mice, CD-I
inhalation - vapor
Dose
(mg/m3)
Incidence of
centrilobular
hvpertrophv
Dose
(mg/m3)
Incidence of
centrilobular
hvpertrophv
male (40/group): 0, 500,1,750, 5,000 ppm (0,
2,090, 7,320, 20,900 mg/m3)b; female (40/group):
0, 500, 1,750, 5,000 ppm (0, 2,090, 7,320,
0
0/15
0
0/13
20,900 mg/m3)b
dynamic whole body chamber; 6 hr/d, 5 d/wk for
2,090
0/15
2,090
2/15
13 wk; generation method, analytical
7,320
2/15
7,320
1/15
concentration, and method reported
20,900
8/10*
20,900
9/14*
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Toxicological Review ofETBE
Reference and study design
Results
Mivata et al. (2013); JPEC (2008c)
Male
Female
rat, CRL:CD(SD)
oral -gavage
male (15/group): 0, 5, 25,100, 400 mg/kg-d;
female (15/group): 0, 5, 25,100, 400 mg/kg-d
Dose Incidence of Dose
(mg/kg-d) centrilobular (mg/kg-d)
hvoertrophv
Incidence of
centrilobular
hvpertrophv
daily for 26 wk
0
0/15
0
0/15
5
0/15
5
0/15
25
0/15
25
0/15
100
0/15
100
0/15
400
6/15*
400
6/15*
Saito et al. (2013); JPEC (2010b)
Male
rat, Fischer 344
inhalation - vapor
male (50/group): 0, 500,1,500, 5,000 ppm (0,
Dose
(mg/m3)
Acidophilic Basophilic
foci in liver foci in liver
Bile duct
hyperplasia
Centrilobular
hvpertrophv
2,090, 6,270, 20,900 mg/m3)b; female (50/group):
0
31/50 18/50
48/50
0/50
0, 500, 1,500, 5,000 ppm (0, 2,090, 6,270,
20,900 mg/m3)b
2,090
28/50 10/50
44/50
0/50
dynamic whole body inhalation; 6 hr/d, 5 d/wk for
6,270
36/49 13/49
46/49
0/49
104 wk; generation method, analytical
concentration, and method reported
20,900
Female
39/50* 33/50*
41/50
0/50
Dose
Acidophilic Basophilic
Bile duct
Centrilobular
(mg/m3)
foci in liver foci in liver
hyperplasia
hvpertrophv
0
2/50 36/50
5/50
0/50
2,090
1/50 31/50
8/50
0/50
6,270
4/50 32/50
7/50
0/50
20,900
2/50 28/50
6/50
0/50
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Toxicological Review ofETBE
Reference and study design
Results
Suzuki et al. (2012); JPEC (2010a)
Male
rat, Fischer 344
oral - water
male (50/group): 0, 625, 2,500,10,000 ppm (0, 28,
Dose
(mg/kg-
dl
Acidophilic
foci in liver
Basophilic Bile duct
foci in liver hyperplasia
Centrilobular
hvpertrophv
121, 542 mg/kg-d)a; female (50/group): 0, 625,
2,500,10,000 ppm (0, 46, 171, 560 mg/kg-d)a
0
14/50
14/50
49/50
0/50
daily for 104 wk
28
12/50
18/50
47/50
0/50
121
17/50
20/50
48/50
0/50
542
13/50
22/50
47/50
0/50
Female
Dose
Acidophilic
Basophilic Bile duct
Centrilobular
(mg/kg-
d)
foci in liver
foci in liver hyperplasia
hvpertrophv
0
2/50
36/50
1/50
0/50
46
2/50
25/50*
4/50
0/50
171
1/50
31/50
4/50
0/50
560
0/50
30/50*
3/50
0/50
Weng et al. (2012)
Male
Female
mice, C57BL/6
inhalation - vapor
male (5/group): 0, 500,1,750, 5,000 ppm (0,
2,090, 7,320, 20,900 mg/m3)b; female (5/group):
Dose
(mg/m3
Incidence of
) centrilobular
hvoertrophv
Dose
(mg/m3)
Incidence of
centrilobular
hvpertrophv
0, 500, 1,750, 5,000 ppm (0, 2,090, 7,320,
0
1/5
0
0/5
20,900 mg/m3)b
dynamic whole body chamber, 6 hr/d, 5 d/wk for
2,090
0/5
2,090
0/5
13 wk; generation methods not reported, but
7,320
0/5
7,320
1/5
analytical methods (gas chromatograph) and
concentration reported
20,900
5/5*
20,900
5/5*
Weng et al. (2012)
Male
Female
mice, Aldh2-/-
inhalation - vapor
male (5/group): 0, 500,1,750, 5,000 ppm (0,
2,090, 7,320, 20,900 mg/m3)b; female (5/group):
Dose
(mg/m3
Incidence of
) centrilobular
hvpertrophv
Dose
(mg/m3)
Incidence of
centrilobular
hvpertrophv
0, 500, 1,750, 5,000 ppm (0, 2,090, 7,320,
0
0/5
0
0/5
20,900 mg/m3)b
dynamic whole body chamber, 6 hr/d, 5 d/wk for
2,090
3/5
2,090
0/5
13 wk; generation methods were not reported,
7,320
2/5
7,320
0/5
but analytical methods (gas chromatograph) and
concentration reported
20,900
5/5*
20,900
4/5*
1 Conversion performed by study authors.
2 b4.18 mg/m3 = 1 ppm.
3 NR: not reported; *: result is statistically significant (p < 0.05) based on analysis of data by study authors.
4 for controls, no response relevant; for other doses, no quantitative response reported.
5
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Table 1-11. Evidence pertaining to liver biochemistry effects in animals
exposed to ETBE
Reference and study design
Results
JPEC (2008b)
rat, CRL:CD(SD)
inhalation - vapor
male (NR): 0,150, 500,1,500, 5,000 ppm
(0, 627, 2,090, 6,270, 20,900 mg/m3)b;
female (NR): 0,150, 500,1,500,
5,000 ppm (0, 627, 2,090, 6,270,
20,900 mg/m3)
dynamic whole body chamber; 6 hr/d,
5 d/wk for 13 wk; generation method,
analytical concentration, and method
reported
Response relative to control
Male
Dose
(mg/m3)
0
627
2,090
6,270
20,900
Female
Dose
(mg/m3)
0
627
2,090
6,270
20,900
ALT
9%
0%
5%
12%
ALT
-1%
11%
-5%
26%
ALP
13%
12%
-12%
-9%
ALP
-3%
-12%
-7%
5%
AST
3%
1%
-7%
4%
AST
2%
-95%
12%
0%
GGT
11%
0%
11%
-100%
GGT
25%
12%
25%
25%
Mivata et al. (2013); JPEC (2008c)
rat, CRL:CD(SD)
oral -gavage
male (15/group): 0, 5, 25,100,
400 mg/kg-d; female (15/group): 0, 5, 25,
100, 400 mg/kg-d
daily for 180 d
Response relative to control
Male
Dose
(mg/kg-d) MI ME ASI GGT
0
-
-
-
-
5
10%
2%
16%
25%
25
48%
12%
19%
50%
100
13%
-7%
20%
25%
400
35%
27%
23%
100%
Female
Dose
(mg/kg-d)
ALT
ALP
AST
GGT
0
-
-
-
-
5
11%
6%
10%
40%
25
21%
-21%
13%
20%
100
46%
-18%
19%
0%
400
21%
-19%
4%
-20%
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Reference and study design
Results
Saito et al. (2013); JPEC (2010b)
rat, Fischer 344
inhalation - vapor
male (50/group): 0, 500,1,500, 5,000 ppm
(0, 2,090, 6,270, 20,900 mg/m3)b; female
(50/group): 0, 500, 1,500, 5,000 ppm (0,
2,090, 6,270, 20,900 mg/m3)b
dynamic whole body inhalation; 6 hr/d,
5 d/wk for 104 wk; generation method,
analytical concentration, and method
reported
Response relative to control
Male
ALT
53%
-3%
24%
ALP
0%
-21%*
-5%
AST
29%
-16%
-2%*
22%
10%
18%*
GGT
33%
50%*
200%*
Dose
(mg/m3)
0
2,090
6,270
20,900
Female
Dose
(mg/m3) MI ALP ASI GGT
0 - - - -
2,090 2% 12%
6,270 -5% -4%
20,900 4%* 4%
50%
0%
150%
Suzuki et al. (2012); JPEC (2010a)
rat, Fischer 344
oral - water
male (50/group): 0, 625, 2,500,
10,000 ppm (0, 28, 121, 542 mg/kg-d)a;
female (50/group): 0, 625, 2,500,
10,000 ppm (0, 46, 171, 560 mg/kg-d)a;
daily for 104 wk
Response relative to control
Male
Dose
(mg/kg-d) MI ALP ASI GGT
0
-
-
-
-
28
-17%
-5%
-21%
0%
121
2%
3%
-3%
43%*
542
-4%
0%
-1%
29%
Female
Dose
(mg/kg-d)
ALT
ALP
AST
GGT
0
-
-
-
-
46
-10%
-16%
-19%
0%
171
-15%
2%
-17%
0%
560
-26%
-15%
-46%*
33%
1 Conversion performed by study authors.
2 b4.18 mg/m3 = 1 ppm.
3 NR: not reported; *: result is statistically significant (p < 0.05) based on analysis of data by study authors.
4 -: for controls, no response relevant; for other doses, no quantitative response reported.
5 (n): number evaluated from group.
6 Percent change compared to controls calculated as 100 x [(treated value - control value) -f control value],
7 Abbreviations: ALT = alanine aminotransferase, ALP = alkaline phosphatase, AST = aspartate aminotransferase,
8 GGT = gamma-glutamyl transferase.
9
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
¦ = exposures at which the endpoint was reported statistically significant by study authors
~ =exposures at which the endpoint was reported not statistically significant by study authors
Relative Liver
Weight
PO Male rat; 16wks (A)
PO Female rat; 16wks (A)
PO Male rat; 18wks (B)
PO Female rat; 18wks (B)
F1 Male rat; GD 0-adult (B)
F1 Female rat; GD 0-adult (B)
Male rat; 23wks (C)
Male rat; 26wks (D)
Female rat; 26wks (D)
0 B-
i-
~ ~ ¦
~ ~ a
o—b—a
¦
—i ¦
Centrllobular
Hypertrophy
PO Male rat; 18wks (B)
PO Female rat; 18wks (B)
Male rat; 26wks (D) ¦
Female rat; 26wks (D)
Male rat; 104wks (E)
Female rat; 104wks (E) ¦
-B-
0—~—~
~ DO
-H-
H=h
C3-
-B-
Serum
Liver
Enzymes
Male rat; ALT, AST, ALP, GGT;26wks (D)
Female rat; ALT, AST, ALP, GGT;26wks (D)
Male rat; ALT, AST, ALP;104wks (E) ¦
Male rat; t GGT;104w'--
Female rat; ALT, ALP, GGT;10 ! I (E)
Female rat; I AST;104wks (E)
-B-
-a-
-ffl-
13-
-H
13 HD-
10 100
Dose (mg/kg-day)
1,000
10,000
Sources; (A) Fujii et al„ 2010; JPEC, 2008e (B) Gaoua, 2004b (€} Hagiwara et al„ 2011 (D) Miyata etal., 2013;
JPEC, 2008c (E) Suzuki et al, 2012; JPEC, 2010a
Figure 1-9. Exposure-response array of noncancer liver effects following oral
exposure to ETBE.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
¦ = exposures at which the endpoint was reported statistically significant by study authors
~ =exposures at which the endpoint was reported not statistically significant by study authors
Relative Liver
Weight
Male rats; 13wks (A)
Female rats; 13wks (A)
Male rats; 13wks, 28d recovery (A)
Female rats; 13wks, 28d recovery (A)
Ccntrilobular
Hypertrophy
Male rats; 13wks (A)
Female rats; 13wks (A)
Male rats; 13wks, 28d recovery (.A)
Female rats; 13wks, 28d recovers I \)
Male rats; 13 wks (B)
Female rats; 13 wks (B)
Male mice; 13 wks (B)
Female mice; 13 wks (B)
Male mice; 13 wks (D)
Female mice; 13 wks (D)
Male Aldh2-/- mice; 13 wks [D]
Female Aldh2-/- mice; 13 wks (D)
Male rats; 104wks [C)
Female rats; 104wks (C)
O-
B-
O-
a-
o-
C3-
o-
~h
o-
Q-
-B-
-B-
-B-
-B-
-B-
-B-
~
~
-B
-B
Hi
Hi
-B
-B
Male rats; ALT, AST, ALP, GGT; 13wks (A)
Female rats; ALT, AST, ALP, GGT; 13wks [A)
Male rats; ALT; 104wks (C)
Male rats; I AST; "104wks (C)
Serum Liver Male rats; J.ALP; 104wks (C)
Enzymes
Male rats; TGGT; 104wks (C)
Female rats; ALT, ALP, GGT; 104wks (C)
Female rats; T'AST; 104wks (C)
O-
Q-
-B-
-B-
a-
ta-
B-
o-
C3-
o-
-B-
-B-
-B-
-B-
-B-
-B
-B
-B
¦
-B
100 1,000 10,000
Exposure Concentration (mg/m3)
100,000
Sources: (A) JPEC, 2008b (B) Medinsky et al, 1999; Bond et al„ 1996 (€} Saito et al, 2013; JPEC, 2010b (D) Weng
eta!,, 2012
Figure 1-10. Exposure-response array of noncancer liver effects following
inhalation exposure to ETBE.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Table 1-12. Evidence pertaining to liver tumor effects in animals exposed to
ETBE
Reference and study design
Results
Hepatocellular Adenoma and Carcinoma
Hagiwara et al. (2015)
rat, Wistar
oral -gavage
male (30/group): 0,100, 300, 500,1,000 mg/kg-d
daily for 19 wk following 2-wk tumor initiation by
N-ethyl-N-hydroxyethylnitrosamine (EHEN)
Incidence
Male
Dose
(mg/kg-d)
0
100
Adenoma
4/30
5/30
Carcinoma
0/30
2/30
Adenoma or
carcinoma
4/30
7/30
300
8/30
0/30
8/30
500
8/30
3/30
10/30
1,000
15/30*
5/30*
17/30*
Suzuki et al. (2012); JPEC (2010a)
rat, Fischer 344
oral - water
male (50/group): 0, 625, 2,500,10,000 ppm (0,
28,121, 542 mg/kg-d)a; female (50/group): 0, 625,
2,500,10,000 ppm (0, 46, 171, 560 mg/kg-d)a
daily for 104 wk
Incidence
Male
Dose
(mg/kg-d)
0
28
Adenoma
2/50
0/50
Carcinoma
2/50
0/50
Adenoma or
carcinoma
4/50
0/50
121
0/50
0/50
0/50
542
0/50
0/50
0/50
Female
Dose
(mg/kg-d)
Adenoma
Carcinoma
Adenoma or
carcinoma
0
0/50
0/50
0/50
46
0/50
0/50
0/50
171
0/50
0/50
0/50
560
1/50
0/50
1/50
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Toxicological Review ofETBE
Reference and study design
Results
Saito et al. (2013); JPEC (2010b)
Incidence
rat, Fischer 344
Male
inhalation - vapor
Dose
Adenoma or
male (50/group): 0, 500,1,500, 5,000 ppm (0,
(mg/m3)
Adenoma
Carcinoma
carcinoma
2,090, 6,270, 20,900 mg/m3)b; female (50/group):
0, 500, 1,500, 5,000 ppm (0, 2,090, 6,270,
0
0/50
0/50
0/50
20,900 mg/m3)b
2,090
2/50
0/50
2/50
dynamic whole body inhalation; 6 hr/d, 5 d/wk for
104 wk; generation method, analytical
6,270
1/50
0/50
1/50
concentration, and method reported
20,900
Female
Dose
9/50*
1/50
10/50*
Adenoma or
(mg/m3)
Adenoma
Carcinoma
carcinoma
0
1/50
0/50
1/50
2,090
0/50
0/50
0/50
6,270
1/50
0/50
1/50
20,900
1/50
0/50
1/50
Liver Neoplasm
Hagiwara et al. (2011); JPEC (2008d)
Incidence
rat, Fischer 344
Male
oral -gavage
male (30/group): 0, 300,1,000 mg/kg-d
daily for 23 wk following a 4-wk tumor initiation
Dose
(mg/kg-d)
Liver
neoplasm
by DMBDDC
0
1/30
+ no DMBDD initiation
300
1,000
0+
1,000+
1/30
6/30*
0/12
0/12
Maltoni et al. (1999)
Incidence
rat, Sprague-Dawley
Male
Female
oral - gavage
male (60/group): 0, 250,1,000 mg/kg-d; female
(60/group): 0, 250,1,000 mg/kg-d
4 d/wk for 104 wk; observed until natural death
Dose
(mg/kg-d)
0
Liver
neoplasm
0/60
Dose
(mg/kg-d)
0
Liver
neoplasm
0/60
250
0/60
250
0/60
NOTE: Tumor data not reanalyzed bv Malarkev
1,000
0/60
1,000
0/60
and Bucher (2011).
1 Conversion performed by study authors.
2 b4.18 mg/m3 = 1 ppm.
3 cDiethylnitrosamine (DEN), N-butyl-N-(4-hydroxybutyl)nitrosamine (BBN), N-methyl-N-nitrosourea (MNU), 1,2-
4 dimethylhydrazine dihydrochloride (DMH), and N-bis(2-hydroxypropyl)nitrosamine (DHPN).
5 *: result is statistically significant (p < 0.05) based on analysis of data by study authors.
6 for controls, no response relevant; for other doses, no quantitative response reported.
This document is a draft for review purposes only and does not constitute Agency policy.
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1 Mode of Action Analysis - Liver Effects
2 Key characteristics of carcinogens
3 Mechanistic information was grouped into 10 "key characteristics" useful for summarizing
4 and organizing the mechanistic data relevant to carcinogens fSmith etal.. 20161. The evidence
5 available for each characteristic is summarized in Table 1-13. Altogether, 5 of the 10 key
6 characteristics were found to have pertinent positive literature. ETBE was found to have the
7 potential for the formation of electrophilic metabolites, but it was concluded that there was
8 inadequate evidence that ETBE induced any of the remaining 9 key characteristics.
9 Table 1-13. Positive evidence of key characteristics of cancer for ETBE.
Characteristic
Evidence
1. Is electrophilic or can be metabolically activated
to electrophiles
Metabolized extensively to acetaldehyde in the
liver.1-3
2. Is genotoxic
Inadequate evidence to draw a conclusion from 12
studies examining micronucleus, DNA strand breaks,
chromosomal aberration, and gene mutation
assays2,3
3. Alters DNA repair or causes genomic instability
No positive studies identified
4. Induces epigenetic alterations
No positive studies identified
5. Induces oxidative stress
Inadequate evidence to draw a conclusion from 3
studies examining 8-OHdG, 8-hOGGl formation3,4
6. Induces chronic inflammation
No positive studies identified
7. Is immunosuppressive
No positive studies identified
8. Modulates receptor-mediated effects
Inadequate evidence to draw a conclusion from 2
studies examining PPAR, CAR, and PXR activation5
9. Causes immortalization
No positive studies identified
10. Alters cell proliferation, cell death, or nutrient
supply
Inadequate evidence to draw a conclusion from 3
studies examining basophilic, acidophilic foci and
cellular proliferation5
10 1See Supplemental Information section B.1.3.
11 2See Supplemental Information section B.2.2.
12 3See Acetaldehyde-mediated liver toxicity and genotoxicity in this section.
13 4See Oxidative stress in this section.
14 5See Receptor-mediated effects in this section.
15 Toxicokinetic considerations relevant to liver toxicity and tumors
16 ETBE is metabolized by cytochrome P450 (CYP) enzymes to an unstable hemiacetal that
17 decomposes spontaneously into tert-butanol and acetaldehyde fBernauer etal.. 19981.
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2
3
4
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7
8
9
10
11
12
13
14
15
16
17
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Toxicological Review ofETBE
Acetaldehyde is further metabolized in the liver by ALDH2, while tert-butanol undergoes systemic
circulation and ultimate excretion in urine. Thus, following ETBE exposure, the liver is exposed to
both acetaldehyde and tert-butanol, so the liver effects caused by tert-butanol (described in the
more detail in the draft IRIS assessment of tert-butanol) and acetaldehyde are relevant to
evaluating the liver effects observed after ETBE exposure.
tert-Butanol induces thyroid tumors in mice and kidney tumors in male rats, but has not
been observed to affect the incidence of rodent liver tumors following a 2-year oral exposure.
Although some data suggest tert-butanol could be genotoxic, the overall evidence is inadequate to
establish a conclusion. One study reported that tert-butanol might induce centrilobular
hypertrophy in mice after 2 weeks (Blancketal.. 20101: however, no related liver pathology was
observed in other repeat-exposure rodent studies including both subchronic and 2-year bioassays.
Although Blanck etal. (20101 reported some limited induction of mouse liver enzymes following
short-term tert-butanol exposure, no corresponding evidence exists in rats following any exposure
duration. Therefore, a role for tert-butanol in liver carcinogenesis ofETBE appears unlikely. No
MOA information is available for tert-butanol-induced noncancer liver effects.
In comparison, acetaldehyde associated with the consumption of alcoholic beverages is
genotoxic and mutagenic (IARC. 1999a). and acetaldehyde produced in the liver as a result of
ethanol metabolism has been suggested to be a contributor to ethanol-related liver toxicity and
cancer (Setshedi et al.. 20101. Additional discussion on the potential role of acetaldehyde in the
liver carcinogenesis ofETBE is provided below.
Receptor-mediated effects
ETBE exposure consistently increased relative liver weights in male and female rats and
increased hepatocellular adenomas and carcinomas in males (Saito etal.. 2013: TPEC. 2010b). In
addition to the transiently increased centrilobular hypertrophy, which is one possible indication of
liver enzyme induction, chronic exposure induced focal proliferative lesions that could be more
directly related to tumorigenesis. Notably, the centrilobular hypertrophy was only increased in rats
of both sexes via both oral and inhalation exposure at subchronic time points; it was not observed
via any exposure route at 2 years. Liver tumors were only observed in one sex (males) following
one route of exposure (inhalation), however, indicating that subchronic hypertrophy is not
associated with later tumor development This process was investigated in several studies to
determine whether nuclear receptor activation is involved.
Centrilobular hypertrophy is induced through several possible mechanisms, many of which
are via activation of nuclear hormone receptors such as peroxisome proliferator-activated receptor
a (PPARa), pregnane X receptor (PXR), and the constitutive androstane receptor (CAR). The
sequence of key events hypothesized for PPARa induction of liver tumors is as follows: activation of
PPARa, upregulation of peroxisomal genes, induction of gene expression driving PPARa-mediated
growth and apoptosis, disrupted cell proliferation and apoptosis, peroxisome proliferation,
preneoplastic foci, and tumors (Klaunig et al.. 20031. The sequence of key events hypothesized for
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2
3
4
5
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7
8
9
10
11
12
13
14
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18
19
20
21
22
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Toxicological Review ofETBE
CAR-mediated liver tumors is as follows: CAR activation, altered gene expression as a result of CAR
activation, increased cell proliferation, clonal expansion leading to altered foci, and liver adenomas
and carcinomas (Elcombe etal.. 20141. PXR, which has no established MOA, is hypothesized to
progress from PXR activation to liver tumors in a similar manner as CAR. This progression would
include PXR activation, cell proliferation, hypertrophy, CYP3A induction, and clonal expansion
resulting in foci development. One study that orally exposed male rats to low and high
concentrations ofETBE reported that several key sequences in the PPARa, PXR, and CAR pathways
were affected fKakehashi etal.. 20131.
PPAR
Limited evidence suggests that ETBE could activate PPAR-mediated events fKakehashi et
al.. 20131. For instance, mRNA expression was significantly elevated for PPARa and PPARy after 1
week of exposure but not after 2 weeks. In addition, several PPARa-mediated proteins involved in
lipid and xenobiotic metabolism were upregulated in the liver after 2 weeks of exposure such as
ACOX1, CYP4A2, and ECH1. Additional effects in the PPAR pathway such as DNA damage (8-OHdG)
and apoptosis (ssDNA) also were significantly increased after 2 weeks at the highest concentration
ofETBE. Cell proliferation was increased after 3 days fKakehashi et al.. 2 0151. unchanged after 1
week, significantly decreased after 2 weeks fKakehashi et al.. 20131 and increased after 28 days
fKakehashi etal.. 20151. The number of peroxisomes per hepatocyte was increased greater than
fivefold after 2 weeks of treatments. Finally, the incidences of preneoplastic basophilic and
acidophilic foci were significantly increased in males after 2 years of inhalation exposure to ETBE
fSaito etal.. 2013: TPEC. 2010bl.
Several measures required for a full evaluation of the PPAR MOA were absent. Selective
clonal expansion and gap junction intercellular communication were not examined in any study. No
evidence is available in wild-type or PPARa-null mice to demonstrate if PPARa gene expression
changes in KO mice. The high dose ofETBE (2,000 mg/kg-day) which induced the most consistent
changes in PPARa, Cyp4a, Cypla, and Cyp3a in the the oral gavage study fKakehashi et al.. 20131
yielded a higher internal metabolic rate in the liver (3.98 mg ETBE/hr) than from the 20,700 mg
ETBE/m3 inhalation dose (3.34 mg ETBE/hr) that increased liver tumors in the 2-year inhalation
study fSaito etal.. 2013: TPEC. 2010bl. Only Cyp2b genes associated with PPARa expression were
affected at the low gavage dose (300mg/kg-day), thus demonstrating poor dose-response
relationships between PPAR-mediated genes and downstream effects. Finally, PPAR agonists
typically decrease rates of apoptosis early in the process, which is in contrast to the increased rate
of apoptosis observed after 2 weeks ofETBE exposure fKakehashi etal.. 20131. Perturbation of
apoptosis is required for this MOA, indicating that this MOA might not be operative. Overall, these
data are inadequate to conclude that ETBE induces liver tumors via a PPARa MOA.
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3
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7
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9
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13
14
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16
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CAR/PXR
Kakehashi etal. (20131 reported several CAR- and PXR-mediated events following ETBE
exposure. After 2 weeks of exposure at the high dose ofETBE, CAR- and PXR-regulated xenobiotic
metabolic enzymes were upregulated, including Cyp2bl, Cyp2b2, Cyp3al, and Cyp3a2 as determined
by mRNA or protein expression. Other PXR/CAR-regulated genes such as Sultldl, Ugt2b5, and
Ugtlal also had elevated mRNA expression after 1 and 2 weeks of exposure, which all suggest
activation of CAR and PXR However, with the exception of Cyp2b, these genes were only increased
at the high dose, which yielded an internal rate of ETBE metabolism (3.98 mg/hr) that was greater
than the metabolism rate (3.34 mg/hr) associated with liver tumors, which demonstrates poor dose
response concordance. Histological evidence (preneoplastic foci) supporting increased liver cell
proliferation is available following chronic, but not subchronic, exposures (Saito etal.. 2013: TPEC.
2010b). Several data gaps were not evaluated, such as a lack of clonal expansion and gap junction
communication. These data provide evidence that CAR and PXR are activated at high
concentrations in the liver following acute ETBE exposure; however, due to crosstalk of CAR and
PXR on downstream effects such as cell proliferation, preneoplastic foci, and apoptosis, determining
the relative contribution of each pathway in tumorigenesis is not possible. Furthermore, the data do
not provide enough information to determine dose-response relationships or temporal
associations, which are critical for establishing an MOA. Finally, the available data from these
studies do not allow for parsing which effects are induced by PPAR or CAR/PXR activation.
Altogether, these data are inadequate to conclude that ETBE induces liver tumors via a CAR/PXR
MOA.
Acetaldehvde-mediated liver toxicity and ge no toxicity
Another possible MOA for increased tumors could be due to direct genotoxicity and
mutagenicity resulting from the production of acetaldehyde in the liver, the primary site for ETBE
metabolism. Acetaldehyde produced as a result of metabolism of alcohol consumption is considered
carcinogenic to humans, although evidence is not sufficient to show that acetaldehyde formed in
this manner causes liver carcinogenesis flARC. 20121. Acetaldehyde administered directly has been
demonstrated to increase the incidence of carcinomas following inhalation exposure in the nasal
mucosa and larynx of rats and hamsters. Furthermore, acetaldehyde has induced sister chromatid
exchanges in Chinese hamster ovary cells, gene mutations in mouse lymphomas, and DNA strand
breaks in human lymphocytes (IARC 1999a). Acetaldehyde has been shown to have an inhibitory
effect on PPARa transcriptional activity fVenkata etal.. 20081. although no effect of acetaldehyde on
CAR or PXR activation has been established. Additionally, the acetaldehyde metabolic enzyme
aldehyde dehydrogenase 2 (ALDH2) is polymorphic in the human population, which contributes to
enhanced sensitivity to the effects of acetaldehyde among some subpopulations such as people of
East Asian origin flARC. 2012: Brennan etal.. 2004). IARC ("20121 found that ALDH2 status was
associated with increased esophageal cancer. Although IARC (20121 found inconclusive evidence
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for a contribution of ALDH2 to liver cancer, Eriksson (20151 concluded that reduced aldehyde
metabolism is associated with liver cancer by further analyzing the ALDH2 compositions of the
controls in the case-control studies.
Several studies have examined the role of acetaldehyde and the metabolizing enzyme
ALDH2 in genotoxicity and centrilobular hypertrophy following ETBE exposure. Ninety-day
inhalation exposure to ETBE significantly increased the incidence of centrilobular hypertrophy in
male AIdh2 knockout (KO) mice compared with wild type (WT), while females appeared to be
similarly sensitive to controls fWengetal.. 20121. Hepatocyte DNA damage as determined by DNA
strand breaks and oxidative base modification was increased at the highest concentration of ETBE
exposure in the WT males, but not in WT females. Measures of DNA damage were all statistically
significantly exacerbated in both male and female AIdh2 KO mice fWeng etal.. 20121. Further
demonstrating enhanced genotoxic sensitivity in males compared with females, erythrocyte
micronucleus assays and oxidative DNA damage (8-hOGGl) in leukocytes were observed to be
statistically significantly increased and dose responsive only in male AIdh2 KO mice fWeng etal..
20131. Together, although these data suggest a potential role for acetaldehyde in the increased liver
tumor response observed in male rats exposed to ETBE, the available data are inadequate to
conclude that ETBE induces liver tumors via acetaldehyde-mediated mutagenicity.
Oxidative Stress
Studies with pertinent information to the evaluation of oxidative stress are limited to two
studies measuring oxidative DNA damage in leukoyctes and hepatocytes in mice fWeng etal.. 20121
and one study in the liver of rats fKakehashi etal.. 20131. Hepatocytes in male mice had increased
levels of 8-OHdG after 13 weeks of inhalation exposure to the concentration ofETBE that induced
liver tumors following 2 years of inhalation exposure. No significant dose response was reported.
Similarly, 8-OHdG was increased after 2 weeks of oral gavage in rats fKakehashi etal.. 20131 at a
concentration two-fold greater than that inducing rat liver tumors in two-stage initiation-
promotion assays fHagiwara etal.. 2015: Hagiwara etal.. 20111. In addition, as discussed in the
previous paragraph, oxidative DNA damage was also induced in AIdh2 KO mice fWeng etal.. 20131.
Altogether, the data are not sufficient to establish temporal or dose response concordance.
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Overall Conclusions on MO A for Liver Effects
Several reviews of the available mechanistic data suggest that the PPAR, PXR, and CAR
pathways induce liver tumors in a manner not relevant to humans fElcombe etal.. 2014: Klaunig et
al.. 20031. although this conclusion has been questioned fGuvton et al.. 20091. The database is
inadequate to determine if nuclear receptor-mediated pathways (i.e., PPAR and CAR/PXR)
contribute to the tumorigenesis observed in ETBE-treated male rats. Furthermore, centrilobular
hypertrophy was observed at the same concentrations that induced liver weight changes in rats of
both sexes after 13-week inhalation and 26-week oral exposure, yet liver tumors were observed
only following oral exposure in male rats. This observation suggests that these transient effects are
not associated with the observed rat liver tumorigenesis. Therefore, given the available data, ETBE-
induced liver tumors in male rats are considered relevant to humans.
Evidence suggests that metabolism ofETBE to acetaldehyde could contribute to ETBE-
induced liver carcinogenesis. For instance, enhancement of ETBE-induced liver toxicity and
genotoxicity has been reported in AIdh2-deficient mice, which have an impaired ability to
metabolize acetaldehyde fWengetal.. 2013: Wengetal.. 20121. Additionally, because lack of ALDH2
activity is directly relevant to the substantial human subpopulation that is deficient in the ALDH2
isozyme flARC. 20121. these data suggest a role for acetaldehyde in ETBE-induced liver
tumorigenesis. The database, however, is inadequate to conclude that ETBE induces liver tumors
via acetaldehyde-mediated mutagenic MOA.
Integration of Liver Effects
Liver effects were observed in oral and inhalation studies with exposure durations of
13 weeks to 2 years. Evidence for ETBE-induced noncancer liver effects is available from rat and
mouse studies that include centrilobular hypertrophy, increased liver weights, and changes in
serum liver enzyme levels. Based on dose-related increases in relative liver weights and transient
increases in hepatocellular hypertrophy in male and female rats, and considering the poor temporal
correlation of serum biomarkers and pathological lesions indicative of accumulating damage,
evidence of liver effects associated with ETBE exposure is suggestive. The hazard and dose-
response conclusions regarding these noncancer endpoints associated with ETBE exposure are
further discussed in Section 1.3.1.
The carcinogenic effects observed include increased hepatocellular adenomas and
carcinomas in males in a 2-year bioassay and ETBE-promoted liver tumorigenesis after 23 weeks
following mutagen pretreatment Although only one carcinoma was observed, rodent liver
adenomas could progress to malignancy, eventually forming carcinomas fLiau etal.. 2013:
McConnell etal.. 19861. Mechanistic data on the role of PPAR, PXR, and CAR activation in liver
tumorigenesis were inadequate to conclude that these pathways mediate tumor formation.
Additional mechanistic studies in transgenic mice suggest that lack of Aldh2 enhances ETBE-
induced liver toxicity and genotoxicity, which is consistent with the observed genotoxicity being
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mediated by the ETBE metabolite acetaldehyde, although the database is inadequate to conclude
that ETBE induces liver tumors via acetaldehyde-mediated mutagenic MOA. The hazard and dose-
response conclusions regarding the liver tumors associated with ETBE exposure are further
discussed as part of the overall weight of evidence for carcinogenicity in Section 1.3.2.
1.2.3. Reproductive Effects
Synthesis of Effects Related to Male Reproduction
The database examining male reproductive effects following ETBE exposure contains no
human data but is comprised of animal data from rats and mice. Effects on male reproduction,
including fertility, male reproductive organ weights, histopathology, sperm parameters, and
hormone levels were evaluated in a one-generation oral study fFuiii etal.. 20101. a two-generation
oral study fGaoua. 2004b). 13- and 9-week inhalation studies fWeng etal.. 20141. and a 14-day oral
study fde Pevster etal.. 20091. Additional data on male reproductive organ weights and
histopathology were obtained from two 2-year carcinogenicity studies [oral: Suzuki etal. T20121:
TPEC f2010al: inhalation: Saito etal. T20131: TPEC f2010bl]. a medium term carcinogenicity study
(23-week oral exposure) f Hagiwara et al.. 2011: TPEC. 2008d). a 180-day oral study fMivata etal..
2013: TPEC. 2008c). a 90-day inhalation study flPEC. 2008bl. and a 13-week inhalation study
fMedinskv etal.. 19991. These studies were conducted in Sprague-Dawley rats, Fischer 344 rats,
CD-I mice, and C57BL/6 mice, and the design, conduct, and reporting of each study were of
sufficient quality to inform human health hazard assessment. Selected endpoints from these studies
are summarized in Table 1-14.
The one- and two-generation reproductive toxicity studies found no effects on copulation,
fertility, or sperm parameters in adult male Sprague-Dawley rats exposed to ETBE by oral gavage at
concentrations up to 1,000 mg/kg-day for 10 weeks prior to mating fFuiii etal.. 2010: Gaoua.
2004b), nor in F1 male offspring exposed during gestation, lactation, and post-weaning diet fGaoua.
2004b). No dose-related changes in testicular histopathology were observed in F0 or F1 males
fGaoua. 2004bl. Furthermore, no dose-related histopathological changes or significant changes in
absolute male reproductive organ weight were observed in the 2-year carcinogenicity studies in
Fischer 344 rats at oral doses up to 542 mg/kg-day f Suzuki etal.. 2012: TPEC. 2010al or at
inhalation exposure concentrations up to 20,900 mg/m3 f Saito etal.. 2013: TPEC. 2010b): in the
medium term carcinogenicity study in Fischer 344 rats fHagiwara etal.. 2011: TPEC. 2008d): in the
180-day oral study in Sprague-Dawley rats at doses up to 400 mg/kg-day fMivata etal.. 2013: TPEC.
2008c); in the 90-day inhalation study in Sprague-Dawley rats at doses up to 20,900 mg/m3 flPEC.
2008b); or in the 14-day oral study in Fischer 344 rats at doses up to 1,800 mg/kg-day fde Pevster
etal.. 20091. In some cases, dose-related increases in relative organ weights were observed,
including significant increases in relative testis weight fFuiii etal.. 2010: TPEC. 2010b: Gaoua.
2004b) and relative prostate weight fGaoua. 2004b) at the highest doses tested, which may have
been attributable to reduced body weight gain in these groups.
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In contrast, testicular degeneration was observed in two 13-week ETBE inhalation studies
in which rats and mice were exposed to concentrations ranging from 2,090-20,900 mg/m3. In
Fischer 344 rats, a statistically significant increase in the percentage of seminiferous tubules with
spermatocyte degeneration was observed; however, there were no significant microscopic findings
in CD-I mice under these same exposure conditions and no changes in male reproductive organ
weights in rats or mice (Medinskv etal.. 19991. In C57BL/6 wild type and AIdh2 KO mice, there was
a dose-related increase in the incidence of atrophy of seminiferous tubules (described by the
authors as "slight" or "extremely slight" atrophy), with a greater incidence of atrophy occurring in
Aldh2 KO mice compared to wild type fWengetal.. 20141. ETBE-exposed mice also had significant
decreases in sperm head numbers and sperm mobility (expressed as the percentage of motile
sperm, percentage of static sperm, and percentage of sperm with rapid movement) and a significant
increase in sperm DNA damage (expressed as strand breaks and oxidative DNA damage), with
effects on sperm parameters reaching statistical significance at lower exposure concentrations in
AIdh2 KO mice (2,090 mg/m3) compared to wild type (7,320-20,090 mg/m3). Significantly
decreased epididymis weight was observed in AIdh2 KO mice but not wild type mice.
Wengetal. f20141 also conducted a 9-week inhalation study using lower ETBE exposure
concentrations (209-2,090 mg/m3) and three mouse genotypes (wild type, AIdh2 KO, and AIdh2
heterogeneous). Wild type mice had little to no change in male reproductive organ weights or
sperm parameters at any of the tested concentrations, whereas significant effects were observed on
sperm count, sperm mobility, and sperm DNA damage in AIdh2 KO and heterogeneous mice at
exposure concentrations as low as 836 mg/m3 ETBE. Aldh2 heterogeneous mice had significantly
decreased relative testis and epididymis weight in the 20,090 mg/m3 exposure group. Taken
together, the results ofWengetal. f20141 indicate that populations with inactive AIdh2 variants are
more susceptible to male reproductive toxicity following exposure to ETBE.
Although testicular lesions were not found in the 14-day oral study in Fischer 344 rats (de
Pevster etal.. 20091. plasma estradiol levels in these animals were increased by up to 106%
compared to controls. Plasma testosterone in the 1,800 mg/kg-day dose group was decreased by
34% compared to controls, but the difference was not statistically significant and was not observed
in any other ETBE dose group. The authors conducted a separate in vitro experiment to evaluate
testosterone production in isolated Sprague-Dawley rat Leydig cells and found reduced
testosterone production in ETBE-treated cells compared to controls (data not shown in evidence
table).
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Table 1-14. Evidence pertaining to male reproductive effects in animals
exposed to ETBE
Reference and Study Design
Resu Its
Male Fertility
Fuiiietal. (2010); JPEC (2008e)
F0 Generation-Parent
rat, Sprague-Dawley
oral -gavage
F0, male and female
(24/sex/group): 0,100, 300,
Dose (mg/kg-
Copulation
Absolute
change from
Fertility index
Absolute
change from
dl
index (%)
control (%)
i%l
control (%)
1,000 mg/kg-d; dosed daily for
0
100
-
87.5
-
17 wks, from 10 wks premating
100
91.7
-8.3
100
12.5
to lactation day 21
300
95.8
-4.2
95.7
8.2
1,000
100
0
91.7
4.2
Gaoua(2004b)
F0 Generation-Parent
rat, Sprague-Dawley
oral - gavage
F0, male and female
(25/sex/group): 0, 250, 500,
Dose (mg/kg-
Male mating
Absolute
change from
Male fertility
Absolute
change from
dl
index3 (%)
control (%)
index" (%)
control (%)
1,000 mg/kg-d
0
100
-
92
-
dosed daily for 18 wks from 10
250
100
0
84
-8
wks premating until weaning of
F1 pups
500
100
0
88
-4
Fl, male and female (24—
25/group): 0, 250, 500, 1,000
1,000
100
0
100
8
mg/kg-d
dosed daily from PND 22 until
Fl Generation-Offspring
weaning of F2 pups
F0 Generation-Parent
Dose (mg/kg-
Male mating
Absolute
change from
Male fertility
Absolute
change from
dl
index3 (%)
control (%)
index" (%)
control (%)
0
96
-
92
-
250
96
0
92
0
500
100
4
88
-4
1,000
96
0
96
4
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Reference and Study Design
Resu Its
Testicular Histopathology
Medinskv et al. (1999); Bond et
Incidence of spermatocyte
Incidence of sloughed
al. (1996b)
Dose (mg/m3)
degeneration
epithelium
rat, Fischer 344
inhalation - vapor
0
11/11
7/11
male (48/group): 0, 500,1,750,
2,090
11/11
3/11
5,000 ppm
(0, 2,090, 7,320,
7,320
11/11
3/11
20,900 mg/m3)a; female
20,900
10/11
7/11
(48/group): 0, 500, 1,750,
5,000 ppm
(0, 2,090, 7,320,
20,900 mg/m3)a
Dose (mg/m3)
Mean seminiferous tubules with
spermatocyte degeneration (%)
Absolute change from control
i%l
dynamic whole body chamber;
0
2.1
-
6 hr/d, 5 d/wk for 13 wk
2,090
2.4
0
7,320
7.8*
6
20,900
12.7*
Mean seminiferous tubules with
11
Absolute change from control
Dose (mg/m3)
lumenal debris (%)
i%l
0
2.1
-
2,090
0.7
-1
7,320
2.8
1
20,900
1
-1
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Reference and Study Design
Resu Its
Weng et al. (2014)
Wild Type Mice; 13-week Exposure
mice, C57BL/6
inhalation - vapor
male (5/group): 0, 500,1,750,
5,000 ppm (0, 2,090, 7,320,
Dose
(mg/kg-d)
Incidence of
"extremelv slight"
atrophv
Total incidence of
Incidence of atrophv of
"slight" atrophv seminiferous tubules
20,900 mg/m3)a
0
1/5
0/5
1/5
dynamic whole body
inhalation; 6 h/d, 5 d/wk for 13
2,090
0/5
0/5
0/5
wk; methods described in
7,320
2/5
0/5
2/5
Weng et al. (2012)
20,900
3/5
0/5
3/5
Knockout Mice (Aldh2-/-); 13-week Exposure
Incidence of
Total incidence of
Dose
"extremelv slight"
Incidence of
atrophv of
(mg/kg-d)
atrophv
"slight" atrophv seminiferous tubules
0
2/5
0/5
2/5
2,090
2/5
3/5
5/5
7,320
4/5
1/5
5/5
20,900
3/5
2/5
5/5
Sperm Parameters
Gaoua(2004b)
F0 Males
rat, Sprague-Dawley
oral -gavage
F0, male and female
(25/sex/group): 0, 250, 500,
Dose
Mean epididvmal
spermatozoa
% change
Mean epididvmal
sperm motilitv
Absolute
change from
(mg/kg-d)
count (n) ± SD
from control
(%) ± SD
control (%)
1,000 mg/kg-d
0
923 ± 200
-
99.7 ± 1.5
-
dosed daily for 18 wks from
10 wks premating until
250
938 ± 205
2
100 ±0
0
weaning of F1 pups
500
935 ±159
1
98.6 ±4
-1
Fl, male and female (24—
25/group): 0, 250, 500,
1,000
918 ±194
-1
97.6 ±6.6
-2
1,000 mg/kg-d
Mean epididvmal
dosed daily from PND 22 until
sperm with
Mean testicular
weaning of F2 pups
normal
Absolute
sperm heads
Dose
morphology
change from
(106/gram of
% change from
(mg/kg-d)
(%) ± SD
control (%)
testis) ± SD
control
0
93 ± 19
-
114.8 ± 18.7
-
250
93 ±19
0
109 ± 13.1
-5
500
97 ±2
4
108.1 ± 18.6
-6
1,000
96 ±2
3
109.8 ± 16.5
-4
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Reference and Study Design
Resu Its
Gaoua (2004b) (continued)
Mean dailv
testicular sperm
production
Dose
(106/gram of
% change
N (epididvmal
N (other sperm
(mg/kg-d)
testis)
from control
sperm count)
parameters)
0
18.8 ±3.1
-
25
25
250
17.9 ±2.2
-5
25
25
500
17.7 ±3.1
-6
25
25
1,000
18 ±2.7
-4
24
25
F1 Males
Mean epididvmal
spermatozoa
Mean epididvmal
Absolute
Dose
count (n)
% change
sperm motilitv
change from
(mg/kg-d)
±SD
from control
(%) ± SD
control (%)
0
725 ±150
-
84.6 ±34.1
-
250
673 ±197
-7
87.1 ±31.6
3
500
701 ± 97
-3
93.3 ±22
9
1,000
688 ±177
-5
88.3 ±29.4
4
Mean epididvmal
sperm with
Mean testicular
normal
Absolute
sperm heads
Dose
morphology
change from
(106/gram of
% change from
(mg/kg-d)
(%) ±SD
control (%)
testis) ± SD
control
0
84 ±30
-
100.6 ± 36.7
-
250
86 ±28
2
97.8 ±32.3
-3
500
86 ±27
2
105.3 ± 27.2
5
1,000
88 ±24
4
99.8 ±38.9
-1
Mean dailv
testicular sperm
production
Dose
(106/gram of
% change
N (epididvmal
N (other sperm
(mg/kg-d)
testis)
from control
sperm count)
parameters)
0
16.5 ± 6
-
22
24
250
16 ±5.3
-3
24
25
500
17.3 ±4.5
5
23
24
1,000
16.4 ±6.4
-1
24
25
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Reference and Study Design
Resu Its
Weng et al. (2014)
Wild Type Mice; 13-week Exposure
mice, C57BL/6
inhalation - vapor
male (5/group): 0, 500,1,750,
5,000 ppm (0, 2,090, 7,320,
20,900 mg/m3)a
Dose
(mg/m3)
Mean sperm
head numbers
(testis)
(x 106/g) ±SD
% change
from control
Motile sperm
(epididvmal) ± SE
Absolute
change from
control (%)
dynamic whole body
0
166.62 ±21.9
-
67.34 ±3.45
-
inhalation; 6 hr/d, 5 d/wk for
13 wk; methods described in
2,090
167.74 ± 28.02
1
69.64 ± 3.45
2
Weng et al. (2012)
7,320
167.78 ±25.52
1
62.73 ± 1.73
-5
20,900
150.94 ± 23.07
-9
Absolute
58.13 ±2.30
% Sperm with
-9
Absolute
Dose
% Static sperm
change from
rapid movement
change from
(mg/m3)
(epididvmal)
control (%)
(epididvmal)
control (%)
0
32.57 ± 3.00
-
55.00 ± 3.75
-
2,090
30.86 ± 3.86
-2
56.25 ±3.13
1
7,320
37.29 ± 1.71
5
49.38 ±3.13
-6
20,900
42.43 ± 2.57
Epididvmal
sperm DNA
breaks (tail
10
46.25* ±2.50
Epididvmal
sperm DNA
damage
(measurement of
-9
Dose
intensity in
% change
8-OH-dG in
% change from
(mg/m3)
comet assav)
from control
comet assav)
control
0
4.91 ±0.34
-
3.46 ±0.45
-
2,090
5.91 ±0.35
20
4.23 ±0.22
23
7,320
7.60* ±0.69
55
5.16* ±0.46
49
20,900
7.91* ±0.52
61
6.55 ± 1.13
89
Knockout Mice (Aldh2-/-); 13-week Exposure
Mean sperm
head numbers
Absolute
Dose
(testis)
% change
Motile sperm
change from
(mg/m3)
(x 106/g) ±SD
from control
(epididvmal) ± SE
control (%)
0
169.15 ±28.33
-
75.07 ± 2.88
-
2,090
127.08 ± 17.32
-25
61.23 ±5.03
-14
7,320
124.6* ± 11.96
-26
61.05* ±5.75
16
20,900
124.72* ± 18.72
-26
57.27* ±5.77
20
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Reference and Study Design
Resu Its
Weng et al. (2014) (continued)
Absolute
% Sperm with
Absolute
Dose
% Static sperm
change from
rapid movement
change from
(mg/m3)
(epididvmal)
control (%)
(epididvmal)
control (%)
0
25.46 ±2.56
-
66.74 ±2.17
-
2,090
40.34 ±5.14
15
51.54 ±2.84
-15
7,320
41.51* ±5.57
16
47.74* ±5.66
-19
20,900
45.27* ±5.58
Epididvmal
sperm DNA
breaks (tail
20
45.03* ±3.97
Epididvmal
sperm DNA
damage
(measurement of
-22
Dose
intensity in
% change
8-OH-dG in
% change from
(mg/m3)
comet assav)
from control
comet assav)
control
0
4.90 ±0.52
-
3.64 ±0.61
-
2,090
7.71 ±0.69
58
5.45 ±0.15
50
7,320
10.44* ± 0.78
113
7.65* ±0.61
110
20,900
9.46* ±0.69
93
7.95* ± 1.52
119
Weng et al. (2014)
Wild Type Mice; 9-Week Exposure
mice, C57BL/6
inhalation - vapor
male (NR): 0, 50, 200, 500 ppm
(0, 209, 836, 2,090 mg/m3)a
dynamic whole body
Dose
Mean sperm
head numbers
(testis)
% change
Motile sperm
Absolute
change from
(mg/m3)
(x 106/g) ±SD
from control
(epididvmal) ± SE
control (%)
inhalation; 6 hr/d, 5 d/wk for 9
0
199.62 ±27.22
-
85.82 ±4.26
-
wk; methods described in
Weng et al. (2012)
209
173.35 ±23.35
-13
78.72 ± 1.42
-7
836
170.47 ± 25.37
-15
82.27 ±2.13
-4
2,090
173.13 ± 16.28
-13
Absolute
80.14 ± 1.42
% Sperm with
-6
Absolute
Dose
% Static sperm
change from
rapid movement
change from
(mg/m3)
(epididvmal)
control (%)
(epididvmal)
control (%)
0
13.02 ± 3.38
-
71.11 ±2.78
-
209
21.74 ±2.96
9
65.56 ±2.22
-6
836
17.78 ±2.11
5
67.22 ±2.22
-4
2,090
16.36 ± 1.68
3
67.22 ±2.78
-4
This document is a draft for review purposes only and does not constitute Agency policy.
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Weng et al. (2014) (continued)
Epididvmal
Epididvmal
sperm DNA
sperm DNA
damage
breaks (tail
(measurement of
Dose
intensity in
% change
8-OH-dG in
% change from
(mg/m3)
comet assav)
from control
comet assav)
control
0
4.10 ±0.26
-
3.88 ±0.30
-
209
4.04 ±0.10
-2
3.73 ±0.15
-4
836
4.40 ±0.26
7
4.25 ±0.30
10
2,090
4.59 ±0.26
12
4.48 ±0.37
15
Knockout Mice (Aldh29-week Exposure
Mean sperm
head numbers
Absolute
Dose
(testis)
% change
Motile sperm
change from
(mg/m3)
(x 106/g) ±SD
from control
(epididvmal) ± SE
control (%)
0
216.19 ± 12.46
-
84.17 ±2.88
-
209
198.21 ± 20.54
-8
83.45 ± 2.88
-1
836
180.71* ±23.5
-16
77.70 ±2.88
-6
2,090
165.8* ±43.52
-23
69.06 ± 6.47
-15
Absolute
% Sperm with
Absolute
Dose
% Static sperm
change from
rapid movement
change from
(mg/m3)
(epididvmal)
control (%)
(epididvmal)
control (%)
0
14.57 ± 1.71
-
69.79 ±2.84
-
209
16.29 ±4.29
2
68.65 ±3.97
-1
836
21.43 ± 3.00
7
63.55 ±2.27
-6
2,090
30.00* ± 6.00
15
52.20* ±5.11
-18
Epididvmal
Epididvmal
sperm DNA
sperm DNA
damage
breaks (tail
(measurement of
Dose
intensity in
% change
8-OH-dG in
% change from
(mg/m3)
comet assav)
from control
comet assav)
control
0
4.65 ±0.17
-
3.66 ±0.30
-
209
4.67 ±0.09
0
3.96 ±0.30
8
836
5.71* ±0.34
23
4.48 ±0.30
22
2,090
7.01* ±0.26
51
4.85* ±0.22
33
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Reference and Study Design
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Weng et al. (2014) (continued)
Haplotype Mice (Aldh2 heterogeneous); 9-week Exposure
Mean sperm
head numbers
Absolute
Dose
(testis)
% change
Motile sperm
change from
(mg/m3)
(x 106/g) ±SD
from control
(epididvmal) ± SE
control (%)
0
202.76 ± 14.59
-
85.61 ±2.16
-
209
202.26 ±26.31
0
85.61 ±2.16
0
836
109.53* ±21.56
-46
73.38* ±3.60
-12
2,090
96.31* ±33.4
-53
76.98* ±3.60
-9
Absolute
% Sperm with
Absolute
Dose
% Static sperm
change from
rapid movement
change from
(mg/m3)
(epididvmal)
control (%)
(epididvmal)
control (%)
0
15.00 ± 1.71
-
70.14 ±2.24
-
209
15.00 ±2.14
0
68.59 ±2.24
-2
836
27.43* ±3.86
12
49.42* ± 6.24
-21
2,090
24.00* ± 3.00
9
58.08* ± 1.69
-12
Epididvmal
Epididvmal
sperm DNA
sperm DNA
damage
breaks (tail
(measurement of
Dose
intensity in
% change
8-OH-dG in
% change from
(mg/m3)
comet assav)
from control
comet assav)
control
0
3.51 ±0.25
-
4.04 ±0.22
-
209
3.70 ±0.34
5
4.45 ±0.14
10
836
5.32* ±0.43
52
4.86 ±0.43
20
2,090
5.86* ±0.42
67
5.34* ±0.50
32
Organ Weights
Fuiiietal. (2010); JPEC (2008e)
F0 Parents-Absolute Organ Weights
rat, Sprague-Dawley
oral -gavage
Dose
Mean testis
% change
Mean epididymis
% change from
FO, male and female
(mg/kg-d)
weight (g) ± SD
from control
weight (mg) ± SD
control
(24/sex/group): 0,100, 300,
1,000 mg/kg-d; dosed daily for
0
3.47 ±0.31
-
1371±136
-
17 wks, from 10 wks premating
100
3.48 ±0.28
0
1360 ± 83
-1
to lactation day 21
300
3.57 ±0.24
3
1381 ±73
1
1,000
3.57 ±0.31
3
1349 ± 95
-2
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
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Resu Its
Fuiiietal. (2010); JPEC (2008e)
Mean prostate
Mean seminal
(continued)
Dose
weight (mg)
% change
vesicle weight (g)
% change from
(mg/kg-d)
±SD
from control
± SD
control
0
787 ±180
-
2.16 ±0.23
-
100
778 ±158
-1
2.1 ±0.32
-3
300
752 ±172
-4
2.19 ±0.24
1
1,000
816 ±136
4
2.19 ±0.23
1
F0 Parents-Relative Organ Weights
Mean
Mean testis:
Absolute
epididymis: bodv
Absolute
Dose
bodv weight
change from
weight ratio (%)
change from
(mg/kg-d)
ratio (%) ± SD
control (%)
± SD
control (%)
0
0.554 ±0.065
-
219 ± 30
-
100
0.572 ±0.062
0.02
223 ± 18
4
300
0.589 ±0.076
0.03
228 ± 25
9
1,000
0.61* ±0.074
0.06
230 ± 24
Mean seminal
11
Mean prostate:
Absolute
vesicle: bodv
Absolute
Dose
bodv weight
change from
weight ratio (%)
change from
(mg/kg-d)
ratio (%) ± SD
control (%)
± SD
control (%)
0
125 ± 28
-
0.345 ± 0.054
-
100
128 ± 30
3
0.343 ±0.051
0.00
300
124 ± 30
-1
0.361 ±0.052
0.02
1,000
139 ± 23
14
0.373 ± 0.042
0.03
Gaoua(2004b)
F0 Parents-Absolute Organ Weights
rat, Sprague-Dawley
oral -gavage
FO, male and female
(25/sex/group): 0, 250, 500,
Dose
(mg/kg-d)
Mean testis
weight (left) (g)
± SD
% change
from control
Mean testis
weight (right) (g)
± SD
% change from
control
1,000 mg/kg-d
0
1.78 ±0.116
-
1.76 ±0.105
-
dosed daily for 18 wks from
10 wks premating until
250
1.73 ±0.181
-3
1.76 ±0.179
0
weaning of F1 pups
500
1.78 ±0.142
0
1.76 ±0.13
0
Fl, male and female (24—
25/group): 0, 250, 500,
1,000
1.75 ±0.237
-2
1.79 ±0.126
2
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
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Gaoua (2004b) (continued)
Mean epididymis
Mean epididymis
Dose
weight (left) (g) ±
% change
weight (right) (g)
% change from
(mg/kg-d)
SD
from control
± SD
control
0
0.77008 ± 0.054
-
0.78148 ± 0.053
-
250
0.77092 ± 0.077
0
0.78698 ±0.092
1
500
0.77784 ± 0.067
1
0.77492 ± 0.062
-1
1,000
0.80988 ±0.189
5
0.77528 ±0.056
-1
Mean seminal
% change
Dose
Mean prostate
% change
vesicle weight ±
from
(mg/kg-d)
weight ± SD
from control
SD
control N
0
1.41 ±0.272
-
2.06 ± 0.309
25
250
1.63 ±0.32
16
2.26 ±0.595
10 25
500
1.37 ±0.285
-3
2.19 ±0.439
6 25
1,000
1.62 ±0.396
15
2.28 ±0.574
11 25
FO Parents-Relative Organ Weights
Mean testis
Mean testis
weight: bodv
Absolute
weight: bodv
Absolute
Dose
weight ratio
change from
weight ratio
change from
(mg/kg-d)
(left) (g) ± SD
control (%)
(right) (g) ±SD
control (%)
0
0.297488 ± 0.029
-
0.29488 ± 0.029
-
250
0.29005 ± 0.025
-0.01
0.29427 ± 0.025
0.00
500
0.307 ± 0.033
0.01
0.30321 ±0.033
0.01
1,000
0.31052 ± 0.049
0.01
0.31497* ±0.029
0.02
Mean epididymis
Mean
weight (left):
Absolute
epididymis: bodv
Absolute
Dose
bodv weight
change from
weight ratio
change from
(mg/kg-d)
ratio (g) ± SD
control (%)
(right) (%) ± SD
control (%)
0
0.12886 ±0.014
-
0.13072 ±0.013
-
250
0.12947 ±0.013
0.00
0.13245 ± 1.014
0.00
500
0.13434 ±0.016
0.01
0.13383 ±0.015
0.00
1,000
0.14209 ±0.027
0.01
0.1367 ±0.012
0.01
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
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Gaoua (2004b) (continued)
Dose
(mg/kg-d)
Mean prostate
weight: bodv
weight ratio ± SD
Absolute
change from
control (%)
Mean seminal
vesicle: bodv
weight ratio ± SD
Absolute
change
from
control
1%1
N
0
0.23582 ± 0.054
-
0.34605 ± 0.066
-
25
250
0.27279 ±0.053
0.04
0.37895 ±0.098
0.03
25
500
0.23656 ±0.054
0.00
0.37615 ±0.073
0.03
25
100
0.28593* ±0.069
0.05
0.40207 ±0.1
0.06
25
F1 Offspring-Absolute Organ Weights
Dose
(mg/kg-d)
Mean testis
weight (left) (g)
± SD
% change
from control
Mean testis
weight (right) (g)
± SD
% change from
control
0
1.79 ±0.11
-
1.84 ±0.137
-
250
1.77 ±0.39
-1
1.75 ±0.337
-5
500
1.84 ±0.21
3
1.86 ±0.226
1
1,000
1.84 ±0.171
3
1.82 ±0.255
-1
Dose
Mean epididymis
weight (left) (g) ±
% change
Mean epididymis
weight (right) (g)
% change from
(mg/kg-d)
SD
from control
± SD
control
0
0.71683 ±0.11
-
0.75575 ±0.041
-
250
0.69636 ±0.123
-3
0.70512 ±0.148
-7
500
0.71904 ±0.123
0
0.75008 ±0.113
-1
1,000
0.6898 ±0.12
-4
0.71244 ±0.127
-6
Dose
(mg/kg-d)
Mean prostate
weight ± SD
% change
from control
Mean seminal
vesicle weight ±
SD
% change
from
control
N
0
1.470 ±0.311
-
1.71 ±0.295
-
24
250
1.48 ± 0.249
1
1.94 ±0.567
13
25
500
1.38 ±0.23
-6
1.86 ± 0.422
9
24
1,000
1.41 ±0.279
-4
1.92 ±0.436
12
25
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
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Gaoua (2004b) (continued)
F1 Offspring-Relative Organ Weights
Mean testis
weight: bodv
Mean testis
weight ratio
Absolute
weight: bodv
Absolute
Dose
(left) (g)
change from
weight ratio
change from
(mg/kg-d)
±SD
control (%)
(right) (g) ±SD
control (%)
0
0.30842 ± 0.065
-
0.31441 ±0.036
-
250
0.30222 ± 0.067
-0.01
0.29746 ± 0.059
-0.02
500
0.30679 ± 0.037
0.00
0.31004 ± 0.04
0.00
1,000
0.31198 ±0.042
Mean epididymis
0.00
0.30958 ± 0.05
0.00
weight (left):
Absolute
ean epididymis
Absolute
Dose
bodv weight
change from
weight (right) (g)
change from
(mg/kg-d)
ratio (g) ± SD
control (%)
± SD
control (%)
0
0.12299 ±0.023
-
0.12915 ±0.012
-
250
0.11863 ±0.021
0.00
0.12002 ±0.025
-0.01
500
0.1198 ±0.021
0.00
0.12492 ±0.018
0.00
1,000
0.11693 ±0.021
-0.01
0.12065 ±0.022
-0.01
Absolute
change
Mean prostate
Absolute
Mean seminal
from
Dose
weight: bodv
change from
vesicle: bodv
control
(mg/kg-d)
weight ratio ± SD
control (%)
weight ratio ± SD
1%1 N
0
0.25136 ±0.057
-
0.29278± 0.055
24
250
0.25239 ±0.043
0.00
0.33038 0.085
0.04 25
500
0.23059 ± 0.043
-0.02
0.3165 ±0.113
0.02 24
1,000
0.2374 ± 0.04
-0.01
0.32424 ± 0.073
0.03 25
Weng et al. (2014)
Wild Type Mice; 13-Week Exposure
mice, C57BL/6
inhalation - vapor
male (5/group): 0, 500,1,750,
5,000 ppm (0, 2,090, 7,320,
20,900 mg/m3)a
Dose
(mg/m3)
Mean testis:
bodv weight
ratio (%)
± SD
Absolute
change from
control (%)
Mean
epididymis: bodv
weight ratio (%)
± SD
Absolute
change from
control (%)
dynamic whole body
0
0.7 ±0.06
-
0.24 ±0.02
-
inhalation; 6 hr/d, 5 d/wk for
13 wk; methods described in
2,090
0.74 ± 0.04
0.04
0.26 ±0.02
0.02
Weng et al. (2012)
7,320
0.67 ±0.09
-0.03
0.25 ±0.01
0.01
20,900
0.7 ±0.02
0.00
0.24 ± 0.02
0.00
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
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Weng et al. (2014)
Knockout Mice (Aldh213-week Exposure
(continued)
Mean testis:
Mean
bodv weight
Absolute
epididymis: bodv
Absolute
Dose
ratio (%)
change from
weight ratio (%)
change from
(mg/m3)
±SD
control (%)
±SD
control (%)
0
0.76 ±0.04
-
0.26 ±0.01
-
2,090
0.71 ±0.11
-0.05
0.24 ±0.02
-0.02
7,320
0.72 ±0.05
-0.04
0.24* ±0.02
-0.02
20,900
0.71 ±0.07
-0.05
0.23* ±0.02
-0.03
Weng et al. (2014)
Wild Type Mice; 9-Week Exposure
mice, C57BL/6
inhalation - vapor
male (NR): 0, 50, 200, 500 ppm
(209, 836, 2,090 mg/m3)a
dynamic whole body
Mean testis:
Mean
Dose
(mg/m3)
bodv weight
ratio (%)
±SD
Absolute
change from
control (%)
epididymis: bodv
weight ratio (%)±
SD
Absolute
change from
control (%)
inhalation; 6 hr/d, 5 d/wk for 9
0
0.8 ±0.12
-
0.26 ±0.03
-
wk; methods described in
Weng et al. (2012)
209
0.77 ±0.09
-0.03
0.25 ±0.03
-0.01
836
0.77 ±0.09
-0.03
0.25 ±0.02
-0.01
2,090
0.78 ±0.08
-0.02
0.25 ±0.02
-0.01
Knockout
Mice (Aldh2-/-); 9-week Exposure
Mean testis:
Mean
bodv weight
Absolute
epididymis: bodv
Absolute
Dose
ratio (%)
change from
weight ratio (%)±
change from
(mg/m3)
±SD
control (%)
SD
control (%)
0
0.8 ±0.06
-
0.27 ±0.02
-
209
0.76 ±0.05
-0.04
0.26 ±0.02
-0.01
836
0.79 ±0.07
-0.01
0.27 ±0.01
0.00
2,090
0.74 ±0.01
-0.06
0.25 ±0.03
-0.02
Haplotype Mice (Aldh2 heterogeneous); 9-week Exposure
Mean testis:
Mean
bodv weight
Absolute
epididymis: bodv
Absolute
Dose
ratio (%)
change from
weight ratio (%)±
change from
(mg/m3)
±SD
control (%)
SD
control (%)
0
0.82 ±0.07
-
0.26 ±0.02
-
209
0.8 ±0.06
-0.02
0.26 ±0.01
0.00
836
0.81 ±0.09
-0.01
0.26 ±0.02
0.00
2,090
0.73 ±0.03
-0.09
0.24 ±0.01
-0.02
This document is a draft for review purposes only and does not constitute Agency policy.
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de Pevster et al. (2009)
Absolute Organ Weights
rat, Fischer 344
oral -gavage
PO, male (12/group): 0, 600,
Dose
(mg/kg-d)
Mean testis
weight (g) ± SD
% change
from control
Mean epididymis
weight (mg) ± SD
% change from
control
1,200,1,800 mg/kg-d
0
2.55 ±0.09
-
0.696 ±0.016
-
daily for 14 days
600
2.53 ±0.05
-1
0.693 ±0.027
0
1,200
2.49 ±0.07
-2
0.701 ±0.026
1
1,800
2.47 ±0.1
Mean prostate
-3
0.663 ±0.029
Mean seminal
-5
Dose
weight (g)
% change
vesicle weight (g)
% change from
(mg/kg-d)
± SD
from control
± SD
control
0
0.238 ±0.018
-
0.781 ±0.022
-
600
0.309 ± 0.034
30
0.733 ± 0.024
-6
1,200
0.252 ±0.018
6
0.749 ± 0.037
-4
1,800
0.269 ±0.036
13
0.701 ±0.041
-10
Mean weight of
Dose
combined accessory sex
(mg/kg-d)
organs (g)±SD % change from control
0
1.712 ±0.041
-
600
1.735 ±0.057
1
1,200
1.702 ± 0.063
-1
1,800
1.633 ±0.059
-5
Relative Organ Weights
Mean
Mean testis:
Absolute
epididymis: bodv
Absolute
Dose
bodv weight
change from
weight ratio (%)
change from
(mg/kg-d)
ratio (%) ± SD
control (%)
± SD
control (%)
0
0.997 ±0.036
-
0.272 ± 0.007
-
600
1.014 ± 0.027
0.02
0.275 ± 0.009
0.00
1,200
1.097 ± 0.03
0.10
0.308 ± 0.009
0.04
1,800
1.097 ± 0.045
0.10
0.294 ±0.014
0.02
Mean prostate:
Absolute
Mean seminal
Absolute
Dose
bodv weight
change from
vesicle: bodv wt.
change from
(mg/kg-d)
ratio (%) ± SD
control (%)
ratio (%) ± SD
control (%)
0
0.092 ± 0.007
-
0.304 ± 0.008
-
600
0.124 ±0.015
0.03
0.292 ±0.012
-0.01
1,200
0.111 ±0.076
0.02
0.328 ±0.012
0.02
1,800
0.123 ±0.021
0.03
0.31 ±0.017
0.01
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
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Resu Its
de Pevster et al. (2009)
(continued)
Mean combined
accessory sex
Dose organs:bodv weight ratio Absolute change from
(mg/kg-d) (%) ± SD control (%)
0 0.668 ±0.018
600 0.691 ± 0.026 0.02
1,200 0.746 ± 0.019 0.08
1,800 0.727 ± 0.035 0.06
Medinskv et al. (1999); Bond et
al. (1996b)
rat, Fischer 344
inhalation - vapor
male (48/group): 0, 500,1,750,
5,000 ppm
(0, 2,090, 7,320,
20,900 mg/m3)a; female
(48/group): 0, 500, 1,750,
5,000 ppm
(0, 2,090, 7,320,
20,900 mg/m3)a
dynamic whole body chamber;
6 hr/d, 5 d/wk for 13 wk
Organ weights of Fisher 344 rats and CD-I mice were not altered by exposure to
ETBE (data not shown).
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Reference and Study Design
Resu Its
Testosterone and Estradiol
de Pevster et al. (2009)
Dose
Mean plasma testosterone
rat, Fischer 344
(mg/kg-d)
N
(ng/ml) ±SE
% change from control
oral -gavage
P0, male (12/group): 0, 600,
0
12
2.07 ± 42
-
1,200,1,800 mg/kg-d
600
12
3.1 ±0.78
50
daily for 14 days
1,200
11
2.61 ±0.55
26
1,800
10
1.36 ±0.39
-34
Dose
Mean plasma estradiol
(mg/kg-d)
N
(pg/ml)
% change from control
0
12
1.085 ±0.1
-
600
12
1.395 ± 0.403
29
1,200
11
2.238* ±0.377
106
1,800
9
2.224* ±0.611
105
1 a4.18 mg/m3 = 1 ppm.
2 bConversion performed by study authors.
3 cMale mating index (%) = (No. males able to mate with at least one female / Total males) x 100.
4 dMale fertility index (%) = (No. males with at least one pregnant partner / Males that mated at least once) x 100
5 *: result is statistically significant (p < 0.05) based on analysis of data by study authors.
6 for controls, no response relevant; for other doses, no quantitative response reported.
7 % change from control = [(treated group value -control value)/control value] x 100.
8 Absolute change from control (%) = control value (%) - treated group value (%).
9
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
¦ = exposures at which the Midpoint was reported statistically significant by study authors
~ = exposures at which the endpoint was reported not statistically significant by study authors
Male Fertility
PO Male rat; copulation; one-gen repro (AJ
PO Male rats copulation; two-gen repro (B|
PI Male rat; copulation; two-gen repro (B)
FO Male rat; fertility; one-gen repro (AJ
F0 Male rat; fertility; two-gen repro (B)
F1 Male rat; fertility; two-gen repro (B)
Sperm PO Male rat; epididymis! sperm count,- one-gen repro (A)
Parameters
FO Male rat; eptdidymal sperm count; two-gen repro (g)
Ft Male rat; cpididymal sperm count; two-gen repro (B)
FO Male rat; epididymal sperm motility; one-gen repro (A)
PO Male rat; epididymal sperm motility; two-gen repro (B)
PI Male rat; cpididy mal sperm motility; two-gen repro (B)
FO Male rat; epididymal sperm morphology; one-gen repro
(A)
FO Male rat; epididvmat sperm morphology; two-gen repro
(B)
F1 Male rat; cpldidymal sperm morphology; two-gen repro
(B)
FO Male rat; testicular sperm heads; one-gen repro (A)
FO Male rat; testicular sperm heads; two-gen repro (13)
PI Male rat; testicular sperm heads; two-gen repro (B)
FO Male rat; daily sperm production; two-gen repro (B) -
F1 Male rat; daily sperm product ton; two-gen repro (BJ •
Testosterone
and Estradiol
Male rat; testosterone; 14 d (C)
Male rat; estradiol; 14 d (C)
o-
_a
o ~ o
~—B—E)
Q-
-S 0
o—a—0
f—I pi—
Q B~
-0
O—B—El
~—a—~
B B-
-0
B—B—H
Q 1 "D—0
0-
-0
G—B—0
o—a—a
L-mJ ^ I, i; If
a—b—0
Q—B—0
O—B—0
Q—B—0
fX-—
10 100
Dose (mg/kg-day)
Sources: (A) Fujii et al,, 2010; JPEC, 2008e (B) Gaoua, 2004b (C) de Peyster et al., 2009
1,000
10,000
Figure 1-11. Exposure-response array of male reproductive effects following
oral exposure to ETBE.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
¦ = exposures at which the endpoint was reported statistically significant by study authors
~ = exposures at which the endpoint was reported not statistically significant by study authors
Testicular
HiStOpathoIogy Male raf; spermatocyte degeneration] 13 wks (A
Male mouse; atrophy of seminiferous tubules; 13 wks (H
Male Aldh2-/- mouse; atrophy of seminiferous tubules; 13 wks(R
Male n ihim i «phy of seminiferous 'i In Us e< wks (H
Male AU1h2 -/- HmpSiy ofseminiferous iahui< s -t wks[H
Male Aldh2+/- mouse atrophy of seminiferous tutu le\ wks (B
Sperm
Parameters
Male mouse; < pulu'vn.is wt (relative); 9 »ks 1 HI
Male Aldh2*/~ mouse; deer epididymis wt (relative); 9 wk> [ Hj
CEH
B-
B-
B-
B-
Q-
B-
B-
B-
B-
B-
B-
B~
Q_
Q-
B-
™0"——El
-B E3
™B> O
E3~
-Bi El
-B ¦
-e-
O-
-B ~
-e—o
Q-
B-
B-
-e—E3
-&¦
_Q_
-B-
-B
-EH E3
-B-
-e—€3
-B-
-£]
100 1,000 10,000
Exposure Concentration (mg/in1)
Sources: (A) Medinskyet al, 1999; Bond et al, 1996b (B) Weng et al, 2014
100,000
Figure 1-12. Exposure-response array of male reproductive effects following
inhalation exposure to ETBE.
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Toxicological Review ofETBE
Mechanistic Evidence
No mechanistic evidence for male reproductive effects was identified by the literature
search.
Integration of Male Reproductive Effects
The male reproductive endpoints examined in this database were not consistently affected
across studies or across doses. The 13-week and 9-week inhalation studies conducted in rats and
mice fWengetal.. 2014: Medinskv etal.. 19991 provide suggestive evidence of ETBE-induced
testicular degeneration and effects on sperm count, sperm mobility, and sperm DNA damage. In
contrast, no male reproductive toxicity was observed in any of the other studies examined in this
database, including one- and two-generation reproductive toxicity studies, 2-year carcinogenicity
studies, and sub-chronic studies. For example, the 2-year inhalation carcinogenicity study (Saito et
al.. 2013: TPEC. 2010bl used the same rat strain, same route of exposure, and similar range of
exposure concentrations as Medinskv et al. f!9991 and did not observe any dose-related effects on
testicular histopathology. Wengetal. (20141. however, found that Aldh2 KO and heterogeneous
mice had consistently reduced numbers of sperm heads and sperm motility as well as reductions in
male reproductive organ weights, suggesting that populations with ALDH2 polymorphisms could
be susceptible to these effects from ETBE exposure (discussed in Section 1.3.3). The 14-day study
by de Pevster et al. (20091 observed increased estradiol and decreased testosterone in ETBE-
exposed rats, which is a potential mechanism for testicular degeneration; however, no effects on
testicular histopathology or organ weight were observed in this study. Collectively, the available
evidence is considered inadequate to draw conclusions regarding the male reproductive toxicity of
ETBE, and male reproductive effects are not carried forward as a hazard.
Synthesis of Effects Related to Female Reproduction
The available evidence for ETBE-induced effects on the female reproductive system
includes no human data. The evidence was obtained primarily from a one-generation reproductive
toxicity study fFuiii etal.. 2010: TPEC. 2008el. a two-generation reproductive toxicity study fGaoua.
2004b), and three developmental toxicity studies fAso etal.. 2014: Asano etal.. 2011: TPEC. 2008h.
ij Gaoua. 2004a). In addition, some evidence was obtained from two 90-day toxicity studies (TPEC.
2008b: Medinskv etal.. 1999: Bond etal.. 1996a). one subchronic (180-day) study (Mivata etal..
2013: TPEC. 2008c). two 2-year carcinogenicity studies (Saito etal.. 2013: Suzuki etal.. 2012: TPEC.
2010a. b), and a short-term study evaluating ETBE-induced oocyte effects (Berger and Horner.
20031. These studies evaluated the effects ofETBE exposure on maternal body weight change fAso
etal.. 2014: Asano etal.. 2011: Fuiii etal.. 2010: TPEC. 2008e. h, i; Gaoua. 2004a. b), fertility, mating,
and pregnancy parameters fFuiii etal.. 2010: TPEC. 2008e: Gaoua. 2004b: Berger and Horner.
2003). fecundity (Aso etal.. 2014: Asano etal.. 2011: Fuiii etal.. 2010: TPEC. 2008e. h, i; Gaoua.
2004a. b), estrous cyclicity fFuiii etal.. 2010: TPEC. 2008e: Gaoua. 2004b). and organ weights (Aso
etal.. 2014: Mivata etal.. 2013: Saito etal.. 2013: Suzuki etal.. 2012: Asano etal.. 2011: Fuiii etal..
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
2010: TPEC. 2010a. b, 2008b. c, e, h, ij Gaoua. 2004b: Medinskv etal.. 1999: Bond etal.. 1996a).
ETBE-induced effects were examined in pregnant rats and rabbits and non-pregnant female rats
after oral or whole body inhalation exposures, and the design, conduct, and reporting of each study
were of sufficient quality to inform human health hazard assessment Selected female reproductive
toxicity endpoints from these studies are summarized in Table 1-15.
The one- and two-generation reproductive toxicity studies and developmental studies
evaluated maternal toxicity and several endpoints related to fertility, pregnancy, and pregnancy
outcomes in rats and rabbits up to 1,000 mg/kg-day ETBE. Maternal toxicity, as shown by
decreased maternal body weight and corrected (for the gravid uterus) body weight, was observed
following gestational exposure to 1,000 mg/kg-day ETBE from GD 5-19; however, this effect was
not observed in another developmental exposure study in which ETBE was administered at the
same dose and exposure duration (Aso etal.. 2014: TPEC. 2008h). Further, administration ofETBE
during the pre-mating through lactation periods in parental and F1 generations fFuiii etal.. 2010:
TPEC. 2008e: Gaoua. 2004b 1 did not affect maternal body weight parameters in rats. Maternal body
weight during the entire pregnancy (GD 0-28) and corrected body weight change were decreased
in rabbits administered 1,000 mg/kg-day ETBE fAsano etal.. 2011: IPEC. 20080: however, the lack
of change in body weight during the treatment period (GD 6-27), the lack of a dose-related
response, and the inherent variability in body weight parameters during pregnancy in rabbits (U.S.
EPA. 1991b) limit the interpretation of this effect ETBE did not affect indices of mating or fertility,
and pre-coital times and gestation lengths were similarly unaffected in rats in the parental fFuiii et
al.. 2010: TPEC. 2008e: Gaoua. 2004b 1 and the F1 generation fGaoua. 2004b). In addition, the
number of corpora lutea in pregnant rats and rabbits fAso etal.. 2014: Asano etal.. 2011: TPEC.
2008h. i), the average estrous cycle length, and the percent of females with normal estrous cycles
fFuiii etal.. 2010: TPEC. 2008e) were not significantly affected by ETBE when compared to control
values. Further supporting these findings, oocyte quality and fertilizability was shown to be
unaffected by ETBE (Berger and Horner. 2003). Litter size was evaluated by Fuiii etal. (2010). TPEC
C2008e). Gaoua f2004b1. Aso etal. C2014). TPEC C2008h). and Asano etal. <"20111. TPEC C2008i). and
no significant, dose-related effects were observed in rats or rabbits following ETBE exposure.
Reproductive organ weights were also reported after oral and inhalation exposures to
ETBE. Gravid uterine weights were not affected following ETBE exposure during gestation in
rabbits fAsano etal.. 2011: TPEC. 2008i) nor were ovary and uterine weights affected after exposure
during pre-mating through lactation periods in rats fFuiii etal.. 2010: TPEC. 2008e). Consistent with
these findings, ovary and uterine weights in non-pregnant females were not affected by ETBE after
90-day inhalation flPEC. 2008b: Medinskv etal.. 1999: Bond etal.. 1996a). 180-day oral fMivata et
al.. 2013: TPEC. 2008c). and 2-year oral fSuzuki etal.. 2012: TPEC. 2010a) exposure assessments. In
a 2-year inhalation study in rats fSaito etal.. 2013: TPEC. 2010b). however, a significant increase in
relative ovary weight was observed at exposures of 1,500 and 5,000 ppm ETBE.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Table 1-15. Evidence pertaining to female reproductive effects in animals
exposed to ETBE
Reference and study design
Results
Maternal Body Weight
Gaoua (2004a)
Net bodv wt
rat, Sprague-Dawley
Bodv wt change ±
% change
change ± SD
% change from
oral -gavage
Dose (mg/kg-d)
SD, GD 5-20 (g)
from control
M
control
P0, female (24/group): 0, 250,
500,1,000 mg/kg-d
0
132 ± 22
-
61.8 ± 13
-
dams exposed from GD 5 to
250
132 ± 12
-2
59.4 ±8.1
-4
GD 19
500
134 ± 19
-1
60 ± 11.3
-3
1,000
120* ±15
-11
51.5* ± 10.3
-17
Gaoua(2004b)
Fl: Bodv wt
rat, Sprague-Dawley
F0: Bodv wt
% change
change ± SD
% change from
oral - gavage
Dose (mg/kg-d)
change ± SD (g)
from control
M
control
F0, male and female
0
132 ± 15
146 ± 21
(25/sex/group): 0, 250, 500,
1,000 mg/kg-d
250
134 ± 14
2
145 ± 15
-1
dosed daily for 18 wks from
10 wks premating until weaning
500
136 ± 25
3
141 ± 21
-3
of F1 pups
1,000
136 ± 12
3
137 ± 12
-6
Fl, male and female (24—
25/group): 0, 250, 500, 1,000
mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Aso et al. (2014); JPEC (2008h)
Bodv wt
rat, Sprague-Dawley
Bodv wt ± SD,
Bodv wt ± SD,
change ± SD,
% change from
oral - gavage
Dose (mg/kg-d)
GD 5 (g)
GD 20 (g)
GD 5-20 (g)
control
female (24/group): 0,100, 300,
1,000 mg/kg-d
0
280.9 ± 16.7
394.4 ± 26.9
113.5
-
dams dosed daily from GD 5 to
100
273.4 ± 10.8
380.3 ± 23.9
106.9
-6
GD 19
C-section GD 20
300
280 ± 13.4
389.8 ±25.9
109.8
-3
1,000
277.7 ± 15.9
382.4 ±27.1
104.7
-8
Fuiiietal. (2010); JPEC (2008e)
F0: Bodv wt
rat, Sprague-Dawley
change ± SD,
% change
oral - gavage
Dose (mg/kg-d)
GD 5-20 (g)
from control
F0, male and female
0
124.9 ± 22
(24/sex/group): 0,100, 300,
1,000 mg/kg-d; dosed daily for
100
119.6 ±20.3
-4
17 wks, from 10 wks premating
300
1,000
135.2 ±21.5
140.2* ± 19.1
8
12
to lactation day 21
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Toxicological Review ofETBE
Reference and study design
Results
Asano et al. (2011); JPEC (2008i)
Bodv wt
Adjusted
rabbit, New Zealand White
change ±
% change
Bodv wt
bodv wt
% change
oral -gavage
SD, GD6-28
from
change ± SD,
change ±
from
female (24/group): 0,100, 300,
Dose (mg/kg-d)
(kgl
control
GD 0-28 (kg)
SD (kg)
control
1,000 mg/kg-d
dams dosed daily from GD 6 to
0
0.26 ±0.12
-
0.40 ±0.12
0.02 ±0.14
-
GD 27
100
0.23 ±0.12
-12
0.35 ±0.12
-0.06 ±0.12
-400
C-section GD 28
300
0.28 ±0.08
8
0.40 ± 0.08
0±0.1
-100
1,000
0.12 ±0.19
-54
0.25* ±0.21
-0.07 ±0.19
-450
Fertility, Mating, and Pregnancy
Fuiiietal. (2010); JPEC (2008e)
Copulation
Fertilitv
rat, Sprague-Dawley
Dose (mg/kg-d)
index0 (%)
indexd (%)
oral - gavage
FO, male and female
0
100
87.5
(24/sex/group): 0,100, 300,
1,000 mg/kg-d; dosed daily for
100
95.8
100
17 wks, from 10 wks premating
300
100
95.8
to lactation day 21
1,000
100
91.7
Gaoua(2004b)
Pregnant/ma
Fertilitv
rat, Sprague-Dawley
Dose
Pregnant/mated Fertilitv index, F0 ted females.
index, Fl
oral - gavage
(mg/kg-d)
females, F0
i%l
£1
(%)
F0, male and female
(25/sex/group): 0, 250, 500,
0
23/25
92
22/25
88
1,000 mg/kg-d
250
21/25
84
22/24
92
dosed daily for 18 wks from
10 wks premating until weaning
500
22/25
88
22/25
88
of F1 pups
1,000
25/25
100
22/23
96
Fl, male and female (24—
25/group): 0, 250, 500, 1,000
mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Aso et al. (2014); JPEC (2008h)
rat, Sprague-Dawley
oral - gavage
Dose Mean no. corpora % change from
(mg/kg-d) lutea ± SD control
female (24/group): 0,100, 300,
0
15.5 ± 1.54
-
1,000 mg/kg-d
dams dosed daily from GD 5 to
100
14.1 ± 1.48
-9
GD 19
300
14.4 ± 1.85
-7
C-section GD 20
1,000
14.6 ± 2.44
-6
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Reference and study design
Results
Litter Size
Fuiiietal. (2010); JPEC (2008e)
Dose
Mean no. duds % change from
rat, Sprague-Dawley
(mg/kg-d)
delivered ± SD (kg) control
oral -gavage
FO, male and female
0
11.8 ±3.2
(24/sex/group): 0,100, 300,
100
10.4 ±3.4 -12
1,000 mg/kg-d; dosed daily for
300
12.1 ±2.3 3
17 wks, from 10 wks premating
to lactation day 21
1,000
13.0 ±1.9 10
Gaoua(2004b)
% change
rat, Sprague-Dawley
Dose
Litter size at birth, from control.
Pregnant/mated % change from
oral - gavage
(mg/kg-d)
IS
IS
females, Fl control, Fl
F0, male and female
0
14.3
13.7
(25/sex/group): 0, 250, 500,
1,000 mg/kg-d
250
14.1 -1
13.7 0
dosed daily for 18 wks from
500
14.9 4
13.7 0
10 wks premating until weaning
of F1 pups
1,000
14.2 -1
14 2
Fl, male and female (24—
25/group): 0, 250, 500, 1,000
mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Aso et al. (2014); JPEC (2008h)
Dose
rat, Sprague-Dawley
(mg/kg-d)
Mean no. live fetuses ± SD (kg)
% change from control
oral - gavage
female (24/group): 0,100, 300,
0
13.6 ± 1.5
-
1,000 mg/kg-d
100
12.0 ±2.65
-12
dams dosed daily from GD 5 to
GD 19
300
12.6 ±2.58
-7
C-section GD 20
1,000
12.3 ±2.8
-10
Asano et al. (2011); JPEC (2008i)
Dose
rabbit, New Zealand White
(mg/kg-d)
Mean no. live fetuses ± SD (kg)
% change from control
oral - gavage
female (24/group): 0,100, 300,
0
7.8 ±3.1
-
1,000 mg/kg-d
100
7.9 ±3.2
1
dams dosed daily from GD 6 to
GD 27
300
8.4 ±2.0
8
C-section GD 28
1,000
6.9 ±3.2
-12
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Reference and study design
Results
Gestation Length
Fuiiietal. (2010); JPEC (2008e)
Dose
rat, Sprague-Dawley
(mg/kg-d)
Mean gestation length ± SD (davs) % change from control
oral -gavage
FO, male and female
0
22.2 ±0.4
-
(24/sex/group): 0,100, 300,
100
22.1 ±0.4
0
1,000 mg/kg-d; dosed daily for
300
22.2 ±0.4
0
17 wks, from 10 wks premating
to lactation day 21
1,000
22.6 ±0.5
2
Gaoua(2004b)
Dose
Gestation length % change from Gestation length
% change from
rat, Sprague-Dawley
(mg/kg-d)
(davs), F0 control, F0
(davs), Fl
control, Fl
oral - gavage
F0, male and female
0
21.7
21.5
-
(25/sex/group): 0, 250, 500,
250
21.5 -1
21.6
0
1,000 mg/kg-d
dosed daily for 18 wks from
500
21.5 -1
21.6
0
10 wks premating until weaning
1,000
21.8 0
21.6
0
of F1 pups
Fl, male and female (24—
25/group): 0, 250, 500, 1,000
mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Estrous Cyclicity
Fuiiietal. (2010); JPEC (2008e)
Mean estrous
rat, Sprague-Dawley
Dose
% Females w/normal estrous
cvcle length ± SD
% change
oral - gavage
(mg/kg-d)
cvcles, F0
(davs)
from control
F0, male and female
0
91.7
4.03 ± 0.09
(24/sex/group): 0,100, 300,
1,000 mg/kg-d; dosed daily for
100
97.1
4.1 ±0.29
2
17 wks, from 10 wks premating
to lactation day 21
300
97.1
4.06 ±0.17
1
1,000
95.8
4.29 ±0.61
6
Organ Weights
Fuiiietal. (2010); JPEC (2008e)
Absolute Weight
rat, Sprague-Dawley
oral - gavage
F0, male and female
Dose
(mg/kg-d)
Meanovarvwt± % change
SD (mg) from control
Mean uterus wt
± SD (mg)
% change from
control
(24/sex/group): 0,100, 300,
0
98.8 ±14.9
468 ± 68
-
1,000 mg/kg-d; dosed daily for
100
92.5 ±16.6 -6
513 ±151
10
17 wks, from 10 wks premating
to lactation day 21
300
95.3 ±11.1 -4
523 ±157
12
1,000
100.9 ± 16.9 2
516 ±136
10
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Reference and study design
Results
Fuiiietal. (2010); JPEC (2008e)
Relative Weight
(continued)
Dose
Mean ovarv wt ±
% change
Mean uterus wt
% change from
(mg/kg-d)
SD(mg)
from control
±SD(mg)
control
0
30.7 ±4.7
-
146 ± 27
-
100
28.6 ± 6
-7
158 ±49
8
300
29.3 ±3.6
-5
162 ± 53
11
1,000
29.9 ±4.9
-3
154 ± 46
5
Gaoua(2004b)
% change
rat, Sprague-Dawley
Dose
Mean ovarv Wt. ±
from
Mean uterus Wt. ±
% change
oral -gavage
(mg/kg-d)
SD (g)
control
SD(g)
from control
FO, male and female (19-
25/sex/group): 0, 250, 500,
Absolute Weight, F0
1,000 mg/kg-d
0
0.168 ±0.025
-
0.54 ± 0.096
-
dosed daily for 18 wks from
10 wks premating until weaning
250
0.167 ±0.027
-1
0.587 ±0.231
9
of F1 pups
500
0.167 ±0.022
-1
0.483 ±0.102
-11
Fl, male and female (19—
25/group): 0, 250, 500,
1,000
0.164 ±0.023
-2
0.576 ±0.218
7
1,000 mg/kg-d
dosed daily from PND 22 until
Absolute Weight, Fl
weaning of F2 pups
0
0.164 ±0.027
-
0.557 ±0.13
-
250
0.172 ±0.028
5
0.577 ±0.161
4
500
0.168 ±0.031
2
0.538 ±0.141
-3
1,000
0.163 ±0.049
-1
0.547 ±0.122
-2
Medinskv et al. (1999); Bond et
Dose
al. (1996b)
(mg/m3)
Mean ovarv wt
±SD(g)
% change from control
rat, Fischer 344
inhalation - vapor
0
0.085 ± 0.022
-
male (48/group): 0, 500,1,750,
2,090
0.095 ±0.016
12
5,000 ppm
(0, 2,090, 7,320,
7,320
0.088 ±0.12
4
20,900 mg/m3)a; female
20,900
0.090 ±0.19
6
(48/group): 0, 500, 1,750,
5,000 ppm
(0, 2,090, 7,320, 20,900 mg/m3)a
dynamic whole body chamber;
6 hr/d, 5 d/wk for 13 wk
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Reference and study design
Results
Asano et al. (2011); JPEC (2008i)
Dose
rabbit, New Zealand White
(mg/kg-d)
Gravid uterus wt ± SD (g)
% change from control
oral -gavage
female (24/group): 0,100, 300,
0
383 ± 98
-
1,000 mg/kg-d
100
398 ±128
4
dams dosed daily from GD 6 to
GD 27
300
403 ± 91
5
C-section GD 28
1,000
323 ±128
-16
Mivata etal. (2013); JPEC
Dose
Mean absolute ovarv wt ± SD
(2008c)
(mg/kg-d)
(mg)
% change from control
rat, Sprague-Dawley
oral - gavage
0
70.0 ± 18.7
-
male (15/group): 0, 5, 25,100,
5
71.0 ±21.7
1
400 mg/kg-d; female
(15/group): 0, 5, 25,100,
25
73.8 ± 16.6
5
400 mg/kg-d
100
67.7 ± 17.7
-3
daily for 180 days
400
Dose
76.6 ± 18.2
Mean relative ovarv wt ± SD
9
(mg/kg-d)
(mg/lOOg)
% change from control
0
20.4 ± 5.4
-
5
21.4 ±5
5
25
21.8 ±4.8
7
100
20.0 ± 4.9
-2
400
22.8 ±5.5
12
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Reference and study design
Results
JPEC (2008b)
% change
% change
rat, Sprague-Dawley
Dose
Mean ovarv wt
from
Mean uterus
from
inhalation - vapor
(mg/m3)
N
±SD (mg)
control
wt ± SD (g)
control
male (10/group): 0,150, 500,
1,500, 5,000 ppm
Absolute Weight, Day 92
(0, 627, 2,090, 6,270,
0
10
91.47 ± 10.26
-
0.709 ±0.222
-
20,900 mg/m3)a; female
(10/group): 0, 150, 500, 1,500,
627
10
87.36 ± 15.83
0
0.819 ±0.38
16
5,000 ppm (0, 627, 2,090, 6,270,
2,090
10
84.92 ± 16.91
0
0.654 ±0.159
-8
20,900 mg/m3)a
dynamic whole body chamber;
6,270
10
78.39 ±9.83
0
0.712 ±0.198
0
6 hr/d, 5 d/wk for 13 wk;
20,900
10
91.94 ±21.84
0
0.702 ± 0.205
-1
generation method, analytical
concentration, and method
Absolute Weight, Day 120
reported
0
627
2,090
6,270
6
82.82 ± 17.89
-
0.965 ±0.332
-
20,900
6
90.38 ± 15.88
9
0.818 ±0.286
-15
Relative Weight, Day 92
0
10
27.19 ±3.8
-
0.21 ±0.066
-
627
10
27.58 ±4.35
1
0.269 ±0.151
28
2,090
10
27.03 ±4.55
0
0.211 ±0.055
0
6,270
10
25 ±2.67
-6
0.228 ±0.061
9
20,900
10
30.39 ± 6.46
9
0.231 ±0.071
10
Relative Weight, Day 120
0
6
25.02 ± 4.03
-
0.298 ±0.107
-
627
-
-
-
-
-
2,090
-
-
-
-
-
6,270
-
-
-
-
-
20,900
6
26.72 ±4.79
7
0.24 ± 0.089
-19
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Reference and study design
Results
JPEC (2010a)
rat, Fischer 344
oral - water
male (50/group): 0, 625, 2,500,
10,000 ppm (0, 28, 121, 542
mg/kg-d)b; female (50/group): 0,
625, 2,500,10,000 ppm (0, 46,
171, 560 mg/kg-d)b
daily for 104 wk
Dose
(mg/kg-d) Mean ovarv wt ± SD (g) % change from control
0 0.194 ±0.238
46 0.18 ±0.146 -7.21649
171 0.153 ±0.035 -21.134
560 0.147 ±0.023 -24.2268
1
2 a4.18 mg/m3 = 1 ppm.
3 bConversion performed by study authors.
4 Population index (%) = (no. of rats with successful copulation/no. of rats paired) x 100.
5 fertility index (%) = (no. females pregnant or no. of males sired/no. of rats with successful copulation) x 100.
6 *: result is statistically significant (p < 0.05) based on analysis of data by study authors.
7 for controls, no response relevant; for other doses, no quantitative response reported.
8 % change from control = (control value - treated group value)/control value] x 100.
9 Absolute change from control (%) = control value (%) - treated group value (%).
10
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
¦ = exposures at which the Midpoint was reported statistically significant by study authors
~ = exposures at which the endpoint was reported not statistically significant by study authors
Pregnancy po Female rat; copulation; one-gen repro (C)
Outcomes
FO Female rat; fertility; one-gen repro [C)
FO Female rat; fertility; two-gen repro (E)
PI Female rat; fertility; two-gcn repro (E)
Female rat; maternal body weight gafn; GD S-19 (D)
FO Female rat; maternal body weight gain; two-gen repro (E)
F1 Female rat; maternal body weight gain; two-gen repro (E)
Female rat; maternal body weight gain; G0 S-19 (B)
FO Female rat; maternal body weight gain; one-gen repro (C)
Female rabbit; maternal body weight gain; G0 6-27 (A)
Female rabbit; gravid uterine weight; GD6-27 (A)
FO Female rat; gestation length; one-gen repro (C)
FO Female rat; gestation length; two-gen repro (E)
Ft Female rat; gestation length; two-gen repro (E)
FO Female rat; litter size; one-gen repro (C)
FO Female rat; litter size; two-gen repro (E)
F1. Female rat; litter size; two-gen lepra (E)
Female rat; litter size; GD 5-19 (B)
Female rabbit; litter size; GD6-27 (A)
FO Female r«it; estrous cyclicfty; one-gen repro (C)
Kstrous Cyclic! ty
10
* maternal weight 14,1m sii»nilkanllv mi reaseil m this study
whereas other studies showed a significant tin rease
Q-
-e-
-B
fcj fcj
~—b—a
~—a—a
~ B—El
Q-
Q-
-B-
B-
Q-
-e
13~
-B
¦a—a
P*~
B—B—€3
—0 0
o-
~ B 0
O B—S
—a a
o-
-B-
100 1,000
Dose (mg/kg-day)
10,000
Sources: (A) Asano et. al., 2011; JPEC. 2008h (B)Aso et al, 2014; JPEC, 2008g (C) Fujiiet al., 2010; JPEC, 2008e
(D) Gaoua, 2004a (E) Gaoua, 2004b
2
3
Figure 1-13. Exposure-response array of female reproductive effects following
oral exposure to ETBE.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
¦ = exposures at which the endpoint was reported statistically significant by study authors
~ = exposures at which the endpoint was reported not statistically significant by study authors
Female Reproductive Organ Weights
Female rat; ovary wt. (absolute); 13 wks (B) ¦ i ~ E3 B
Female rat; ovary wt (absolute); 13 wks (A) 1 13—i B B—s El
Female rat; ovary wt. (relative); 13 wks (A) i E3—i B B—i Q
Female rat; uterine wt. (absolute); 13 wks (A) 1 ~— O —-0—; 0
Female rat; uterine wt. (relative); 13 wks (A) - ~—: B B—: B
100 1,000 10,000 100,000
Exposure Concentration (mg/ra3)
Source: (A) (PEC 2008b (B) Medinskyet al, 1999; Bond et al, 1996b
3 Figure 1-14. Exposure-response array of female reproductive effects following
4 inhalation exposure to ETBE.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Mechanistic Evidence
No mechanistic evidence for female reproductive effects was identified by the literature
search.
Integration of Female Reproductive Effects
The available evidence to assess female reproductive effects consists of one- and two-
generation reproductive toxicity studies, developmental toxicity studies, and 90-day through 2-year
oral and inhalation exposure studies that adequately evaluate the relevant female reproductive
endpoints. These studies show that ETBE does not adversely affect maternal body weight gain,
fertility, mating, pregnancy parameters, or reproductive organ weights in all but one study up to
1,000 mg/kg-day (oral exposure) or 5,000 ppm (whole body inhalation exposure) in the female rat
or rabbit. Relative ovary weights were significantly increased following ETBE inhalation exposure
in one 2-year study but not observed in other 2-year, 180-/90-day, reproductive, or developmental
studies, and the explanation for this observation is unclear without additional information.
Collectively, the available evidence is considered inadequate to draw conclusions regarding the
female reproductive toxicity ofETBE, and female reproductive effects are not carried forward as a
hazard.
1.2.4. Developmental Effects
Synthesis of Effects Related to Development
The database examining developmental effects following ETBE exposure includes no human
data; it is composed of data from toxicology studies conducted in Sprague-Dawley rats or New
Zealand White rabbits in which ETBE was administered via oral gavage. These consisted of three
prenatal developmental toxicity studies [two in rats: (Aso etal.. 2014: TPEC. 2008h) and (Gaoua.
2004a) and one in rabbits: (Asano etal.. 2011: TPEC. 2008i)]. a one-generation reproductive toxicity
study in rats fFuiii etal.. 2010: IPEC. 2008e). and a two-generation reproductive toxicity study in
rats f Gaoua. 2004al. The design, conduct, and reporting of all five studies were of sufficient quality
to inform human health hazard assessment. The highest dose level tested in each study was
1,000 mg/kg-d, the recommended limit dose for prenatal developmental toxicology studies (OECD.
2001: U.S. EPA. 1998cl.
Developmental endpoints evaluated after ETBE exposure include prenatal and postnatal
survival, growth, and morphological development In addition, limited assessments of postnatal
neurological functional development were conducted. Selected developmental toxicity data are
summarized in Table 1-16.
Evidence of effects ofETBE treatment on pre- or postnatal survival was minimal. In the
developmental toxicity study in rats by (Aso etal.. 2014: TPEC. 2008h). increased preimplantation
loss was observed in the treated groups. The percent preimplantation loss in the 1,000 mg/kg-day
dams was 81.8% greater than control, while it was increased 37.9% at 100 mg/kg-day and 21.2%
This document is a draft for review purposes only and does not constitute Agency policy.
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3
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5
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7
8
9
10
11
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13
14
15
16
17
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21
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23
24
25
26
27
28
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38
Toxicological Review ofETBE
at 300 mg/kg-day. Statistical significance was not reported. Increased preimplantation loss was not
observed in the other available developmental toxicity studies in rats or rabbits [fGaoua. 2004al
and (Asano etal.. 2011: IPEC. 20080. respectively]. Postnatal survival was not affected by ETBE
treatment in either the first or second generation of the reproductive toxicity study by Gaoua
(2004b). Viability indices throughout the lactation period were similar between control and treated
groups during both generations of this study. In the one-generation reproductive toxicity study
fFuiii etal.. 2010: IPEC. 2008el. there was evidence of a non-significant decrease (10.5% as
compared to control) in the PND 4 viability index at 1,000 mg/kg-day. Examination of the
individual animal data indicated that total litter loss in three litters had resulted in the majority of
pup deaths that occurred from PND 0-4. For two of these litters, severe maternal toxicity had led to
moribund sacrifice of the dams in early lactation; this is the only evidence in the available ETBE
data where adverse outcomes in the offspring were definitively associated with maternal toxicity.
The third dam with total litter loss had no evidence of treatment-related toxicity.
Neither prenatal nor postnatal growth were affected by ETBE treatment Mean fetal weights
were comparable between control and ETBE-treated groups in the prenatal developmental toxicity
studies in rats and rabbits fAso etal.. 2014: Asano etal.. 2011: IPEC. 2008h. ij Gaoua. 2004al.
Similarly, pup weights from PND 0-21 were not affected by treatment in the reproductive toxicity
studies fFuiii etal.. 2010: IPEC. 2008e: Gaoua. 2004b). Additionally, fFuiii etal.. 2010: IPEC. 2008e)
no effects were observed in the rate of completion of development landmarks in male and female
F1 offspring, specifically pinna detachment on PND 3, incisor eruption on PND 11, and eye opening
on PND 15. Organ weights (brain, spleen, and thymus) were evaluated in PND 21 pups in the one-
and two-generation reproduction studies fFuiii etal.. 2010: IPEC. 2008e: Gaoua. 2004bl: no
significant differences were observed between control and treated groups (not shown in evidence
table). At the termination of adult animals in the reproductive toxicity studies, a number of organ
weights were measured. Sections 1.2.1 and 1.2.2 discuss increased mean kidney and liver weights,
respectively, observed in the adult F1 offspring of the two-generation reproduction study fGaoua.
2004b). The findings in the F1 adults were similar to those in the P adults, indicating an absence of
life stage-related susceptibility for these outcomes.
No evidence existed of treatment-related effects on postnatal morphological assessments
that consisted of PND 1 anogenital distance measurements in F1 and F2 pups fGaoua. 2004bl and
the age of F1 sexual maturation (preputial separation in males and vaginal opening in females)
fFuiii etal.. 2010: IPEC. 2008e: Gaoua. 2004b).
In the prenatal developmental toxicity studies with ETBE (Aso etal.. 2014: Asano etal..
2011: IPEC. 2008h. ij Gaoua. 2004al. the evidence of treatment-related alterations in fetal
development at 1,000 mg/kg-day were sporadic, and there was no consistent pattern of effect
In Aso etal. f20141. a >3-fold increase in the number and percent of rat fetuses with skeletal
variations was noted at 1,000 mg/kg-day compared to control. Examination of the individual litter
data revealed that this increase was primarily attributable to a statistically significant >6-fold
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Toxicological Review ofETBE
increase in the number of fetuses (and >3-fold increase in the number of litters) with rudimentary
lumbar rib at that dose. The study authors dismissed the relevance of this finding, reporting that it
is within a historical control range (1.1-21.2%) for the strain of rat used in the study and because
the effect has sometimes been viewed as transient [e.g., (Chernoff etal.. 1991)]. Nevertheless, the
incidence of this finding is significantly increased as compared to the concurrent control, which is
considered more relevant and preferable to historical control and the finding might have been the
result of an alteration of vertebral development; therefore, it is considered potentially treatment-
related.
In Gaoua f2004al. a statistically significant 37% increase in the number of fetuses with
unossified 4thmetacarpal as compared to control was observed at 1,000 mg/kg-day. Further
evaluation of the fetuses, which were double-stained with alcian blue, revealed that a cartilage
precursor was present, suggesting that the finding represented a treatment-related delay in
development rather than a malformation.
An increase in the number of rabbit fetuses and litters with visceral malformations at 1,000
mg/kg-day was noted in Asano etal. f20111 and TPEC f2008il. This was specifically attributed to
observations of fetuses with absent right atrioventricular valve of the heart. The incidences of this
finding did not achieve statistical or biological significance. Also in Asano etal. (2011) and TPEC
(2008i). a 66% increase in the number of rabbit fetuses with skeletal variations at 1,000 mg/kg-day
as compared to control was found to be primarily attributed to incidences of unossified talus (in 12
fetuses, 6 litters).
Limited evaluation of postnatal functional neurological development in F1 male and female
offspring in reproductive toxicity studies were conducted by Fuiii etal. f20101. TPEC f2008el. and
Gaoua f2004bl. No treatment-related effects were found in assessments of reflex ontogeny, which
included surface righting reflex on PND 5 (Fuiii etal.. 2010: TPEC. 2008e: Gaoua. 2004b). negative
geotaxis on PND 8 (Fuiii etal.. 2010: TPEC. 2008e). cliff avoidance on PND 11 (Gaoua. 2004b). and
air righting reflex on PND 17 (Gaoua. 2004b) or PND 18 (Fuiii etal.. 2010: TPEC. 2008e). Gaoua
f2004bl also conducted tests in F1 males and females of acoustic startle response [postnatal week
(PNW) 4], pupil constriction (PNW 4), and motor activity (PNW 7 and 8). The motor activity testing
was performed using an automated device that measured the number of movements within the
front or back of the cage, back and forth movements, and vertical movements. Two 10-minute trials
were conducted 1 week apart. No treatment-related effects were found.
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Toxicological Review ofETBE
Table 1-16. Evidence pertaining to developmental effects in animals following
exposure to ETBE
Reference and study design
Resu Its
Prenatal Survival
Aso et al. (2014); JPEC (2008h)
% change % change
rat, Sprague-Dawley
Dose
No. No. preimplan-
from % Preimplan- from
oral -gavage
(mg/kg-d)
Litters tation loss
control tation loss3 control
female (24/group): 0,100, 300,
1,000 mg/kg-d
0
21 22
6.6
dams dosed daily from GD 5 to
100
22 25
13.6 9.1 37.9
GD 19
C-section GD 20
300
20 25
13.6 8.0 21.2
1,000
22 39
77.3 12.0 81.8
Dose
(mg/kg-d)
No. resorptions
% Postimplantation lossb
0
18
5.8
100
22
7.2
300
12
4.2
1,000
13
5
Gaoua (2004a)
Dose
No. preimplantation loss
rat, Sprague-Dawley
(mg/kg-d)
No. Litters
% Preimplantation loss3
oral - gavage
female (24/group): 0, 250, 500,
0
21
48 17.8
1,000 mg/kg-d
250
19
36 14.9
dosed daily from GD 5 to GD 19
C-section GD 20
500
20
38 14.3
1,000
22
47 16.8
Dose
(mg/kg-d)
No. Postimplantation loss % Postimplantation lossb
0
14
5.2
250
16
6.6
500
19
7.2
1,000
21
7.5
Asano et al. (2011); JPEC (2008i)
Dose
% Postimplantation
rabbit, New Zealand White
(mg/kg-d)
No. litters % Preimplantation loss3 lossb
oral - gavage
female (24/group): 0,100, 300,
0
22
19.6 11.0
1,000 mg/kg-d
100
22
15.3 11.3
dams dosed daily from GD 6 to
GD 27
300
20
10.7 7.0
C-section GD 28
1,000
23
22.9 8.7
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Toxicological Review ofETBE
Reference and study design
Resu Its
Postnatal Survival
Fuiiietal. (2010); JPEC (2008e)
% change
rat, Sprague-Dawley
from
oral -gavage
Dose
Viability index
Viabilitv index
control
Total litter loss
FO, male and female
(mg/kg-d)
PND0±SD
PND 4 ± SD
(PND 4)
(PND 0-4)c
(24/sex/group): 0,100, 300,
1,000 mg/kg-d; dosed daily for
0
98.9 ±3.7
97.4 ±4.7
-
0
17 wks, from 10 wks premating
100
97.9 ±5.6
96.7 ±8.1
-0.7
0
to lactation day 21
300
99.5 ±2.6
99.6 ± 1.9
2.3
0
1,000
93.6 ± 15.5
87.2 ±29.8
-10.5
3
Dose
(mg/kg-d)
Viabilitv Index - PND 21 ± SD
0
97 ± 11.1
100
95.8 ± 11.4
300
95.7 ± 11.1
1,000
92.5 ±23.1
Gaoua (2004b)
Dose
Viabilitv index
Viabilitv index Total litter loss
Viabilitv index
rat, Sprague-Dawley
(mg/kg-d)
PND 0
PND 4
(PND 0-4)
PND 21
oral - gavage
Fl
F0, male and female
(25/sex/group): 0, 250, 500,
0
100
97.6
0
94.6
1,000 mg/kg-d
250
100
92.9
91.7
dosed daily for 18 wks from
1
10 wks premating until weaning
500
100
82.3
0
96.1
of F1 pups
Fl, male and female (24—
1,000
100
97.7
1
99.5
25/group): 0, 250, 500,
F2
1,000 mg/kg-d
dosed daily from PND 22 until
0
100
97.6
0
97.6
weaning of F2 pups
250
100
94.8
0
98.8
500
100
97.0
3
100
1,000
100
92.9
0
99.3
Prenatal Growth
Aso et al. (2014); JPEC (2008h)
Dose
Mean fetal weight ± SD Mean fetal weight ± SD
rat, Sprague-Dawley
(mg/kg-d)
No. litters
male (g)
female (g)
oral - gavage
female (24/group): 0,100, 300,
0
21
4.1 ±0.3
3.89 ±0.25
1,000 mg/kg-d
100
22
4.14 ±0.33
3.92 ±0.23
dams dosed daily from GD 5 to
GD 19
300
20
4.23 ±0.22
4.01 ±0.22
C-section GD 20
1,000
22
4.14 ±0.34
3.91 ±0.39
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Reference and study design
Resu Its
Gaoua (2004b)
Dose
Mean fetal weight ± SD Mean fetal weight ± SD
rat, Sprague-Dawley
(mg/kg-d)
No. litters
male (g)
female (g)
oral -gavage
FO, male and female
0
21
3.92 ±0.58
3.77 ±0.5
(25/sex/group): 0, 250, 500,
250
19
4.03 ±0.32
3.82 ±0.33
1,000 mg/kg-d
dosed daily for 18 wks from
500
20
3.94 ±0.35
3.75 ±0.32
10 wks premating until weaning
1,000
22
3.91 ±0.33
3.66 ±0.39
of F1 pups
Fl, male and female (24—
25/group): 0, 250, 500, 1,000
mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Asano et al. (2011); JPEC (2008i)
Dose
Mean fetal weight ± SD Mean fetal weight ± SD
rabbit, New Zealand White
(mg/kg-d)
No. litters
male (g)
female (g)
oral - gavage
female (24/group): 0,100, 300,
0
22
33.5 ±4.1
31.5 ±3.7
1,000 mg/kg-d
100
22
33.4 ±6.2
31.5 ±4.8
dams dosed daily from GD 6 to
GD 27
300
20
33.9 ±2.5
32.0 ±3.6
C-section GD 28
1,000
23
32.3 ±6.5
30.1 ±6.0
Postnatal Growth
Fuiiietal. (2010); JPEC (2008e)
Dose
No.
Mean±SD Mean±SD
Mean ± SD
rat, Sprague-Dawley
(mg/kg-d)
litters
PND 0(g) PND 4 precull (g)
PND 21(g)
oral - gavage
F0, male and female
Fl-Male Pup Weight
(24/sex/group): 0,100, 300,
0
21
6.9 ±0.7 11.0 ±2.0
61.3 ±6.3
1,000 mg/kg-d; dosed daily for
17 wks, from 10 wks premating
100
22
6.9 ±0.8 11.0 ±1.8
61.0 ±7.0
to lactation day 21
300
23
6.9 ±0.6 10.8 ±1.4
61.6 ±4.6
1,000
22
7.0 ±0.7 10.4 ±1.7
61.6 ±6.4
Fl-Female Pup Weight
0
21
6.5 ±0.7 10.4 ±1.8
59.3 ±6.4
100
22
6.5 ±0.6 10.4 ±1.6
58.5 ±6.4
300
23
6.5 ±0.6 10.2 ±1.4
58.5 ±6.4
1,000
22
6.6 ±0.6 10.0 ±1.8
59.7 ±5.2
This document is a draft for review purposes only and does not constitute Agency policy.
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Reference and study design
Resu Its
Gaoua (2004b)
rat, Sprague-Dawley
oral -gavage
F0, male and female
(25/sex/group): 0, 250, 500,
1,000 mg/kg-d
dosed daily for 18 wks from
10 wks premating until weaning
of F1 pups
Fl, male and female (24—
25/group): 0, 250, 500,
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Dose
(mg/kg-d)
Mean ± SD
PND Kg)
Fl-Male Pup Weight
Mean ± SD
PND 4 precull (g)
Mean ± SD
PND 21(g)
0
6.8 ±0.7
9.1 ± 1.4
50.1 ±4.9
250
6.7 ±0.6
9.0 ± 1.6
51.7 ±4.1
500
6.5 ±0.7
8.7 ± 1.3
50.5 ±6.7
1,000
7.0 ±0.7
9.3 ± 1.2
52.4 ±4.5
Fl-Female Pup Weight
0
6.4 ±0.6
8.6 ± 1.4
48.1 ±6.1
250
6.4 ±0.6
8.5 ± 1.6
49.5 ±4.3
500
6.0 ±0.6
8.1 ± 1.2
48.2 ±5.9
1,000
6.5 ±0.6
8.9 ± 1.2
50.6 ±4.4
F2-Male Pup Weight
0
6.9 ±0.6
9.5 ± 1.5
51.5 ±7.2
250
6.7 ±0.6
9.3 ± 1.0
52.1 ±4.4
500
6.4 ±0.5
9.2 ± 1.0
50.3 ±5.8
1,000
6.3 ±0.6
9.2 ± 1.4
51.2 ±3.6
F2-Female Pup Weight
0
6.5 ±0.6
8.9 ± 1.3
49.6 ±6.2
250
6.3 ±0.6
8.8 ± 1.0
49.9 ±3.6
500
6.4 ±0.5
8.9 ±0.9
49.0 ±5.5
1,000
6.3 ±0.6
8.7 ± 1.4
49.1 ±3.7
Prenatal Morphology
Aso et al. (2014); JPEC (2008h)
rat, Sprague-Dawley
oral - gavage
female (24/group): 0,100, 300,
1,000 mg/kg-d
dams dosed daily from GD 5 to
GD 19
C-section GD 20
No. fetuses
examined for No. fetuses with
No. fetuses
Dose
No. fetuses
visceral
visceral
with visceral
(mg/kg-d)
(litters)d
anomalies
malformations
variations
0
285(21)
146
3(3)
6(6)
100
263(22)
137
2(2)
8(7)
300
251(20)
132
2(2)
4(4)
1,000
270(22)
139
0
8(7)
This document is a draft for review purposes only and does not constitute Agency policy.
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Reference and study design
Resu Its
Aso et al. (2014); JPEC (2008h)
No. fetuses
% fetuses
(continued)
examined
No. fetuses with No. fetuses
(litters) with
Dose
for skeletal
skeletal
with skeletal
skeletal
(mg/kg-d)
anomalies
malformations variations
variations
0
139
0
9(8)
6.5(38.1)
100
126
0
3(3)
2.4(13.6)
300
119
0
3(3)
2.5(15.0)
1,000
131
0
29(13)
22.1(59.1)
No. fetuses (litters) % fetuses (litters)
Dose
with rudimentary with rudimentary
(mg/kg-d)
lumbar rib
lumbar rib
0
4(4)
2.9(19.0)
100
0
0(0)
300
2(2)
1.7(10.0)
1,000
25*(11)
19.1*(50.0)
Gaoua (2004a)
No. fetuses
rat, Sprague-Dawley
No. fetuses with
examined for
No. fetuses with
oral -gavage
Dose
No. fetuses
external
visceral
visceral
female (24/group): 0, 250, 500,
(mg/kg-d)
(litters)d malformations
anomalies
malformations
1,000 mg/kg-d
255(21)
dosed daily from GD 5 to GD 19
0
0
120
0
C-section GD 20
250
226(19)
1(1)
109
0
500
246(20)
0
116
0
1,000
258(22)
0
122
1(1)
No. fetuses
No. fetuses
examined for
No. fetuses with No. fetuses
Dose
with visceral
skeletal
skeletal
with skeletal
(mg/kg-d)
variations
anomalies
malformations
variations
0
1(1)
135
1(1)
125(21)
250
2(2)
117
2(2)
101(19)
500
1(1)
130
1(1)
116(20)
1,000
3(3)
136
2(2)
112(22)
No. fetuses with
% fetuses with
unossified 4th
unossified 4th
metacarpal
% change from control
metacarpal
0
27(9)
-
20.0(42.9)
250
21(10)
-22.2
17.9(52.6)
500
24(9)
-11.1
18.5(45.0)
1,000
43*(12)
37.2
31.6(54.5)
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Reference and study design
Resu Its
Asano et al. (2011); JPEC (2008i)
No.
No. fetuses with No. fetuses with No. fetuses
rabbit, New Zealand White
Dose
fetuses
external visceral with skeletal
oral -gavage
(mg/kg-d)
(litters)d
malformations malformations malformations
female (24/group): 0,100, 300,
0
171(22)
0
1(1) 5(4)
1,000 mg/kg-d
dams dosed daily from GD 6 to
100
174(22)e
1(1)
1(1) 4(4)
GD 27
C-section GD 28
300
167(20)
0
1(1) 3(2)
1,000
159(23)e
1(1)
3(2) 8(5)
Absent right
Dose
No. fetuses with atrioventricular
% change from
(mg/kg-d)
skeletal variations valve
control
0
9(7) 0
-
100
11(9) 0
0(0)
300
6(6) 1(1)
0.6(5.0)
1,000
15(8) 3(2)
1.9(8.7)
Postnatal Morphology
Fuiiietal. (2010); JPEC (2008e)
Male preputial separation -
Female vaginal opening -
rat, Sprague-Dawley
Dose
No.
age (davs)
age (davs)
oral - gavage
(mg/kg-d)
litters
mean ± SD
mean ± SD
FO, male and female
(24/sex/group): 0,100, 300,
rl
1,000 mg/kg-d; dosed daily for
0
21
41.0 ± 1.7
31.2 ± 1.4
17 wks, from 10 wks premating
to lactation day 21
100
22
41.4 ± 1.1
30.9 ± 1.7
300
23
40.6 ± 1.5
30.5 ±2.2
1,000
19
41.2 ± 1.6
30.3 ±2.1
Gaoua (2004b)
Anogenital distance' - males
Anogenital distance'-
rat, Sprague-Dawley
Dose
No.
(PND 1)
females (PND 1)
oral - gavage
(mg/kg-d)
litters
mean ± SD
mean ± SD
F0, male and female
(25/sex/group): 0, 250, 500,
1,000 mg/kg-d
0
21
2.48 ±0.18
1.53 ±0.18
dosed daily for 18 wks from
250
22
2.45 + 0.17
1.5 + 0.14
10 wks premating until weaning
of F1 pups
500
23
2.4 ±0.21
1.45 ±0.14
Fl, male and female (24—
1,000
20
2.43 + 0.15
1.44 + 0.2
25/group): 0, 250, 500,
1,000 mg/kg-d
F2
dosed daily from PND 22 until
0
21
2.41 ±0.18
1.51 ±0.18
weaning of F2 pups
250
22
2.42 ±0.25
1.47 ±0.19
500
23
2.42 ±0.23
1.51 ±0.17
1,000
20
2.45 ±0.21
1.57 ±0.22
This document is a draft for review purposes only and does not constitute Agency policy.
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Reference and study design
Resu Its
Gaoua (2004b)
(continued)
Dose
(mg/kg-d)
No.
litters
Male preputial separation -
age (davs) - mean ± SD
Female vaginal opening -
age (davs) - mean ± SD
F1
0
25
35 ±2
34 ±3
250
25
34 ±2
34 ±3
500
25
35 ±2
35 ±2
1,000
25
35 ±2
33 ±2
aPercent preimplantation loss = (no. preimplantation embryonic loss/no. corpora lutea) xlOO.
bPercent postimplantation loss = (no. resorptions and dead fetuses/no. implantations) xlOO.
cTwo 1,000 mg/kg-d dams were killed in a moribund condition on PND 2 and 4, thus compromising the survival of
their litters. In a third litter, all pups died between PND 1-4 although there was no evidence of maternal toxicity
throughout the study.
dThe parenthetical number following fetal incidence indicates the associated litter incidence for all findings.
eNo. of fetuses examined for visceral and skeletal anomalies at 100 and 1,000 mg/kg-d were 173 and 158,
respectively, because fetuses with external malformations were excluded.
fAGD/cube root of body weight.
*: result is statistically significant (p < 0.05) based on analysis of data by study authors.
-: for controls, no response relevant; for other doses, no quantitative response reported.
% change from control = (control value - treated group value)/control value] x 100.
Absolute change from control (%) = control value (%) - treated group value (%).
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE—Volume 1 of 2
¦ = exposures at which the endpoint was reported statistically significant by study authors
~ = exposures at which the endpoint was reported not statistically significant by study authors
Survival -
Prenatal
Rat; pre implantation loss; GD5-19 (1
Rat; preimplantation loss; GD 5-19 [D
Rabbit; preimplantation loss; GD 6-27 (A
Rat; postlmplantation loss; GD 5-19 (B
Rat; postlmplantation loss; CjD 5-19 (d;
Rabbit,* postlmplantation loss; GDf><27 (A
Survival -
Postnatal
Growth -
Prenatal
Rat; decreased viability, PND4 orPND 21; one-yen repro (C)
Rat; decreased viability, PND 4 urPND2 1; two-gen repro (K)
Rat; decreased mean letal wt.; GD 5-1 *> (B)
Rat; decreased mean fetal wt; GD 5-19 (D)
Rabbit; decreased mean fetal wt.; GD (S-27 (A)
Growth -
Postnatal
Rat; decreased mean pup wt.; onegen repro (C)
Rat; decreased mean pup wt.; two*gen repro (E) -
Morphology
Prenatal
Rat; fetal external or visceral anomalies; GD 5-19 (B)
Rat; fetal external or visceral anomalies; GD 5-19 (D)
Rabbit; fetal external or visceral anomalies; GD 6-27 (A)
Rat; skeletal variation: rudimentary lumbar ribs; GD 5-19 (B)
Rat, skeletal variation, i udimentary lumbar ribs; GD 5-19 (D)
Rabbit; skeletal vaiiation: rudimentary lumbar ribs; GD6*27
(A)
Rat; skeletal variation: unossified 4th metatarsal; GD 5-19 (B) ¦
Rat; skeletal variation; unossified 4th metatarsal; GD 5-19 (D)
Rabbit; skeletal vanation: unossified 4tb metatarsal; G!>f>-27
. (A)
Rat; alteied HI or F2 anugenital distance (I'ND 1); one-^en
Morphology- repro (C)
Postnatal Rat; altered F1 or F2 anogenital distance (PND 1); two-gen
repro (E)
Rat; altered F1 age of puberty (male or female}; one-gen repro
(C)
Rat; altered F1 age of puberty (male or female); two-gen repro
(i)
Rat, .lilt l'.'I HI iotlo\ ontogeny, ai otistic ;>t-n lie ivsponve.
Functional pupil constriction, rootoi activity; one gen i epru (C)
(Neurological) Rat, altered Ft trflex ontogenj, aioustit Maitle i< sponse,
Development pupil constriction, motor activity; two-gen repro (E)
0 B-
O—0—El
CB~
13 B
•e-
-Q
~ ~ ~
-B-
-EP
H—-B-
-O
~ ~ ~
m—b a
Q B-
B-
-a
B B-
-B
~ DO
-B-
p_g_q]
IB B-
Q-—1 ~ ¦
13—b—a
IB B-
-B
Q B-
-Q
B-Q—*
Q B-
HS1
-a
Q " D 03
HB-
-a
~ ~ a
G 0 El
10 100
Dose (mg/kg-day)
1,000
10,000
Sources: (A) Asano et al. 2011; JPEC, 20Q8h (B)Aso et al. 2014; JPEC,2008g(€} Fujii et al., 2010; JPEC, 2008e
(D) Gaoua, 2004a (E) Gaoua, 2004b
Figure 1-15. Exposure-response array of developmental effects following oral
exposure to ETBE.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
Mechanistic Evidence
No mechanistic evidence for developmental effects was identified by the literature search
Integration of Developmental Effects
The evidence to assess developmental toxicity for ETBE consists of two prenatal
developmental toxicity studies in rats and one in rabbits, a one-generation reproductive toxicity
study in rats, and a two-generation reproductive toxicity study in rats. These studies included
assessments of pre- and postnatal survival, growth, morphology, and functional neurological
development following oral (gavage) administration during sensitive periods of development.
Slight evidence of effects ofETBE treatment on prenatal or postnatal survival consisted of
preimplantation loss in a developmental toxicity study in rats and decreased PND 0-4 pup viability
that was associated with severe maternal toxicity. Pre- and postnatal growth (body weights and
developmental landmarks), anogenital distance, sexual maturation, and evaluation of neurological
function (including reflex ontogeny and assessments of acoustic startle response, pupil constriction,
and motor activity in offspring) were not affected by treatment. Evidence of incidental structural
(visceral and skeletal) fetal anomalies following in utero exposures to ETBE were observed at the
highest dose tested (1,000 mg/kg-day). The findings were limited to increased incidences of
rudimentary lumbar rib (Aso etal.. 2014: TPEC. 2008h) and unossified 4th metatarsal (Gaoua.
2004b) in two rat studies and unossified talus and absent right atrioventricular valve in a rabbit
study fAsano etal.. 2011: IPEC. 20080. The fetal, but not litter, incidences of skeletal findings in rats
(rudimentary lumbar rib and unossified 4th metatarsal) were statistically significant at the highest
dose tested (1,000 mg/kg-day). These skeletal observations were not confirmed in other species.
No inhalation prenatal developmental or reproductive toxicity studies were conducted, thus
potential effects of inhalation exposure on pre- and postnatal development have not been
characterized. Overall, the available evidence is considered inadequate to draw conclusions
regarding the development toxicity ofETBE, and developmental effects are not carried forward as a
hazard.
1.2.5. Carcinogenicity (Other than in the Kidney or Liver)
Synthesis of Carcinogenicity Data (Other than in the Kidney or Liver)
This section reviews the studies that investigated whether exposure to ETBE can cause
cancers (other than in the kidney or liver) in humans or animals. The evidence pertaining to
tumorigenicity in the kidney and liver was previously discussed in Sections 1.2.1 and 1.2.2,
respectively. The database for ETBE carcinogenicity consists of only animal data: three 2-year
studies, one 23-week and one 31-week two-stage (i.e., "initiation, promotion") cancer bioassay
performed in rats (Hagiwara et al.. 2013: Saito etal.. 2013: Suzuki etal.. 2012: Hagiwara etal.. 2011:
Malarkev and Bucher. 2011: TPEC. 2010a. b; Maltoni etal.. 1999) (see Table 1-17, Table 1-18; Figure
1-16, Figure 1-17). Interpretation of the study results reported by Maltoni etal. (1999) is
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review ofETBE
complicated by the nonstandard histopathological diagnoses used and the greater than expected
mortality in treated groups and controls compared with other laboratories. Low survival rates at
104 weeks (approximately 25%) in control groups confound these data because whether tumors in
the control group were not observed due to premature death cannot be determined. In response to
these and other concerns, a pathology working group sponsored by EPA and the National
Toxicology Program (NTP) reviewed the histopathological data fMalarkev and Bucher. 20111. In
addition to recalculating tumor incidences, the working group found that the respiratory infections
in the study animals confound interpretation of leukemia and lymphoma. Thus, the Malarkev and
Bucher f20111 data were used when considering carcinogenicity in place of the published Maltoni
etal. (19991 study, and leukemia and lymphoma incidences from this study were not considered.
Following 2-year exposure to ETBE, the incidence of leiomyomas was increased in the
uterus of Sprague-Dawley rats in the high-dose group (Maltoni et al.. 19991. Malignant
schwannomas in the uterus were increased only at the lowest dose, and no significant trend was
observed. These neoplasms arise from nervous tissue and are not specific to uterine tissue.
Leiomyomas and a carcinoma were observed in uterine/vaginal tissue, but no significant trend was
observed fMalarkev and Bucher. 20111. A statistically significant and dose-dependent increase in
incidence of neoplastic lesions was observed in the thyroid of F344 male rats following subchronic
exposure to ETBE after a 4-week tumor initiation exposure to DMBDD (Hagiwara et al.. 2 0111:
incidences of colon and urinary bladder neoplasms also were statistically significantly increased
fHagiwara etal.. 20131. Forestomach papilloma or hyperplasia incidence was elevated statistically
significantly, while no cases were reported in control animals receiving 4 weeks of mutagenic
treatment. This finding is consistent with the rarity of forestomach squamous cell papillomas in
untreated animals (historical control rate = 0.08% in untreated male F344/N rats after 2 years;
(NTP. 20111: comparability with JPEC controls unknown). Exposure to ETBE via gavage in the
absence of prior DMBDD treatment did not significantly induce tumor development in any organs
evaluated (Hagiwara etal.. 20111. Increased tumorigenesis in these tissues was not reported
following 2 years of exposure to ETBE alone via drinking water or inhalation in male or female
F344 rats fSaito etal.. 2013: Suzuki etal.. 2012: TPEC. 2010bl.
Mechanistic Evidence
Available mechanistic evidence was previously discussed in the context of kidney and liver
tumors (Sections 1.1.1 and 1.1.2). Aside from genotoxicity testing results, generally relevant to
tumorigenesis in any tissue location (discussed in the Supplemental Information), no further
mechanistic evidence was identified relevant to uterine, thyroid, colon, forestomach, or urinary
bladder carcinogenesis.
Integration of Carcinogenicity Evidence
The evidence for carcinogenic effects other than liver or kidney is solely from rat studies.
ETBE exposure following mutagen administration increased the incidence of thyroid adenomas or
This document is a draft for review purposes only and does not constitute Agency policy.
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1 carcinomas, colon adenomas or carcinomas, forestomach papillomas, and urinary bladder
2 carcinomas in male rats. Confidence in the data demonstrating an increase in the incidence of
3 schwannomas is reduced due to the lack of a dose-response in Sprague-Dawley rats and lack of a
4 similar effect reported in F344 rats from two other well-conducted 2-year studies, or in F344 or
5 Wistar rats from the two-stage subchronic cancer bioassays. The hazard and dose-response
6 conclusions regarding these carcinomas and adenomas associated with ETBE exposure are further
7 discussed as part of the overall weight of evidence for carcinogenicity in Section 1.3.2.
This document is a draft for review purposes only and does not constitute Agency policy.
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1 Table 1-17. Evidence pertaining to ETBE promotion of mutagen-initiated
2 tumors in animals
Reference and Dosing Protocol
Results by Endpoint
Colon Adenoma or Carcinoma
Hagiwara et al. (2011); JPEC (2008d)
rat, Fischer 344
oral -gavage
male (30/group): 0, 300,1,000 mg/kg-d
daily for 23 wk following a 4-wk tumor initiation by
DMBDDa
+no DMBDD initiation
Dose (mg/kg-d) Response
(incidence)
Male 0 25/30
300 21/30
1,000 28/30*
0+ 0/12
1,000+ 0/12
Forestomach Papillomas or Hyperplasia
Hagiwara et al. (2011); JPEC (2008d)
rat, Fischer 344
oral - gavage
male (30/group): 0, 300,1,000 mg/kg-d
daily for 23 wk following a 4-wk tumor initiation by
DMBDD3
+no DMBDD initiation
Dose (mg/kg-d) Response
(incidence)
Male 0 0/30
300 6/30*
1,000 6/30*
0+ 0/12
1,000+ 0/12
Thyroid Gland Adenoma or Carcinoma
Hagiwara et al. (2011); JPEC (2008d)
rat, Fischer 344
oral - gavage
male (30/group): 0, 300,1,000 mg/kg-d
daily for 23 wk following a 4-wk tumor initiation by
DMBDD3
+no DMBDD initiation
Dose (mg/kg-d) Response
(incidence)
Male 0 8/30
300 17/30*
1,000 20/30*
0+ 0/12
1,000+ 0/12
Urinary Bladder Carcinoma
Hagiwara et al. (2013)
rat, F344/DuCrlCrlj
oral - water
male (30/group): 0,100, 300, 500,1,000 mg/kg-d
daily for 31 wk beginning 1 wk after a 4-wk
exposure to BBN
Dose (mg/kg-d) Response
(incidence)
Male 0 5/30
100 7/30
300 6/30
500 14/30*
1,000 9/26
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Reference and Dosing Protocol
Results by Endpoint
Urinary Bladder Papilloma
Hagiwara et al. (2013)
rat, F344/DuCrlCrlj
oral - water
male (30/group): 0,100, 300, 500,1,000 mg/kg-d
daily for 31 wk beginning 1 wk after a 4-wk
exposure to N-butyl-N-(4-hydroxybutyl) (BBN)
Response
Dose (mg/kg-d) (incidence)
Male 0 21/30
100 13/30
300 17/30
500 17/30
1,000 21/26
Urinary Bladder Papilloma or Carcinoma
Hagiwara et al. (2013)
rat, F344/DuCrlCrlj
oral - water
male (30/group): 0,100, 300, 500,1,000 mg/kg-d
daily for 31 wk beginning 1 wk after a 4-wk
exposure to N-butyl-N-(4-hydroxybutyl) (BBN)
Response
Dose (mg/kg-d) (incidence)
Male 0 24/30
100 18/30
300 20/30
500 25/30
1,000 21/26
Urinary Bladder Papillomatosis
Hagiwara et al. (2011); JPEC (2008d)
rat, Fischer 344
oral -gavage
male (12/group): 0,1,000 mg/kg-d
daily for 23 wk following a 4-wk tumor initiation by
DMBDDa
+no DMBDD initiation
Response
Dose (mg/kg-d) (incidence)
Male 0 0/30
300 0/30
1,000 10/30*
0+ 0/12
1,000+ 2/12
1 aDiethylnitrosamine (DEN), N-butyl-N-(4-hydroxybutyl)nitrosamine (BBN), N-methyl-N-nitrosourea (MNU), 1,2-
2 dimethylhydrazine dihydrochloride (DMH), and N-bis(2-hydroxypropyl)nitrosamine (DHPN).
3
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Table 1-18. Evidence pertaining to carcinogenic effects (in tissues other than
liver or kidney) in animals exposed to ETBE
Reference and study design
Resu Its
Thyroid adenom as/aden oca rein om as
JPEC (2010a); Suzuki et al.
(2012)
rat, Fischer 344
oral - water
male (50/group): 0, 625, 2,500,
10,000 ppm (0, 28, 121,
542 mg/kg-d)a; female
(50/group): 0, 625, 2,500,
10,000 ppm (0, 46, 171,
560 mg/kg-d)a
daily for 104 wk
Incidence
Male
Female
Thvroid
Dose
Thvroid
Dose
follicular
Thvroid
Thvroid follicular
follicular
adenocarcino
follicular
(mg/kg-d)
adenocarcinoma
adenoma
(mg/kg-d)
ma
adenoma
0
0/50
1/50
0
0/50
0/50
28
1/50
0/50
46
1/50
0/50
121
0/50
0/50
171
0/50
0/50
542
0/50
0/50
560
0/50
0/50
JPEC (2010b);Saito et al. (2013)
rat, Fischer 344
inhalation - vapor
male (50/group): 0, 500,1,500,
5,000 ppm (0, 2,090, 6,270,
20,900 mg/m3)b; female
(50/group): 0, 500,1,500,
5,000 ppm (0, 2090, 6270,
20,900 mg/m3)b
dynamic whole body
inhalation; 6 hr/d, 5 d/wk for
104 wk; generation method,
analytical concentration, and
method reported
Incidence
Male
Female
Thvroid
Thvroid
follicular
Thvroid
Dose
Thvroid follicular
follicular
Dose
adenocarcino
follicular
(mg/m3)
adenocarcinoma
adenoma
(mg/m3)
ma
adenoma
0
0/50
1/50
0
1/50
0/50
2,090
0/50
0/50
2,090
1/50
0/50
6,270
0/50
1/50
6,270
1/50
0/50
20,900
0/50
2/50
20,900
0/50
0/50
Maltoni et al. (1999)
rat, Sprague-Dawley
oral -gavage
male (60/group): 0, 250,
1,000 mg/kg-d; female
(60/group): 0, 250,
1,000 mg/kg-d
4 d/wk for 104 wk; observed
until natural death
NOTE: Tumor data not
reanalyzed by Malarkev and
Bucher (2011).
Incidence
Male
Female
Dose
Dose
(mg/kg-d)
Thvroid adenocarcinoma
(mg/kg-d)
Thvroid adenocarcinoma
0
0/60
0
0/60
250
0/60
250
0/60
1,000
0/60
1,000
1/60
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Reference and study design
Resu Its
Endometrial/Uterine carcinogenic effects
JPEC (2010a);Suzuki et al.
Incidence
(2012)
Female
rat, Fischer 344
Dose
Endometrial
Uterine
Uterine
female (50/group): 0, 625,
(mg/kg-d)
stromal sarcoma
adenocarcinoma
fibroma
2,500, 10,000 ppm (0, 46, 171,
0
6/50
1/50
1/50
560 mg/kg-d)a
46
9/50
0/50
0/50
daily for 104 wk
171
3/50
2/50
0/50
560
7/50
2/50
0/50
JPEC (2010b);Saito et al. (2013)
Incidence
rat, Fischer 344
Female
inhalation - vapor
Dose
Endometrial
Uterine
female (50/group): 0, 500,
(mg/m3)
stromal sarcoma
adenocarcinoma
1,500, 5,000 ppm (0, 2,090,
0
2/50
2/50
6,270, 20,900 mg/m3)b
2/50
3/50
2,090
dynamic whole body
inhalation; 6 hr/d, 5 d/wk for
6,270
3/50
1/50
104 wk; generation method,
20,900
2/50
4/50
analytical concentration, and
method reported
Malarkev and Bucher (2011);
Incidence
Maltoni et al. (1999)
Female
rat, Sprague-Dawley
Carcinoma of
Schwannoma
oral -gavage
Dose
the uterus/ Uterine Uterine
of the
Uterine
female (60/group): 0, 250,
(mg/kg-d)
vagina leiomvoma leiomvosarcoma
uterus/vagina
carcinoma
1,000 mg/kg-d
reanalvsis of data from Maltoni
0
0/60
0/60
1/60
0/60
0/60
et al. (1999) for which animals
250
1/60
0/60
0/60
7/60
1/60
were dosed 4 d/wk for 104 wk
1,000
0/60
3/60
0/60
2/60
0/60
1
2 Conversion performed by study authors.
3 b4.18 mg/m3 = 1 ppm.
4 ^Statistically significant (p < 0.05) based on analysis of data conducted by study authors.
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¦ = exposures at which the endpoint was reported statistically significant by study authors
~ = exposures at which the endpoint was reported not statistically significant by study authors
Male rat oral cavity;104wks (Q
Female rat uterine malignancies; 104wks (C)
Male rat all tissues; 104wks (D)
Female rat all tissues; 104wks (D)
Male rat all tissuc?s;23wks without DMBDD initiation (A)
Male rat co!on;23wks following 4wk initiation with DMBDD
(A)
Male rat forestomach or hypcrplasia;23wks following 4wk
initiation with DMBDD (A)
Male rat thyroid;23wks following 4wk initiation with
DMBDD (A)
Male rat urinary bladder carcinoma;31wks following 4wk
initiation with BBN (B)
Male rat urinary bladder papllloma;31wks following 4wk
initiation with BBN (B)
Male rat urinary bladder papillamatosis;23wks following
4wk initiation with DMBDD (A)
Q——0
O : B H
~
ta b—¦—a
B B—B—a
o-
10 100 1,000
Dose (mg/kg-day)
10,000
Sources: (A) Hagiwara et al, 2011; JPEC 2008d (B) Hagiwara et al, 2013 [C) Malarkey and Bucher, 2011
(reanalysis of Maltoni et al, 1999) Maltoni et al, 1999; (D) Suzuki et al., 2012; JPEC, 2010a
Figure 1-16. Exposure-response array of carcinogenic effects following oral
exposure to ETBE.
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¦ = exposures at which the endpoint was reported statistically significant by study authors
~ =exposures at which the endpoint was reported not statistically significant by study authors
Female rat; thyroid adenoma/adenocarcinoma;
104wks (A)
Male rat; thyroid adenoma/adenocarcinoma;
104wks (A)
Female rat; uterine malignancies; 104wks (A)
100 1,000 10,000 100,000
Exposure Concentration (mg/m ')
Source: (A) Saito et al, 2013; JPEC, 2010b
Figure 1-17. Exposure-response array of carcinogenic effects following
inhalation exposure to ETBE.
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1.2.6. Other Toxicological Effects
The database for other effects includes 11 rodent studies, some of which reported
decreased body weight, increased adrenal weights, altered spleen weights, and increased mortality.
All selected studies used inhalation, oral gavage, or drinking water exposure for 90 days or more.
Shorter-duration, multiple-exposure studies that examined immunological endpoints also were
included. The design, conduct, and reporting of each study were reviewed, and each study was
considered adequate.
At this time, the available evidence is considered inadequate to draw conclusions regarding
these other toxic effects following ETBE exposure. For more information, see Appendix B.3.
1.3. INTEGRATION AND EVALUATION
1.3.1. Effects Other Than Cancer
Kidney effects were identified as a potential human hazard ofETBE exposure based on
several endpoints in male and female rats, including kidney weight increases, urothelial
hyperplasia, and—to a lesser extent—exacerbated CPN, and increases in serum markers of kidney
function such as cholesterol, BUN, and creatinine. These effects are similar to the kidney effects
observed with tert-butanol exposure (e.g., CPN and transitional epithelial hyperplasia) and MTBE
(e.g., CPN and mineralization) fATSDR. 19961. Changes in kidney parameters were consistently
observed but the magnitude of change was generally moderate, while males had greater severity of
effects compared to females. MOA analysis determined data are insufficient to conclude that the
a2u-globulin-process operates in male rats. The endpoints associated with a2U-globulin nephropathy
such as linear mineralization, however, were not considered for dose-response analysis because
these endpoints have an unknown relevance to humans. On the other hand, endpoints considered
part of CPN were considered for dose-response analysis since the individual lesions associated with
CPN also occur in the human kidney and exacerbation of one or more of these lesions may reflect a
type of injury relevant to the human kidney. Urothelial hyperplasia was induced in male rats after
2-year inhalation or oral exposure fSaito etal.. 2013: Suzuki etal.. 2012: TPEC. 2010a. b) and was
not confounded by age, as indicated by a complete absence of the lesion in study controls.
Additionally, the robust dose-response relationship and weak correlation with CPN suggest that
urothelial hyperplasia is an effect primarily related to ETBE treatment Urothelial hyperplasia in
male rats, increased severity of CPN, increased blood biomarkers in male and female rats, and
increased kidney weights in male and female rats are considered the result ofETBE exposure,
independent a2U-globulin, and relevant for assessing human health hazard. These effects, therefore,
are suitable for consideration for dose-response analysis and derivation of reference values, as
discussed in Section 2.
Evidence is suggestive that liver effects are associated with ETBE exposure. Increased liver
weight in male and female rats was consistently observed across studies. Centrilobular
hypertrophy was observed at the same concentrations that induced liver weight changes in rats of
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both sexes after 13-week inhalation and 26-week oral exposures. Hypertrophy, however, was not
observed in any 2-year study rat study, suggesting a transient effect. No other histopathological
findings were observed, and only one serum marker of liver toxicity (GGT) was elevated, although
other markers (AST, ALT, and ALP) were not The magnitude of change for these noncancer liver
effects was considered modest and, except for organ weight data, did not exhibit consistent dose-
response relationships. Mechanistic data suggest ETBE exposure leads to activation of several
nuclear receptors, but evidence that nuclear receptor-mediated pathways contribute to the
tumorigenesis observed in ETBE-treated males is inadequate, thus these data remain relevant for
human noncancer hazard identification. Due to the uncertainty that the liver weight increases were
indicative of a liver hazard, no liver effects were considered further for dose-response analysis and
the derivation of reference values.
At this time, there is inadequate information to draw conclusions regarding male
reproductive effects, female reproductive effects, developmental effects, or other toxic effects as
human hazards of ETBE exposure.
1.3.2. Carcinogenicity
Summary of Evidence
In F344 rats, administration ofETBE via inhalation increased the incidence of
hepatocellular adenomas or carcinomas (only one carcinoma observed) at the highest dose tested
in males; hepatocellular tumors were not induced in females fSaito etal.. 20131. Following gavage
or drinking water exposure, liver tumors were not increased in Sprague-Dawley or F344 rats of
either sex fSuzuki etal.. 2012: Maltoni etal.. 19991. Toxicokinetic analysis comparing oral and
inhalation exposures from these three studies using metabolized dose ofETBE or metabolized dose
of tert- butanol (one of the two primary breakdown products ofETBE) demonstrated that these two
routes of exposure yielded comparable internal concentrations (see Supplemental Information,
Appendix B.1.5.4). This observation suggests thatthe lack of carcinogenic effects via oral exposure
is likely not due to a difference in administered dose. Therefore, the observed lack of a tumor
response following oral exposure suggests that ETBE might not cause significant induction of rat
tumors via the oral route. Statistically significant increases in liver tumor incidence, however, were
observed in the livers of male F344 and Wistar rats in initiation-promotion studies, after 19-23
weeks ofETBE exposure via oral gavage, following an initial 2-4-week mutagen exposure
(Hagiwara etal.. 2015: Hagiwara etal.. 20111. Furthermore, colon, thyroid, forestomach, and
urinary bladder tumorigenesis also was promoted by oral ETBE exposure in male F344 rats
f Hagiwara etal.. 2013: Hagiwara etal.. 20111. Incidence of kidney tumors in rats was not
significantly increased following 2 years of oral or inhalation exposure to ETBE alone, nor did ETBE
promote kidney tumorigenesis in male F344 rats; however, increased renal tubule tumors were
promoted in male Wistar rats following mutagen administration. No studies have evaluated chronic
ETBE exposure in mice via any route.
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The Cancer Guidelines (U.S. EPA. 2005a) emphasize that knowledge of the biochemical and
biological changes preceding tumor development could inform whether a cancer hazard exists and
might help in understanding events relevant to potential mode of carcinogenic action. As discussed
in Section 1.2.2, the evidence for the nuclear hormone receptor MOAs (i.e., PPARa, PXR, or CAR)
was inadequate to determine what role, if any, these pathways play in ETBE-induced liver
carcinogenesis. Centrilobular hypertrophy could be induced through several possible mechanisms,
including nuclear receptor activation, but centrilobular hypertrophy was not associated with
tumorigenesis. The data are also inadequate to show that tert-butanol, an ETBE metabolite formed
in the liver with acetaldehyde (Section 1.1.2), activates nuclear receptors, increases centrilobular
hypertrophy, or induces proliferative liver lesion formation. The observations of proliferation and
apoptosis had little temporal coherence, suggesting that these proposed downstream key events
were not related to nuclear receptor activation. Acetaldehyde-mediated genotoxicity also was
evaluated as a possible MOA. ALDH2 deficiency enhanced ETBE-induced genotoxicity in
hepatocytes and leukocytes from exposed mice; although suggestive, the available data overall are
inadequate to conclude that ETBE induces liver tumors via acetaldehyde-mediated mutagenicity.
An MOA for liver carcinogenesis could not be established, and in the absence of information to
indicate otherwise (U.S. EPA. 2005b). the liver tumors induced by ETBE are relevant to human
hazard identification.
As mentioned in Sections 1.1.2 through 1.1.4, ETBE is primarily metabolized into
acetaldehyde and tert-butanol, a compound also formed by MTBE metabolism; the rodent bioassays
from both MTBE and tert-butanol could provide supplementary information on the carcinogenicity
ofETBE. For MTBE, the most recent cancer evaluation by a national or international health agency
is from IARC f!999cl IARC reported that oral gavage exposure in Sprague-Dawley rats resulted in
testicular tumors in males and lymphomas and leukemias (combined) in females; inhalation
exposure in male and female F344 rats resulted in renal tubule adenomas in males; and inhalation
exposure in male and female CD-I mice resulted in hepatocellular adenomas in females (IARC.
1999c). For tert-butanol, a draft IRIS assessment under development concurrently with this
assessment reports that drinking water exposure in F344 rats resulted in renal tubule tumors,
mostly adenomas, in males; drinking water exposure also increased the incidence of thyroid
follicular cell adenomas in female B6C3Fi mice and adenomas or carcinomas (only one carcinoma
observed) in males.
Integration of evidence
This evidence leads to consideration of two hazard descriptors under EPA's cancer
guidelines fU.S. EPA. 2005al The descriptor likely to be carcinogenic to humans is appropriate when
the evidence is "adequate to demonstrate carcinogenic potential to humans" but does not support
the descriptor carcinogenic to humans. One example from the cancer guidelines is "an agent that has
tested positive in animal experiments in more than one species, sex, strain, site, or exposure route,
with or without evidence of carcinogenicity in humans." The database for ETBE does not appear to
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match the conditions of this example, having increased tumor incidences only in male rats, and only
via inhalation; however, this conclusion is limited by the lack of studies evaluating chronic exposure
by any route in another species (e.g., mice).
Alternatively, the descriptor suggestive evidence of carcinogenic potential is appropriate
when the evidence raises "a concern for potential carcinogenic effects in humans" but is not
sufficient for a stronger conclusion, and covers a spectrum of evidence associated with varying
levels of concern for carcinogenicity. Such evidence can range from a positive cancer result in the
only study on an agent to a single positive cancer result in an extensive database that includes
negative studies in other species. The results for ETBE raise a concern for cancer, but the effects
were limited primarily to one tissue (liver), at one dose (highest), and in one sex/species
combination (male rats), which were almost entirely benign. Although MTBE also was associated
with liver tumorigenesis in male and female mice, no data are available for comparison with ETBE,
which has not been evaluated in chronic mouse bioassays. Furthermore, results between ETBE- and
tert-butanol- or MTBE-associated tumorigenesis in rats have little coherence, as ETBE did not
induce renal tubule tumorigenesis.
Knowledge of the biochemical and biological changes preceding tumor development also
might provide important insight for determining whether the cancer descriptor for a particular
agent (and route of exposure) is appropriate (U.S. EPA. 2005a). Although the guidelines do not
provide specific recommendations on how to incorporate results from 2-stage "initiation-
promotion" carcinogenesis studies, these studies are considered along with standard 2-year
bioassays by IARC flARC. 20151. Across three initiation-promotion studies, orally administered
ETBE enhanced tumorigenesis in multiple tissues in male rats pre-exposed to mutagens, including
kidney, liver, forestomach, thyroid, colon, and urinary bladder. Although the ETBE metabolite tert-
butanol similarly induced tumors in two of the tissues (kidney tumors in rats, thyroid tumors in
mice), and ETBE alone caused liver toxicity and tumorigenesis in 2-year rat inhalation bioassays, no
treatment-related toxicity has been reported in the rat forestomach, thyroid, colon, or urinary
bladder following chronic exposure to either ETBE or tert-butanol independently. Furthermore, no
systemic MOA has been identified for ETBE, which could explain the potentiation of mutagen-
induced carcinogenesis in the forestomach, thyroid, colon, and urinary bladder. This suggests that
the available database is severely limited with regard to informing molecular mechanisms of ETBE
carcinogenesis. The available evidence suggests that populations exposed to mutagenic agents prior
to, or concomitant with, oral ETBE exposure might be more susceptible to chemically induced
carcinogenesis than predicted by the results ofETBE 2-year rodent oral bioassays alone.
These considerations, interpreted in light of the cancer guidelines, support the conclusion of
suggestive evidence of carcinogenic potential for ETBE. This finding is based primarily on a positive
carcinogenic response in the liver at one dose in a single animal study, along with significant
increases in focal pre-neoplastic liver lesions and mechanistic data, including the metabolism of
ETBE to acetaldehyde in the liver, and the mutagenic and genotoxic effects of acetaldehyde.
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Although the available guidelines do not provide instruction for incorporating initiation-promotion
bioassay data, this evidence also appears consistent with the descriptor of suggestive evidence of
carcinogenic potential.
The descriptor, suggestive evidence of carcinogenic potential, applies to all routes of human
exposure. Inhalation administration of ETBE to male rats induced tumors beyond the point of initial
contact, as discussed in Section 1.2.2. Although the results from the oral exposure 2-year ETBE
bioassays on rats were negative (mice were not tested), the increased liver tumorigenesis reported
in two strains of male rats following oral ETBE exposure across three two-stage "initiation-
promotion" cancer bioassays, and the enhanced systemic genotoxicity reported in the absence of
ALDH2 in transgenic mice, together provide additional biological plausibility for carcinogenicity
following oral ETBE exposure (see Sections 1.2.2 and 1.2.5). Together with the enhanced
carcinogenicity reported in multiple other male rat tissues following oral exposure in 2-stage
initiation-promotion bioassays, the evidence implicating acetaldehyde in the human carcinogenicity
associated with ethanol consumption coupled with the increased genotoxicity observed in ALDH2-
deficient transgenic mice exposed to ETBE (see Section 1.3.3), this evidence was decisive in
extending the weight of evidence descriptor to the oral route. According to the cancer guidelines
(U.S. EPA. 2005a). this information provides sufficient basis to apply the cancer descriptor
developed from inhalation studies to other exposure routes.
Biological considerations for dose-response analysis
Regarding hazards to bring forward to Section 2 for dose-response analysis, the observed
liver tumors are relevant to human cancer hazard. The results from MOA analysis could inform
dose-response analysis and extrapolation approaches fU.S. EPA. 2005al. As discussed above, the
evidence was inadequate to determine the role of nuclear receptor activation in liver
carcinogenesis, due in part to a lack of coherence between nuclear receptor activation and
proliferation or apoptosis, key events in these pathways. Evidence also was inadequate to conclude
that ETBE induces liver tumors via acetaldehyde-mediated mutagenic MOA, due in part to a paucity
of evidence specifically evaluating intermediate key events following ETBE exposure in rats. No
other systemic cancer MOAs were identified. In the absence of MOA information to indicate
otherwise, dose-response analysis should use linear extrapolation fU.S. EPA. 2005al. The Saito et al.
(2013) inhalation study was considered suitable for dose-response analysis, as it is part of a well-
designed GLP study (OECD Guideline 451) that evaluated multiple dose levels (TPEC. 2010b). The
study included histological examinations for tumors in many different tissues, contained three
exposure levels and controls, contained adequate numbers of animals per dose group
(~50/sex/group), treated animals for up to 2 years, and included detailed reporting of methods
and results.
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1.3.3. Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes
Genetic polymorphisms of ALDH2, the enzyme that oxidizes acetaldehyde to acetic acid,
might affect potential ETBE liver toxicity. The virtually inactive form, ALDH2*2, is responsible for
alcohol intolerance and is found in about one-half of East Asian populations fBrennan. 20021. This
variant is associated with slow metabolism of acetaldehyde and, hence, extended exposure to a
genotoxic compound. Other studies also have linked ALDH2 polymorphisms to hepatocellular
cancers in humans fEriksson. 20151. With respect to ETBE exposure, the ALDH2*2 variant should
increase any type of risk associated with acetaldehyde produced by ETBE metabolism because it
will prolong internal exposure to this metabolite. As demonstrated in several in vivo and in vitro
genotoxic assays in AIdh2 KO mice or cells, genotoxicity was significantly increased compared with
wild-type controls following ETBE exposure to similar doses where both cancer and noncancer
effects were observed following chronic rodent exposure bioassays fWeng etal.. 2014: Wengetal..
2013: Wengetal.. 2012: Wengetal.. 20111. Studies in AIdh2 KO mice observed elevated blood
concentrations of acetaldehyde following ETBE exposure compared with wild-type mice fWeng et
al.. 20131. increased alterations to sperm and male reproductive tissue fWeng etal.. 20141. and
increased incidence of centrilobular hypertrophy fWeng etal.. 2013: Wengetal.. 20121. Notably, a
consistent finding in these studies was increased severity of genotoxicity in males compared with
females, which corresponds with increased incidence of hepatic tumors only in male rats fSaito et
al.. 2013: TPEC. 2010bl. No MOA information exists to account for the sex discrepancies in genotoxic
effects. Finally, IARC f!999al and IARC T20121 identified acetaldehyde produced as a result of
ethanol metabolism as contributing to human carcinogenesis in the upper aerodigestive tract and
esophagus following ethanol ingestion, with effects amplified by slower acetaldehyde metabolism.
Altogether, these data present plausible evidence that diminished ALDH2 activity yields health
effect outcomes that are more severe than those organisms with fully functional ALDH2.
No other specific potential polymorphic-related susceptibility issues were reported in the
literature. CYP2A6 is likely to be the P450 isoenzyme in humans to cleave the ether bond in ETBE. It
also exists in an array of variants, and at least one variant (2A6*4) clearly has no catalytic activity
fFukami et al.. 20041: however, the effect of this variability on ETBE toxicity is unknown. In
addition, the data on ETBE-induced mutagenicity are inconclusive.
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2. DOSE-RESPONSE ANALYSIS
2.1. ORAL REFERENCE DOSE FOR EFFECTS OTHER THAN CANCER
The reference dose (RfD) (expressed in units of mg/kg-day) is defined as an estimate (with
uncertainty spanning perhaps an order of magnitude) of a daily exposure to the human population
(including sensitive subgroups) that is likely to be without an appreciable risk of deleterious effects
during a lifetime. It can be derived from a no-observed-adverse-effect level (NOAEL), lowest-
observed-adverse-effect level (LOAEL), or the 95% lower bound on the benchmark dose (BMDL),
with uncertainty factors (UF values) generally applied to reflect limitations of the data used.
2.1.1. Identification of Studies and Effects for Dose-Response Analysis
Studies were evaluated using general study quality characteristics as discussed in
Section 1.1.1; see also U.S. EPA (2002) to help inform the selection of studies from which to derive
toxicity values.
Human studies are preferred over animal studies when quantitative measures of exposure
are reported and the reported effects are determined to be associated with exposure. No human
occupational or epidemiological studies of oral exposure to ETBE, however, are available.
Animal studies were evaluated to determine which studies provided (1) the most relevant
routes and durations of exposure, (2) multiple exposure levels that informed the shape of the dose-
response curve, and (3) sufficient sample size to detect effects at low exposure levels (U.S. EPA.
20021. The database for ETBE includes several chronic and subchronic studies, mostly in rats,
showing effects in the kidney that are suitable for use in deriving oral reference values. In general,
lifetime exposures are preferred over subchronic exposures.
Kidney Toxicity
Kidney effects were identified as a potential human hazard of ETBE-induced toxicity based
on findings in male and female rats (summarized in Section 1.3.1). Kidney toxicity was observed
across several chronic and subchronic studies following oral and inhalation exposure, based on
findings of organ weight changes, histopathology (urothelial hyperplasia), and altered serum
biomarkers (cholesterol, creatinine, BUN) in rats. The strongest and most consistent findings across
oral exposure routes and durations were for absolute kidney weight changes and urothelial
hyperplasia; thus, only these endpoints were analyzed for dose-response. Kidney effects observed
after chronic exposure, such as urothelial hyperplasia, could affect the ability of the kidney to filter
waste, and changes in kidney weight could serve as a general indication of renal toxicity. In the case
of kidney weight changes, numerous chronic and subchronic studies investigated this endpoint
following oral and inhalation exposure fMivata etal.. 2013: Saito etal.. 2013: Suzuki etal.. 2012:
Hagiwaraetal.. 2011: Fuiii etal.. 2010: TPEC. 2010b. 2008b. c; Gaoua. 2004b: Medinskvetal.. 19991.
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Chronic studies of oral exposure reported urothelial hyperplasia to be increased with treatment in
male rats (Saito etal.. 2013: Suzuki etal.. 2012: TPEC. 2010a. b).
Hagiwara etal. (20111. with only one dose group, was not considered further given its
concordance with several other rat studies that had multiple groups. Additionally, as discussed in
Section 1.1.1, 2-year organ weight data were not considered suitable due to the prevalence of age-
associated confounders. Therefore, the urothelial hyperplasia data were the only endpoint from the
2-year studies flPEC. 2010al selected data published as Suzuki etal. f20121. and absolute kidney
weight was the only endpoint from the 13- to 26-week studies that were considered for dose-
response analysis. These data and the absolute kidney weights from the remaining studies, TPEC
(2008c) selected data published as Mivataetal. (20131. Gaoua (2004b). Fuiii etal. (20101. are
discussed further below.
In the 2-year drinking water study f Suzuki etal.. 2012: TPEC. 2010al. male and female F344
rats (50/sex/dose group) were exposed to doses of 0, 28,121, or 542 mg/kg-day. Increased
incidence of urothelial hyperplasia was observed only in males and significantly increased at 121
and 542 mg/kg-day. Effects were not observed in similarly exposed females, thus female
hyperplasia was not modeled.
In the TPEC (2008c) 26-week gavage study, male and female Crl:CD(SD) rats (15/sex/dose
group) were exposed to daily doses of 0, 5, 25,100, or 400 mg/kg-day. Absolute kidney weight was
significantly increased in males and females treated with 400 mg/kg-day. Abnormal
histopathological findings in the kidney (basophilic tubules and hyaline droplets) were observed in
male rats, but not in female rats.
In the Gaoua (2004b) two-generation reproductive toxicity study, Sprague-Dawley rats
(25/sex/dose group) were exposed via gavage to doses of 0, 250, 500, or 1,000 mg/kg-day;
treatment commenced 10 weeks before mating and continued throughout the 2-week mating
period, gestation, and the end of lactation (PND 21) for 18 weeks. Absolute kidney weights were
significantly increased in all dose groups in P0 males, but not in P0 females, which was associated
with the presence of acidophilic globules in renal tissue from 5/6 males examined. In addition,
tubular basophilia (4/6), peritubular fibrosis (3/6), and proteinaceous casts (1/6) were observed
in kidneys of male rats at the high dose. Similar microscopic effects in females were not observed,
thus P0 female kidney weights were not modeled. Absolute kidney weights were increased in F1
males at 500 and 1,000 mg/kg-day and females at 1,000 mg/kg-day.
In the Fuiii etal. f20101 one-generation reproductive toxicity study, male and female
Crl:CD(SD) rats (24/sex/dose group) were exposed via gavage to doses of 0,100, 300, or
1,000 mg/kg-day beginning 10 weeks prior to F0 mating and continuing throughout the
reproductive period (mating, gestation, lactation). Treatment durations were stated to be
approximately 16 weeks for males and 17 weeks for females but ranged up to 20 weeks in animals
that took longer to mate. Kidney weights were significantly increased in F0 males and females at
1,000 mg/kg-day.
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2.1.2. Methods of Analysis
No biologically based dose-response models are available for ETBE. In this situation, a range
of dose-response models was evaluated to determine how best to model the dose-response
relationship empirically in the range of the observed data. The models in EPA's Benchmark Dose
Software (BMDS) were applied. Consistent with EPA's Benchmark Dose Technical Guidance
Document fU.S. EPA. 20121. the BMD and the BMDL are estimated using a benchmark response
(BMR) to represent a minimal, biologically significant level of change. In the absence of information
regarding what level of change is considered biologically significant, a BMR of 10% change from the
control mean (relative deviation; RD) for kidney weight and urothelial hyperplasia data is used to
estimate the BMD and BMDL and to facilitate a consistent basis of comparison across endpoints,
studies, and assessments. When modeling was feasible, the estimated BMDLs were used as points of
departure (PODs); the PODs are summarized in Table 2-1. Details, including the modeling output
and graphical results for the model selected for each endpoint are presented in Appendix C of the
Supplemental Information to this Toxicological Review.
Human equivalent doses (HEDs) for oral exposures were derived from the PODs according
to the hierarchy of approaches outlined in EPA's Recommended Use of Body Weight3/4 as the Default
Method in Derivation of the Oral Reference Dose fU.S. EPA. 20111. The preferred approach is
physiologically based pharmacokinetic (PBPK) modeling. Other approaches include using chemical-
specific information in the absence of a complete PBPK model. As discussed in Appendix B of the
Supplemental Information, several rat PBPK models for ETBE have been developed and published,
but a validated human PBPK model for ETBE for extrapolating doses from animals to humans is not
available. In lieu of chemical-specific models or data to inform the derivation of human equivalent
oral exposures, body-weight scaling to the % power (BW3/4) is applied to extrapolate toxicologically
equivalent doses of orally administered agents from adult laboratory animals to adult humans to
derive an oral RfD. BW3/4 scaling was not used for deriving HEDs from studies in which doses were
administered directly to early postnatal animals because of the absence of information on whether
allometric (i.e., body weight) scaling holds when extrapolating doses from neonatal animals to adult
humans due to presumed toxicokinetic or toxicodynamic differences between lifestages fU.S. EPA.
2011: Hattis etal.. 20041.
Consistent with EPA guidance fU.S. EPA. 20111. the PODs estimated based on effects in adult
animals are converted to HEDs using a standard dosimetric adjustment factor (DAF) derived as
follows:
DAF = (BWa1/4 / BWh1/4)
where:
BWa = animal body weight
BWh = human body weight
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1 Using a standard BWa of 0.25 kg for rats and a BWh of 70 kg for humans fU.S. EPA. 19881.
2 the resulting DAF for rats is 0.24. Applying the DAF to the POD identified for effects in adult rats
3 yields a PODhed as follows (see Table 2-1):
4
5 PODhed = Laboratory animal dose (mg/kg-day) x DAF
6
7 Table 2-1 summarizes the sequence of calculations leading to the derivation of a human-
8 equivalent POD for each data set discussed above.
9 Table 2-1. Summary of derivation of points of departure following oral
10 exposure for up to 2 years
Endpoint and Reference
Species/
Sex
Model3
BMR
BMD
(mg/kg-d)
BMDL
(mg/kg-d)
PODadj"
(mg/kg-d)
PODhed0
(mg/kg-d)
Kidney
Increased urothelial
hyperplasia; 2-year
Suzuki etal. (2012); JPEC
(2010a)
Male Fischer
rats
Quanta 1-
Linear
10%
ER
79.3
60.5
60.5
14.5
Increased absolute kidney
weight; 26-week
JPEC (2008c); Mivata et al.
(2013)
Male
Sprague-
Dawley rats
Linear
10%
RD
176
115
115
27.6
Increased absolute kidney
weight; 26-week
JPEC (2008c); Mivata et al.
(2013)
Female
Sprague-
Dawley rats
Exponential
(M4)
10%
RD
224
57
57
13.7
Increased absolute kidney
weight (P0 generation);
18-week
Gaoua (2004b)
Male
Sprague-
Dawley rats
Hill
10%
RD
244
94
94
22.6
Increased absolute kidney
weight (F1 generation); in utero
through lactation and breeding
Gaoua (2004b)
Male
Sprague-
Dawley rats
Polynomial
3°
10%
RD
318
235
235
235
Increased absolute kidney
weight (F1 generation); in utero
through lactation and breeding
Gaoua (2004b)
Female
Sprague-
Dawley rats
Exponential
(M2)
10%
RD
978
670
670
670
Increased absolute kidney
weight (P0 generation); 16-
week
Fuiii et al. (2010)
Male
Sprague-
Dawley rats
Hill
10%
RD
435
139
139
33.4
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Endpoint and Reference
Species/
Sex
Model3
BMR
BMD
(mg/kg-d)
BMDL
(mg/kg-d)
PODadj"
(mg/kg-d)
PODhed0
(mg/kg-d)
Increased absolute kidney
weight (PO generation); 17-
week
Fuiii et al. (2010)
Female
Sprague-
Dawley rats
Polynomial
2°
10%
RD
1,094
905
905
217
1 aFor modeling details, see Appendix C of the Supplemental Information.
2 bFor studies in which animals were not dosed daily, administered doses were adjusted to calculate the TWA daily
3 doses prior to BMD modeling. This adjustment, however, was not required for the studies evaluated.
4 CHED PODs were calculated using BW3/4scaling (U.S. EPA, 2011).
5 ER = extra risk, RD = relative deviation.
6 2.1.3. Derivation of Candidate Values
7 Consistent with EPA's A Review of the Reference Dose and Reference Concentration Processes
8 (U.S. EPA. 2002: Section 4.4.51. five possible areas of uncertainty and variability were considered
9 when determining the application of UF values to the PODs presented in Table 2-1. An explanation
10 follows.
11 An intraspecies uncertainty factor, UFh, of 10 was applied to all PODs to account for
12 potential differences in toxicokinetics and toxicodynamics in the absence of information on the
13 variability of response in the human population following oral exposure to ETBE fU.S. EPA. 20021.
14 An interspecies uncertainty factor, UFa, of 3 (10°5 = 3.16, rounded to 3) was applied to PODs
15 that used BW3/4 scaling to extrapolate oral doses from laboratory animals to humans. Although
16 BW3/4 scaling addresses some aspects of cross-species extrapolation of toxicokinetic and
17 toxicodynamic processes, some residual uncertainty remains. In the absence of chemical-specific
18 data to quantify this uncertainty, EPA's BW3/4 guidance fU.S. EPA. 20111 recommends using an
19 uncertainty factor of 3. For PODs that did not use BW3/4 such as early-life effects, an interspecies
20 uncertainty factor, UFa, of 10 was applied fU.S. EPA. 20111.
21 A subchronic-to-chronic uncertainty factor, UFs, differs depending on the exposure
22 duration. For studies of 16- to 26-week duration, the magnitude of change observed in kidney
23 weights was similar to the effect observed at 104 weeks. This suggests a maximum effect could
24 have been reached by 16-26 weeks. The 104-week kidney data, however, are confounded due to
25 age-associated factors, so this comparison might not be completely reliable. Additionally, some but
26 not all markers of kidney toxicity appear more severely affected by ETBE at 2 years compared with
27 observations at 16-26 weeks (e.g., histopathology, BUN) fSuzuki etal.. 2012: TPEC. 2010al. Thus, a
28 UFs of 3 was applied for studies of 16- to 26-week duration to account for this uncertainty, and a
29 UFs of 1 was applied to 2-year studies.
30 A LOAEL-to-NOAEL uncertainty factor, UFl, of 1 was applied to all PODs derived because the
31 current approach is to address this factor as one of the considerations in selecting a BMR for
32 benchmark dose modeling. In this case, BMRs of a 10% change in absolute kidney weight and a
33 10% extra risk of urothelial hyperplasia were selected assuming that they represent minimal
34 biologically significant response levels.
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A database uncertainty factor, UFd, of 1 was applied to all PODs. The ETBE oral toxicity data
set includes a 2-year toxicity study in rats (Suzuki etal.. 2012: TPEC. 2010al. a 26-week toxicity
study in rats (Mivata etal.. 20131. prenatal developmental toxicity studies in rats and rabbits (Aso
etal.. 2014: Asano etal.. 20111. and both single- and multigene ration reproductive studies and
developmental studies in rats fFuiii etal.. 2010: Gaoua. 2004a. b). The ETBE data set does not
indicate immunotoxicity fBanton etal.. 2011: Li etal.. 20111. Additionally, the available mouse
study observed less severe effects than those in rats, suggesting that mice are less sensitive than
rats. Although most of the studies are in rats, the ETBE oral database adequately covers all major
systemic effects, including reproductive and developmental effects, and does not suggest that
additional studies would lead to identification of a more sensitive endpoint or a lower POD.
Furthermore, the effects observed in inhalation studies support the effects observed in the oral
studies. Therefore, an uncertainty factor for the database, UFd, of 1 was applied.
Table 2-2 is a continuation of Table 2-1 and summarizes the application of UF values to each
POD to derive a candidate value for each data set, preliminary to the derivation of the
organ/system-specific RfDs. These candidate values are considered individually in selecting a
representative oral reference value for a specific hazard and subsequent overall RfD for ETBE.
Figure 2-1 graphically presents the candidate values, UFs, and PODhed values, with each bar
corresponding to one data set described in Table 2-1 and Table 2-2.
Table 2-2. Effects and corresponding derivation of candidate values
Endpoint and Reference
PODhed
(mg/kg-d)
POD
type
UFa
UFh
UFl
UFs
UFd
Composite
UF
Candidate
value
(mg/kg-d)
Kidney
Increased urothelial hyperplasia;
male rat; 2-year
Suzuki et al. (2012): JPEC (2010a)
14.5
BMDLio
3
10
1
1
1
30
5 x 10 1
Increased absolute kidney weight;
male rat; 26-week
JPEC (2008c): Mivata et al. (2013)
27.6
BMDLio%
3
10
1
3
1
100
3 x 10 1
Increased absolute kidney weight;
female rat; 26-week
JPEC (2008c): Mivata et al. (2013)
13.7
BMDLio%
3
10
1
3
1
100
1 x 10 1
Increased absolute kidney weight;
P0 male rat; 18-week
Gaoua(2004b)
22.6
BMDLio%
3
10
1
3
1
100
2 x 10 1
Increased absolute kidney weight;
F1 male rat; in utero through
lactation and breeding
Gaoua(2004b)
235
BMDLio%
10
10
1
3
1
300
8 x 10 1
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Endpoint and Reference
PODhed
(mg/kg-d)
POD
type
UFa
UFh
UFl
UFS
UFd
Composite
UF
Candidate
value
(mg/kg-d)
Increased absolute kidney weight;
F1 female rat; in utero through
lactation and breeding
Gaoua(2004b)
670
BMDLio%
10
10
1
3
1
300
2x 10°
Increased absolute kidney weight;
male rat; 16-week
Fuiii et al. (2010)
33.4
BMDLio%
3
10
1
3
1
100
3 x 10 1
Increased absolute kidney weight;
female rat; 17-week
Fuiii et al. (2010)
217
BMDLio%
3
10
1
3
1
100
2x 10°
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Increased urothelial hyperplasia; male
rat; 2 year
Suzuki et ai. (2012); J PEC (2010a)
Increased absolute kidney weight; male
rat; 26 week
J PEC (2008c)
Increased ab\ulul<> kidney weight;
fero.ik r H; 26 week
H'FC (2008c)
Increased absolute kidney weight; male
rat; IfUveek
Gaoua (UtXMb)
Increased absolute kidney weight; F1
raaie rat; in utero through lactation and
breeding
Gaoua (2004b)
Increased absolute kidney weight; F1
female rat; in utero thi-uugh lactation
ami breeding
(.i.wiu (200'Ib)
Increased absolute kidney weight; male
rat; 18 week
Fuji! (2010)
Increased .ihsnhik- kidney weight;
ferrule rat; 17 week
FuttiUOlO)
mg/itg-day
1000
^ Candidate RfD
• P0Dhed
:: Composite DF
Figure 2-1. Candidate values with corresponding POD and composite UF. Each
bar corresponds to one data setdescribed in Table 2-1 and Table 2-2.
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2.1.4. Derivation of Organ/System-Specific Reference Doses
Table 2-3 distills the candidate values from Table 2-2 into a single value for each organ or
system. Organ- or system-specific RfDs are useful for subsequent cumulative risk assessments that
consider the combined effect of multiple agents acting at a common site.
Kidney Toxicity
For ETBE, candidate values were derived for increases in urothelial hyperplasia or absolute
kidney weight in male or female rats, spanning a range from 1 x 101 to 2 x 10° mg/kg-day, for an
overall 20-fold range. Selection of a point estimate considered multiple aspects, including study
design and consistency across estimates. As stated previously, reference values based on lifetime
exposure are preferred over subchronic exposures. The only candidate reference value based on
data from a 2-year oral study is that for urothelial hyperplasia in male rats (Saito etal.. 2013:
Suzuki etal.. 2012: TPEC. 2010a. b). Consistent with the above, the composite UF for urothelial
hyperplasia was the lowest of all the candidate values, which provides greater confidence in the
selection of the candidate. This lesion is a specific indicator of kidney toxicity and is synonymous
with the transitional epithelial hyperplasia in the renal pelvis observed after chronic tert-butanol
exposure in both male and female rats fNTP. 1995a], Furthermore, the Toxicological Review of tert-
butanol identified transitional epithelial hyperplasia in the kidney as the highest POD lending
support that this endpoint is a specific indicator of kidney toxicity following ETBE exposure. On the
other hand, kidney weight changes represent a nonspecific effect, and the data available on kidney
weight changes have greater composite UF values than the hyperplasia value, in part because they
are derived from studies of 16- to 2 6-week duration, which are shorter than lifetime exposures.
Collectively, these observations suggest that the most appropriate basis for a kidney-
specific RfD would be the increased incidence of urothelial hyperplasia in male rats from the 2-year
oral study (Suzuki etal.. 2012: TPEC. 2010a). To estimate an exposure level below which kidney
toxicity from ETBE exposure is not expected to occur, the candidate value for increased incidence of
urothelial hyperplasia in male rats (5 x 101 mg/kg-day) was selected as the kidney-specific
reference dose for ETBE. Confidence in this RfD is high. The POD is based on benchmark dose
modeling, and the candidate value is derived from a well-conducted GLP study, involving a
sufficient number of animals per group, assessing a wide range of kidney endpoints.
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Table 2-3. Organ/system-specific RfDs and overall RfD for ETBE
Effect
Basis
RfD
(mg/kg-day)
Study exposure
description
Confidence
Kidney
Incidence of urothelial
hyperplasia Suzuki et al.
(2012); JPEC (2010a)
5 x 10 1
Chronic
High
Overall RfD
Kidney
5 x 101
Chronic
High
2.1.5. Selection of the Overall Reference Dose
For ETBE, only kidney effects were identified as a hazard and carried forward for dose-
response analysis; thus, only one organ/system-specific reference dose was derived. Therefore, the
kidney-specific RfD of 5 x 10"1 mg/kg-day is the overall RfD for ETBE. This value is based on
increased incidence of urothelial hyperplasia in male rats exposed to ETBE.
The overall reference dose is derived to be protective of all types of effects for a given
duration of exposure and is intended to protect the population as a whole, including potentially
susceptible subgroups fU.S. EPA. 20021. Decisions concerning averaging exposures over time for
comparison with the RfD should consider the types of toxicological effects and specific lifestages of
concern. Fluctuations in exposure levels that result in elevated exposures during these lifestages
could lead to an appreciable risk, even if average levels over the full exposure duration were less
than or equal to the RfD. In the case of ETBE, no specific potential for early lifestage susceptibility to
ETBE exposure was identified, as discussed in Section 1.3.3.
2.1.6. Confidence Statement
A confidence level of high, medium, or low is assigned to the study used to derive the RfD,
the overall database, and the RfD, as described in Section 4.3.9.2 of EPA's Methods for Derivation of
Inhalation Reference Concentrations and Application of Inhalation Dosimetry fU.S. EPA. 19941. The
overall confidence in this RfD is high. Confidence in the principal study fSuzuki etal.. 2012: TPEC.
2010a) is high. This study was well conducted, complied with OECD guidelines for GLP studies,
involved a sufficient number of animals per group (including both sexes), and assessed a wide
range of tissues and endpoints. Confidence in the database is high. The available studies evaluated a
comprehensive array of endpoints, and that additional studies would lead to identification of a
more sensitive endpoint is not indicated. Reflecting high confidence in the principal study and high
confidence in the database, confidence in the RfD is high.
2.1.7. Previous IRIS Assessment
No previous oral assessment for ETBE is available in IRIS.
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2.2. INHALATION REFERENCE CONCENTRATION FOR EFFECTS OTHER
THAN CANCER
The inhalation RfC (expressed in units of mg/m3) is defined as 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. It can be derived from a NOAEL, LOAEL, or the 95% lower
bound on the benchmark concentration (BMCL), with UF values generally applied to reflect
limitations of the data used.
2.2.1. Identification of Studies and Effects for Dose-Response Analysis
Kidney effects were identified as a potential human hazard ofETBE exposure based on
studies in experimental animals (summarized in Section 1.3.1). These studies were evaluated using
general study quality characteristics as discussed in Section 6 of the Preamble and in Section 1.1.1;
see also U.S. EPA f20021 to help inform the selection of studies from which to derive toxicity values.
Rationale for selection of studies and effects representative of this hazard is summarized below.
Human studies are generally preferred over animal studies as the basis for reference values
when quantitative measures of exposure are reported and the reported effects are determined to
be associated with exposure. Data on the effects of inhaled ETBE in humans is limited to a small
number of 2-hour inhalation studies at doses up to 208.9 mg/m3 (Nihlen et al.. 1998b: Vetrano.
1993). These studies were not considered for dose-response assessment because they are of acute
duration and investigated toxicokinetics.
The database for ETBE includes inhalation studies and data sets that are potentially suitable
for use in deriving inhalation reference values. Specifically, effects associated with ETBE exposure
in animals include observations of organ weight and histological changes in the kidney in chronic
and subchronic studies in male and female rats.
Kidney Toxicity
Evidence exists supporting kidney effects following ETBE exposure in rats, including organ
weight changes, histopathology (urothelial hyperplasia and CPN), and altered serum biomarkers
(creatinine, BUN, cholesterol). The most consistent, dose-related findings across multiple studies
were for kidney weight changes, CPN severity, and urothelial hyperplasia. In the case of kidney
weight changes, numerous chronic and subchronic studies investigated this endpoint following
inhalation exposure (Suzuki etal.. 2012: Hagiwara et al.. 2011: Fuiii etal.. 2010: TPEC. 2010b.
2008b. c; Gaoua. 2004b: Medinskv etal.. 1999). For urothelial hyperplasia and CPN, a 2-year study
by inhalation fSaito etal.. 2013: TPEC. 2010bl exposure reported this effect to be increased with
treatment in male rats. Therefore, CPN and urothelial hyperplasia data were the only endpoints
from the 2-year studies and kidney weights were the only endpoint from 13-week studies that were
considered for dose-response analysis (Saito etal.. 2013: TPEC. 2010b). Changes in serum
biomarkers lacked consistency and strength of association and were therefore not considered for
modeling.
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In the Saito etal. f20131 2-year inhalation study, male and female F344 rats (50/sex/dose
group) were exposed to concentrations of 0, 2,090, 6,270, or 20,900 mg/m3 fTPEC. 2010b).
Increased incidences of urothelial hyperplasia were only observed in males and significantly
increased at 6,270 and 20,900 mg/m3. Similar effects were not observed in females, thus the female
data were not modeled. Increased severity of CPN was significantly increased in males and females
at 20,900 mg/m3.
In the TPEC f2008bl 13-week whole-body inhalation study, male and female Crl:CD(SD) rats
were exposed to concentrations of 0, 627, 2,090, 6,270, or 20,900 mg/m3 for 6 hours/day,
5 days/week (65 exposures total). Significant increases in absolute kidney weights occurred in
male rats exposed to 6,270 or 20,900 mg/m3 ETBE compared with controls, while changes in
female rats were not statistically significant, and were not modeled.
In the Medinskv et al. f!9991 13-week whole-body inhalation study, male and female F344
rats were exposed to concentrations of 0, 2,090, 7,320, or 20,900 mg/m3 for 6 hours/day,
5 days/week. Kidney weights were increased at the highest two doses in both male and females.
Slight, but statistically significant, increases in various clinical chemistry parameters were
observed; however, these effects were reported to be of uncertain toxicological significance and
were not modeled.
2.2.2. Methods of Analysis
No biologically based dose-response models are available for ETBE. In this situation, dose-
response models thought to be consistent with underlying biological processes were evaluated to
determine how best to model the dose-response relationship empirically in the range of the
observed data. Consistent with this approach, all models available in EPA's BMDS were evaluated.
Consistent with EPA's Benchmark Dose Technical Guidance Document fU.S. EPA. 2012). the BMC and
the 95% BMCL were estimated using BMR to represent a minimal, biologically significant level of
change. As noted in Section 2.1.2, a 10% relative change from the control mean (relative deviation;
RD) was used as a BMR for absolute kidney weight, and a BMR of 10% extra risk was considered
appropriate for the quantal data on incidences of urothelial hyperplasia. When modeling was
feasible, the estimated BMCLs were used as points of departure (PODs); the PODs are summarized
in Table 2-4. Further details including the modeling output and graphical results for the model
selected for each endpoint are found in Appendix C of the Supplemental Information to this
Toxicological Review.
Because the RfC is applicable to a continuous lifetime human exposure but is derived from
animal studies featuring intermittent exposure, EPA guidance fU.S. EPA. 19941 provides
mechanisms for: (1) adjusting experimental exposure concentrations to a value reflecting
continuous exposure duration (ADJ) and (2) determining a human equivalent concentration (HEC)
from the animal exposure data. The former employs an inverse concentration-time relationship to
derive a health-protective duration adjustment to time-weight the intermittent exposures used in
the studies. The modeled benchmark concentration from the animal exposures in both inhalation
studies fTPEC. 2008b: Medinskv etal.. 19991 were adjusted to reflect a continuous exposure by
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multiplying concentration by (6 hours/day) 4- (24 hours/day) and (5 days/week) 4 (7 days/week)
as follows:
BMCLadj = BMCL (mg/m3) x (6 -h 24) x (5 4 7)
BMCL (mg/m3) x (0.1786)
The RfC methodology provides a mechanism for deriving an HEC from the duration-
adjusted POD (BMCLadj) determined from the animal data. The approach takes into account the
extra-respiratory nature of the toxicological responses and accommodates species differences by
considering blood:air partition coefficients for ETBE in the laboratory animal (rat or mouse) and
humans. According to the RfC guidelines fU.S. EPA. 19941. ETBE is a Category 3 gas because extra-
respiratory effects were observed. Therefore, the duration-adjusted BMCLadj is multiplied by the
ratio of animal/human blood:air partition coefficients (La/Lh). As detailed in Appendix B.2.2 of the
Supplemental Information, the values reported in the literature for these parameters include an La
of 11.6 for Wistar rats fKaneko etal.. 20001 and an Lh in humans of 11.7 fNihlen et al.. 19951. This
allowed a BMCLhec to be derived as follows:
BMCLhec = BMCLadj (mg/m3) x (La 4 Lh) (interspecies conversion)
= BMCLadj (mg/m3) x (11.6 4 11.7)
= BMCLadj (mg/m3) x (0.992)
Table 2-4 summarizes the sequence of calculations leading to the derivation of a human-
equivalent POD (PODhec) for each inhalation data set discussed above.
Table 2-4. Summary of derivation of PODs following inhalation exposure
Endpoint and
Reference
Species/
Sex
Model3
BMR
BMC
(mg/m3)
BMCL (mg/m3)
PODadj"
(mg/m3)
PODhec0
(mg/m3)
Kidney
Increased urothelial
hyperplasia; 2-year
Saito et al. (2013);
JPEC (2010b)
Male F344
rats
Gamma
10%
2,734
1,498
268
265
Increased CPN
severity; 2-year
Saito et al. (2013);
JPEC (2010b)
Male and
female F344
rats
NOAELd: 6270 mg/m3
1,120
1,110
Increased absolute
kidney weight;
13- week
JPEC (2008b)
Male
Sprague-
Dawley rats
NOAELd: 627 mg/m3
10% 1" in kidney weight
112
111
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Endpoint and
Reference
Species/
Sex
Model3
BMR
BMC
(mg/m3)
BMCL (mg/m3)
PODadj"
(mg/m3)
PODhec0
(mg/m3)
Increased absolute
kidney weight;
13-week
JPEC (2008b)
Female
Sprague-
Dawley rats
Linear
10% RD
28,591
16,628
2,969
2,946
Increased absolute
kidney weight;
13-week
Medinskv et al.
(1999)
Male F344
rats
Hill
10% RD
6,968
2,521
450
447
Increased absolute
kidney weight;
13-week
Medinskv et al.
(1999)
Female F344
rats
Exponenti
al (M4)
10% RD
5,610
3,411
609
604
aFor modeling details, see Appendix C of the Supplemental Information.
bPODs were adjusted for continuous daily exposure: PODadj = POD x (hours exposed per day -f 24 hr) x (days
exposed per week 4 7 days).
cPODhec calculated by adjusting the PODadj by the DAF (=0.992) for a Category 3 gas (U.S. EPA, 1994).
dNOAEL was used due to lack of suitable model fit (see Appendix C).
2.2.3. Derivation of Candidate Values
In EPA's A Review of the Reference Dose and Reference Concentration Processes fU.S. EPA.
2002: Section 4.4.51. also described in the Preamble, five possible areas of uncertainty and
variability were considered. An explanation follows:
An intraspecies uncertainty factor, UFh, of 10 was applied to all PODs to account for
potential differences in toxicokinetics and toxicodynamics in the absence of information on the
variability of response in the human population following inhalation exposure to ETBE fU.S. EPA.
20021.
An interspecies uncertainty factor, UFa, of 3 (10°5 = 3.16, rounded to 3) was applied to all
PODs to account for residual uncertainty in the extrapolation from laboratory animals to humans in
the absence of information to characterize toxicodynamic differences between rodents and humans
after inhalation exposure to ETBE. This value is adopted by convention where an adjustment from
animal to a human equivalent concentration has been performed as described in EPA's Methods for
Derivation of Inhalation Reference Concentrations and Application of Inhalation Dosimetry fU.S. EPA.
19941.
A subchronic to chronic uncertainty factor, UFs, differs depending on the exposure duration.
For rodent studies, exposure durations of 90 days (or 13 weeks) are generally considered
subchronic. Furthermore, the magnitude of change in absolute kidney weights appeared to increase
in male and female rats exposed for 26 weeks compared with 13-18 weeks, when results across
oral and inhalation exposures were evaluated based upon of internal blood concentrations (see
Figure 1-2), suggesting that toxicity would be expected to increase with exposure durations greater
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than 13 weeks. Therefore, a UFs of 10 was applied for studies of 13 weeks. A UFs of 1 was applied to
2-year studies.
A LOAEL to NOAEL uncertainty factor, UFl, of 1 was applied to all PODs derived because the
current approach is to address this factor as one of the considerations in selecting a BMR for
benchmark dose modeling. In this case, BMRs of a 10% change or a NOAEL in absolute kidney
weight or CPN and a 10% extra risk of urothelial hyperplasia were selected under an assumption
that they represent minimal biologically significant changes.
A database uncertainty factor, UFd, of 1 was applied to all PODs. The ETBE inhalation
toxicity database includes a 2-year toxicity study in rats fSaito etal.. 2013: TPEC. 2010bl and
13-week toxicity studies in mice and rats fTPEC. 2008b: Medinskv etal.. 19991. There are no
developmental or multi-generation reproductive studies by the inhalation route; however,
considering systemic effects such as these are anticipated to be similar via oral or inhalation
exposure to ETBE, first pass effects are not indicated by the available data, and no evidence is
available to suggest that untransformed ETBE would have a significant role in toxicity, the oral
studies of prenatal developmental toxicity in rats and rabbits fAso etal.. 2014: Asano etal.. 20111.
and single- and multi-generation reproductive toxicity and developmental toxicity in rats fFuiii et
al.. 2010: Gaoua. 2004a. b) are available to inform the inhalation database. Similarly, the oral ETBE
data set does not indicate immunotoxicity and differences in outcome would not be anticipated for
inhalation exposures fBanton etal.. 2011: Li etal.. 20111. Although most of the studies are in rats,
the available mouse study observed effects that were less severe than those in rats, suggesting that
mice are not more sensitive than rats. The ETBE inhalation database, supported by the information
from the oral database, adequately covers all major systemic effects, including reproductive,
developmental, immunological and neurological effects, and does not suggest that additional
studies would lead to identification of a more sensitive endpoint or a lower POD. Therefore, a
database UFd of 1 was applied.
Table 2-5 is a continuation of Table 2-4, and summarizes the application of UF values to
each POD to derive a candidate value for each data set. The candidate values presented in the table
below are preliminary to the derivation of the organ/system-specific reference values. These
candidate values are considered individually in the selection of a representative inhalation
reference value for a specific hazard and subsequent overall RfC for ETBE.
Figure 2-2 presents graphically the candidate values, UF values, and PODs, with each bar
corresponding to one data set described in Table 2-4 and Table 2-5.
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1 Table 2-5. Effects and corresponding derivation of candidate values
Endpoint (Sex and species) and
Reference
PODhec
(mg/m3)
POD
type
UFa
UFh
UFl
UFs
UFd
Composite
UF
Candidate
value
(mg/m3)
Kidney
Increased urothelial hyperplasia;
male rat; 2-year
Saito et al. (2013); JPEC (2010b)
265
BMCLio%
3
10
1
1
1
30
9x 10°
Increased CPN severity; male and
female rats; 2-year
Saito et al. (2013); JPEC (2010b)
1,110
NOAEL
3
10
1
1
1
30
4x 101
Increased absolute kidney weight;
male rat; 13-week
JPEC (2008b)
111
NOAEL
3
10
1
10
1
300
4 x 101
Increased absolute kidney weight;
female rat; 13-week
JPEC (2008b)
2,946
BMCLio%
3
10
1
10
1
300
lx 101
Increased absolute kidney weight;
male rat; 13-week
Medinskv et al. (1999)
447
BMCLio%
3
10
1
10
1
300
2x 10°
Increased absolute kidney weight;
female rat; 13-week
Medinskv et al. (1999)
604
BMCLio%
3
10
1
10
1
300
2x 10°
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Increased urothelial hyperplasia:
male rat; 2 year
Saito #t ai. (2013); J PEC (20t Obi
Increased CPN severity;
male and female rals; 2 year
Saitoet at. (20! 3}; j PEC 2010b
Increased absolute kidney weight;
male rat; 1.3 week
J PEC (200 8b)
Increased absolute kidney weight;
i Ml, 13
JPEC (2008b)
Increase! absolute kidney weight;
male Kit; 13 week
Medio$ky et al, (1999)
Increased absolute kidney weight;
female rat; 13 week
Metlinsky et al, (1999)
0.1
~ Candidate EMC
• roote
S'S Composite UF
16 100
mg/m?
looo iocmo
Figure 2-2. Candidate values with corresponding POD and composite UF.
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2.2.4. Derivation of Organ/System-Specific Reference Concentrations
Table 2-6 distills the candidate values from Table 2-5 into a single value for the kidney.
Organ- or system-specific reference values can be useful for subsequent cumulative risk
assessments that consider the combined effect of multiple agents acting at a common site.
Kidney Toxicity
For ETBE, candidate values were derived for increased kidney weight in both sexes of rats,
and urothelial hyperplasia in males, spanning a range from 4 x 101 to 4 x 101 mg/m3, for an overall
100-fold range. To estimate an exposure level below which kidney toxicity from ETBE exposure is
not expected to occur, the candidate RfC for increased incidence of urothelial hyperplasia in male
rats (9 x 10° mg/m3) was selected as the kidney-specific RfC for ETBE, consistent with the
selection of the kidney-specific RfD (see Section 2.1.4). As discussed in Section 2.1.4, this lesion is a
more specific and more sensitive indicator of kidney toxicity, compared with the relatively
nonspecific endpoint of kidney weight change, and is synonymous with the transitional epithelial
hyperplasia in the kidney observed after chronic tert-butanol exposure described in NTP f!995al
Finally, the Toxicological Review of tert-butanol identified transitional epithelial hyperplasia in the
kidney as the lowest POD, further supporting this endpoint as a sensitive indicator of kidney
toxicity. Confidence in this kidney-specific RfC is high. The PODs are based on BMD modeling, and
the candidate values are derived from well-conducted studies, involving a sufficient number of
animals per group, including both sexes, and assessing a wide range of kidney endpoints.
Table 2-6. Organ-/system-specific RfCs and overall RfC for ETBE
Effect
Basis
RfC (mg/m3)
Study exposure
description
Confidence
Kidney
Incidence of urothelial
hyperplasia
Saito etal. (2013); JPEC
(2010b)
9 x 10°
Chronic
High
Overall RfC
Kidney
9 x 10°
Chronic
High
2.2.5. Selection of the Overall Reference Concentration
For ETBE, kidney effects were identified as the primary hazard; thus, a single
organ-/system-specific RfC was derived. Therefore, the kidney-specific RfC of 9 x 10° mg/m3 is
selected as the overall RfC, representing an estimated exposure level below which deleterious
effects from ETBE exposure are not expected to occur.
The overall RfC is derived to be protective for all types of effects for a given duration of
exposure and is intended to protect the population as a whole including potentially susceptible
subgroups fU.S. EPA. 20021. Decisions concerning averaging exposures over time for comparison
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with the RfC should consider the types of toxicological effects and specific lifestages of concern.
Fluctuations in exposure levels that result in elevated exposures during these lifestages could lead
to an appreciable risk, even if average levels over the full exposure duration were less than or equal
to the RfC. In the case ofETBE, no specific potential for early lifestage susceptibility to ETBE
exposure was identified, as discussed in Section 1.3.3.
2.2.6. Confidence Statement
A confidence level of high, medium, or low is assigned to the study used to derive the RfC,
the overall database, and the RfC itself, as described in Section 4.3.9.2 of EPA's Methods for
Derivation of Inhalation Reference Concentrations and Application of Inhalation Dosimetry fU.S. EPA.
19941. The overall confidence in this RfC is high. Confidence in the principal study, Saito et al.
f20131: TPEC f2010bl. is high. This study was well conducted, following GLP guidelines that
involved a sufficient number of animals per group (including both sexes), and assessed a wide
range of tissues and endpoints. Confidence in the database is high; the available studies evaluated a
comprehensive array of endpoints, and that additional studies would lead to identification of a
more sensitive endpoint is not indicated. Reflecting high confidence in the principal studies and
high confidence in the database, overall confidence in the RfC for ETBE is high.
2.2.7. Previous IRIS Assessment
No previous inhalation assessment for ETBE is available in IRIS.
2.2.8. Uncertainties in the Derivation of the Reference Dose and Reference Concentration
The following discussion identifies uncertainties associated with the RfD and RfC for ETBE.
To derive the RfD and RfC, the UF approach fU.S. EPA. 2000.19941 was applied to a POD based on
kidney toxicity in rats treated chronically. UFs were applied to the PODs to account for
extrapolating from an animal bioassay to human exposure and for the likely existence of a diverse
human population of varying susceptibility. Default approaches are used for these extrapolations,
given the lack of data to inform individual steps.
The database for ETBE contains no human data on adverse health effects from subchronic
or chronic exposure, and the PODs were calculated from data on the effects ofETBE reported by
studies in rats. The database for ETBE exposure includes three lifetime bioassays in rats, several
reproductive/developmental studies in rats and rabbits, several subchronic studies in rats and
mice, and immunotoxicity assays.
Although the database is adequate for reference value derivation, some uncertainty
associated with the database remains, such as the lack of chronic studies in a species other than rats
(e.g., mice), the lack of developmental/reproductive inhalation studies, and no information
available regarding kidney or liver toxicity in animals with deficient ALDH2 activity.
The toxicokinetic and toxicodynamic differences for ETBE between the animal species from
which the POD was derived and humans are unknown. Although sufficient information is available
to develop a PBPK model in rats to evaluate differences across routes of exposure, the ETBE
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database lacks an adequate model that would inform potential interspecies differences. Generally,
males appear more susceptible than females to ETBE toxicity. The underlying mechanistic basis of
this apparent difference, however, is not understood. Most importantly, which animal species and
sexes are more comparable to humans is unknown.
The ETBE data are insufficient to conclude that the a2U-globulin process is operative;
however, noncancer effects related to a2U-globulin were considered not relevant for hazard
identification and, therefore, not suitable for dose-response consideration. Based on the candidate
RfCs in Table 2-5 and Figure 2-2, increased CPN severity would not be selected as a critical
endpoint, even if the effect were assumed relevant to humans. Changes in absolute kidney weights
for male rats and for female rats in some studies, however, result in lower RfD or RfC values than
urothelial hyperplasia. So, if the a2U-globulin process were determined responsible for all male
kidney toxicity, female kidney weight could be used to derive a POD that is lower than the current
value. If kidney noncancer effects were determined not relevant to humans, absolute kidney
weights would still be a relevant endpoint because subchronic kidney weights were used for dose-
response analysis and CPN severity was elevated only after chronic exposures.
2.3. ORAL SLOPE FACTOR FOR CANCER
The oral slope factor (OSF) is a plausible upper bound on the estimate of risk per
mg/kg-day of oral exposure. The OSF can be multiplied by an estimate of lifetime exposure (in
mg/kg-day) to estimate the lifetime cancer risk.
2.3.1. Analysis of Carcinogenicity Data
As noted in Section 1.3.2, EPA concluded that there is "suggestive evidence of carcinogenic
potential" for ETBE. The Guidelines for Carcinogen Risk Assessment (U.S. EPA. 2005a) state:
When there is suggestive evidence, the Agency generally would not attempt a
dose-response assessment, as the nature of the data generally would not support
one; however when the evidence includes a well-conducted study, quantitative
analysis may be useful for some purposes, for example, providing a sense of the
magnitude and uncertainty of potential risks, ranking potential hazards, or setting
research priorities.
A PBPK model is used to derive oral values from the inhalation POD based on an endpoint
reported in Saito etal. (20131 and TPEC (2010b). A description of the carcinogenicity data is
presented in the discussions of biological considerations for cancer dose-response analysis (see
Section 1.3.2).
2.3.2. Dose-Response Analysis—Adjustments and Extrapolation Methods
The EPA Guidelines for Carcinogen Risk Assessment (U.S. EPA. 2005a) recommend that
determining the method to use for characterizing and quantifying cancer risk from a chemical be
based on what is known about the MOA of the carcinogen and the shape of the cancer dose-
response curve. EPA uses a two-step approach that distinguishes analysis of the observed dose-
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response data from inferences about lower doses fU.S. EPA. 2005al. Within the observed range, the
preferred approach is to use modeling to incorporate a wide range of data into the analysis, such as
through a biologically based model, if supported by substantial data. Without a biologically based
model, as in the case of ETBE, a standard model is used for curve-fitting the data and estimating a
POD. EPA uses the multistage model in IRIS dose-response analyses for cancer fGehlhaus etal..
20111 because it parallels the multistage carcinogenic process and fits a broad array of dose-
response patterns.
The second step, extrapolation to lower exposures from the POD, considers what is known
about the modes of action for each effect As above, a biologically based model is preferred (U.S.
EPA. 2005a). Otherwise, linear low-dose extrapolation is recommended if the MOA of
carcinogenicity is mutagenic or has not been established (U.S. EPA. 2005a). For ETBE, the mode(s)
of carcinogenic action for liver tumors has not been established (see Section 1.3.2). Therefore,
linear low-dose extrapolation was used to estimate human carcinogenic risk.
A PBPK model for ETBE in rats has been applied as described in Appendix B of the
Supplemental Information. Using this model, route-to-route extrapolation of the inhalation BMCL to
derive an oral POD was performed as follows. First, the internal dose in the rat at the inhalation
BMCL (assuming the same periodic exposure profile used by the bioassay) was estimated using the
PBPK model to derive an "internal dose BMDL." Then, the oral dose (assuming a circadian drinking
water exposure profile) that led to the same internal dose in the rat was estimated using the PBPK
model, resulting in a route-to-route extrapolated BMDL.
A critical decision in the route-to-route extrapolation is the selection of the internal dose
metric for establishing "equivalent" oral and inhalation exposures. For ETBE-induced liver tumors,
the four options are the (1) concentration of tert-butanol in blood, (2) rate of tert-butanol
metabolism in the liver, (3) concentration ofETBE in blood, and (4) rate ofETBE metabolism in the
liver fSalazar etal.. 20151. The major systemically available metabolite ofETBE is tert-butanol,
which has not been shown to cause liver toxicity, so tert-butanol blood concentration and tert-
butanol metabolism are not plausible dose metrics. ETBE in the blood also is not supported as a
dose metric because liver concentrations ofETBE are more proximal to the site of interest Liver
concentration for ETBE will lead to a similar route-to-route extrapolation relationship as using liver
metabolism ofETBE because metabolism is a function of the liver concentration. Since metabolism
is saturable and the degree of metabolic saturation can vary with dose and dose-rate, there is likely
to be some difference between using these two metrics for extrapolation. Further, if the BMCL is in
the linear metabolic range, then the route-to-route extrapolation will be independent of the choice
between ETBE concentration in liver and ETBE metabolism. While this computational equivalence
exists, use of the rate of metabolism ofETBE in the liver accounts for the possible role of
acetaldehyde, the other metabolite ofETBE produced in the liver, which is a genotoxic carcinogen.
Consequently, the rate of metabolism ofETBE was selected as the best available basis for route-to-
route extrapolation.
The data modeled and other details of the modeling are provided in Appendix C. The BMDs
and BMDLs recommended for each data set are summarized in Table 2-7. The route-to-route
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extrapolated ETBE BMDL is scaled to an HED according to EPA guidance fU.S. EPA. 2011. 2005a). In
particular, the BMDL was converted to an HED assuming that doses in animals and humans are
toxicologically equivalent when scaled by body weight raised to the 3/4 power. Standard body
weights of 0.25 kg for rats and 70 kg for humans were used (U.S. EPA. 19881. The following formula
was used for the conversion of an oral BMDL to an oral HED:
Scaled HED in mg/kg-d = (BMDL in mg/kg-d) x (0.25/70)1/4
= (BMDL in mg/kg-d) x 0.2445
PODs for estimating low-dose risk were identified at doses at the lower end of the observed
data, corresponding to 10% extra risk.
2.3.3. Derivation of the Oral Slope Factor
The results from route-to-route extrapolation of the male rat liver tumor data fSaito etal..
2013: TPEC. 2010b) are summarized in Table 2-7. The lifetime oral cancer slope factor for humans is
defined as the slope of the line from the lower 95% bound on the exposure at the POD to the control
response (slope factor = BMR/BMDLbmr = 0.1/BMDLio). This slope represents a plausible upper
bound on the true population average risk. Using linear extrapolation from the BMDLio, a human
equivalent oral slope factor was derived as presented in Table 2-7.
A single oral slope factor was derived. The recommended oral slope factor for providing a
sense of the magnitude of potential carcinogenic risk associated with lifetime oral exposure to
ETBE is 1 x 10"3 per mg/kg-day based on the liver tumor response in male F344 rats (Saito etal..
2013: TPEC. 2010b). This slope factor should not be used with exposures exceeding 402 mg/kg-day
(the POD), because above this level the cancer risk might not increase linearly with exposure. The
slope of the linear extrapolation from the central estimate BMDiohed is 0.1/0.2445 x (525 mg/kg-
day)] = 8 x 10 4per mg/kg-day.
Table 2-7. Summary of the oral slope factor derivation
Tumor
Species/Sex
BMR
BMC
(mg/m3)
BMCL
(mg/m3)
Internal
BMC
Dose3
(mg/h)
Internal
BMCL
Doseb
(mg/h)
BMDC
(mg/kg-
d)
POD=
BMDLC
(mg/kg-
d)
BMDLhed"
(mg/kg-
d)
Slope
Factore
(mg/kg-d)1
Hepatocellular
adenomas and
carcinomas
Saito et al.
Male F344
rat
10%
10,884
7,118
2.517
1.977
525
401.8
98.2
1 X 10"3
(2013): JPEC
(2010b)
aAverage rate of ETBE metabolism in rats under 6 hour/day, 5 days/week inhalation exposure at the BMC.
bAverage rate of ETBE metabolism in rats under 6 hour/day, 5 days/week inhalation exposure at the BMCL.
cOral exposure in rats under circadian drinking water ingestion that leads to the same average rate of ETBE
metabolism as 6 hour/day, 5 days/week inhalation exposure in rats at the BMC/BMCL.
Continuous oral exposure human equivalent dose = BMDL x (0.25/70)^.
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eHuman equivalent oral slope factor = 0.1/BMDLhed.
2.3.4. Uncertainties in the Derivation of the Oral Slope Factor
Uncertainty exists when extrapolating data from animals to estimate potential cancer risks
to human populations from exposure to ETBE.
Table 2-8 summarizes several uncertainties that could affect the oral slope factor. Although
the 2-year cancer bioassays did not report an increase in liver tumorigenesis following oral
exposure in rats, increased liver tumorigenesis in male rats was observed in a 2-year inhalation
bioassay and several initiation-promotion bioassays. No other studies are available to replicate
these findings and none examined other animal models (e.g., mice). Additionally, no data in humans
are available to confirm a cancer response in general or the specific tumors observed in the rat
bioassay fSaito etal.. 2013: TPEC. 2010bl Although changing the methods used to derive the oral
slope factor could change the results, standard practices were used due to the lack of a human
PBPK model, and no other data (e.g., MOA) supported alternative derivation approaches.
Table 2-8. Summary of uncertainties in the derivation of the oral slope factor
for ETBE
Consideration and
Impact on Cancer Risk Value
Decision
Justification and Discussion
Selection of tumor type and
relevance to humans:
Rat liver tumors are the basis for
estimating human cancer risk.
Liver tumors in male rats
were selected.
An MOA for liver carcinogenicity could not be
established, so rat liver tumors were
considered relevant to humans U.S. EPA
(2005a).
Selection of data set:
No other 2-year studies are
available.
Saito etal. (2013), JPEC
(2010b) inhalation studv
was selected to derive oral
cancer risks for humans.
Saito et al. (2013), JPEC (2010b) was a well-
conducted study and the only lifetime
exposure bioassay that reported increased
liver tumors. No guidance for quantifying a
lifetime cancer risk arising from promotion of
mutagen-induced tumors is available.
Additional bioassays might add support to
the findings or provide results for different
doses, which could affect the oral slope
factor.
Selection of extrapolation approach:
Different PBPK model could 4, or T*
oral slope factor.
PBPK model-based
extrapolation of inhalation
data was used for oral
slope factor.
The PBPK model accurately predicted ETBE
toxicokinetics. Difference in oral slope factor
derived using the Salazar et al. (2015) and
Borghoff et al. (2016) models was
approximately 20%.
Selection of dose metric:
Alternatives could 4' or T* oral
slope factor.
ETBE metabolism rate as
the dose metric for route-
to-route extrapolation was
converted to HED.
ETBE metabolized is the best-supported dose
metric. It is consistent with a hypothesis that
acetaldehyde plays a role in liver
carcinogenesis of ETBE. It is also consistent
with ETBE concentration in the liver as the
mediator of carcinogenesis (metabolism is
approximately proportional to ETBE liver
concentration). Alternative dose metrics of
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Consideration and
Impact on Cancer Risk Value
Decision
Justification and Discussion
ETBE concentration, te/t-butanol
concentration, or te/t-butanol metabolism
would result in a range of 50% decrease to
25% increase in the oral slope factor.
Interspecies extrapolation of
dosimetry and risk:
Alternatives could 4^ or T* slope
factor (e.g., 3.5-fold 4^ scaling by
body weight] or T* 2-fold scaling by
BW2/3]).
The default approach of
BW3/4 was used.
No data suggest an alternative approach for
ETBE. Because the dose metric was not an
area under the curve, BW3/4 scaling was used
to calculate equivalent cumulative exposures
for estimating equivalent human risks.
Although the true human correspondence is
unknown, this overall approach is expected
to neither overestimate nor underestimate
human equivalent risks.
Dose-response modeling:
Alternatives could 4^ or T* slope
factor.
Used multistage dose-
response model to derive
BMD and BMDL
No biologically based models for ETBE were
available. The multistage model has
biological support and is the model most
consistently used in EPA cancer assessments.
Low-dose extrapolation:
4/ cancer risk estimate would be
expected with the application of
nonlinear low-dose extrapolation.
Linear extrapolation of risk
in low-dose region used
U.S. EPA (1998a).
Linear low-dose extrapolation for agents
without a known MOA is supported U.S. EPA
(2005a).
Statistical uncertainty at POD:
4/ oral slope factor 1.5-fold if BMD
used as the POD rather than BMDL
BMDL (preferred approach
for calculating slope
factor).
Limited size of bioassay results in sampling
variability; lower bound is 95% CI on
administered exposure at 10% extra risk of
liver.
Sensitive subpopulations:
T* oral slope factor to unknown
extent.
Individuals deficient in
ALDH2 are potentially
more sensitive; individuals
pre- or co-exposed to
mutagenic carcinogens
could be more sensitive.
Experiments showed enhanced liver toxicity
and genotoxicity in mice when ALDH2 was
absent. Human subpopulations deficient in
ALDH2 are known to be at enhanced risk of
ethanol-induced cancer mediated by
acetaldehyde. No chemical-specific data are
available, however, to determine the extent
of enhanced susceptibility due to ETBE-
induced carcinogenicity. ETBE promotion of
mutagen-induced tumors in rat tissues not
identified as hazards of ETBE toxicity
suggests that ETBE could enhance
carcinogenesis through an undetermined
MOA. Beyond ALDH deficiency, no chemical-
specific data are available to determine the
range of human toxicodynamic variability or
sensitivity, including the susceptibility of
children. Because determination of a
mutagenic MOA has not been made, an age-
specific adjustment factor is not applied.
2.3.5. Previous IRIS Assessment: Oral Slope Factor
No previous cancer assessment for ETBE is available in IRIS.
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2.4. INHALATION UNIT RISK FOR CANCER
The carcinogenicity assessment provides information on the carcinogenic hazard potential
of the substance in question, and quantitative estimates of risk from inhalation exposure can be
derived. Quantitative risk estimates can be derived from the application of a low-dose extrapolation
procedure. If derived, the inhalation unit risk is a plausible upper bound on the estimate of risk per
|ig/m3 air breathed.
2.4.1. Analysis of Carcinogenicity Data
As noted in Section 1.3.2, there is "suggestive evidence of carcinogenic potential" for ETBE.
A description of the carcinogenicity data is presented in the discussions of biological considerations
for cancer dose-response analysis (see Section 1.3.2). For hepatocellular adenomas and carcinomas,
statistical tests conducted by the study authors found significant dose-response trends by both the
Peto test (incidental tumor test) and the Cochran-Armitage test. Therefore, the hepatocellular
adenomas and carcinomas in male rats were considered for unit risk derivation.
2.4.2. Dose-Response Analysis—Adjustments and Extrapolation Methods
The EPA Guidelines for Carcinogen Risk Assessment fU.S. EPA. 2005al recommend that the
method used to characterize and quantify cancer risk from a chemical be determined by what is
known about the MOA of the carcinogen and the shape of the cancer dose-response curve. EPA uses
a two-step approach that distinguishes analysis of the observed dose-response data from
inferences about lower doses (U.S. EPA. 2005a). Within the observed range, the preferred approach
is to use modeling to incorporate a wide range of data into the analysis, such as through a
biologically based model, if supported by substantial data. Without a biologically based model, as in
the case of ETBE, a standard model is used to curve-fit the data and to estimate a POD. EPA uses the
multistage model in IRIS dose-response analyses for cancer fGehlhaus etal.. 20111 because it
parallels the multistage carcinogenic process and fits a broad array of dose-response patterns.
The second step, extrapolation to lower exposures from the POD, considers what is known
about the modes of action for each effect As above, a biologically based model is preferred (U.S.
EPA. 2005a). Otherwise, linear low-dose extrapolation is recommended if the MOA of
carcinogenicity is mutagenic or has not been established (U.S. EPA. 2005a). For ETBE, the mode(s)
of carcinogenic action for liver tumors has not been established (see Section 1.3.2). Therefore,
linear low-dose extrapolation was used to estimate human carcinogenic risk.
Details of the modeling and the model selection process can be found in Appendix C of the
Supplemental Information. APOD for estimating low-dose risk was identified at the lower end of
the observed data, corresponding to 10% extra risk.
Because the inhalation unit risk is applicable to a continuous lifetime human exposure but
derived from animal studies featuring intermittent exposure, EPA guidance (U.S. EPA. 1994)
provides mechanisms for (1) adjusting experimental exposure concentrations to a value reflecting
continuous exposure duration and (2) determining a human equivalent concentration (HEC) from
the animal exposure data. The former uses an inverse concentration-time relationship to derive a
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health-protective duration adjustment to time weight the intermittent exposures used in the study.
The animal BMCL (Table 2-7) estimated from the inhalation study fSaito etal.. 2013: TPEC. 2010b)
was adjusted to reflect continuous exposure by multiplying it by (6 hours/day) 4- (24 hours/day)
and (5 days/week) 4 (7 days/week) as follows:
BMCLadj = BMCL (mg/m3) x (6 -h 24) x (5 4 7)
= 7,118 mg/m3 x 0.25 x 0.71
= 1,271 mg/m3
The approach to determine the HEC accounts for the extrarespiratory nature of the
toxicological responses and accommodates species differences by considering blood:air partition
coefficients for ETBE in the laboratory animal (rat) and humans. According to the RfC guidelines
fU.S. EPA. 19941. ETBE is a Category 3 gas because extrarespiratory effects were observed. The
values reported in the literature for these parameters include a blood:air partition coefficient of
11.6 for rats fKaneko etal.. 20001 and a blood:air partition coefficient for humans of 11.7 (Nihlenet
al.. 19951. This allowed a BMCLhec to be derived as follows:
BMCLhec = BMCLadj (mg/m3) x (LA 4 Lh) (interspecies conversion)
BMCLadj (mg/m3) x (11.6 4 11.7)
= BMCLadj (mg/m3) x (0.992)
1,271 mg/m3 x (0.992)
= 1,261 mg/m3
2.4.3. Inhalation Unit Risk Derivation
The POD estimate based on the male rat liver tumor data (Saito etal.. 2013: TPEC. 2010b) is
summarized in Table 2-9. The lifetime inhalation unit risk for humans is defined as the slope of the
line from the lower 95% bound on the exposure at the POD to the control response (inhalation unit
risk = 0.1 4 BMCLio). This slope represents a plausible upper bound on the true risk. Using linear
extrapolation from the BMCLio, a human-equivalent inhalation unit risk was derived as presented
in Table 2-9.
A single inhalation unit risk was derived. Therefore, the recommended inhalation unit risk
for providing a sense of the magnitude of potential carcinogenic risk associated with lifetime
inhalation exposure to ETBE is 8 x 10-s per mg/m3, based on the liver tumor response in male
F344 rats fSaito etal.. 2013: TPEC. 2010bl. This unit risk should not be used with continuous
exposures exceeding 1,271 mg/m3 (the POD) because above this level the cancer risk might not
increase linearly with exposure. The slope of the linear extrapolation from the central estimate
BMDio is 0.1 4 0.992 x (1,944 mg/kg-day)] = 5 x 10"5 per mg/m3.
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Toxicological Review ofETBE
1 Table 2-9. Summary of the inhalation unit risk derivation
Tumor
Species/Sex
Selected
Model
BMR
BMCadj
(mg/m3)
POD=
BMCLadj
(mg/m3)
BMCLhec
(mg/m3)
Slope
factor3
(mg/m3)1
Hepatocellular
adenomas or
carcinomas
Saito et al. (2013);
Male F344 rat
1°
Multistage
10%
1,944
1,271
1,261
8 x 10"5
JPEC (2010b)
2 aHuman equivalent slope factor = 0.1/BMCLiohec; see Appendix C of the Supplemental Information for details of
3 modeling results.
4 2.4.4. Uncertainties in the Derivation of the Inhalation Unit Risk
5 Uncertainty exists when extrapolating data from animals to estimate potential cancer risks
6 to human populations from exposure to ETBE.
7 Table 2-10 summarizes several uncertainties that could affect the inhalation unit risk.
8 Although the chronic studies did not report an increase in liver tumorigenesis following oral
9 exposure in rats, no other inhalation studies are available to replicate these findings and none
10 examined other animal models. In addition, no data in humans are available to confirm a general
11 cancer response or the specific tumors observed in the rat bioassay f Saito etal.. 2013: TPEC. 2010b).
12 Although changing the methods used to derive the inhalation unit risk could change the results,
13 standard practices were used due to the lack of a human PBPK model, and no other data (e.g., MOA)
14 supported alternative derivation approaches.
15 Table 2-10. Summary of uncertainties in the derivation of the inhalation unit
16 risk for ETBE
Consideration and
Impact on Cancer Risk Value
Decision
Justification and Discussion
Selection of tumor type and
relevance to humans:
Rat liver tumors are the basis for
estimating human cancer risk.
The liver was selected as
the target organ (U.S. EPA,
2005a).
An MOA for liver carcinogenicity could not be
established, so rat liver tumors were
considered relevant to humans supported
(U.S. EPA, 2005a).
Selection of data set:
No other studies are available.
Saito etal. (2013),JPEC
(2010b) was selected to
derive cancer risks for
humans.
Saito et al. (2013), JPEC (2010b) was a well-
conducted inhalation study and the only
bioassay that reported increased liver
tumors. Additional bioassays might add
support to the findings or provide results for
different (possibly lower) doses, which could
affect the oral slope factor.
Selection of dose metric:
Alternative could 4, inhalation unit
risk.
Administered
concentration was used.
Modeling based on the best-supported PBPK
model-based internal dose metric of ETBE
metabolism decreased the BMCL by 35%.
Interspecies extrapolation of
dosimetry and risk:
The default approach for a
Category 3 gas was used.
No data suggest an alternative approach.
Although the true human correspondence is
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Toxicological Review ofETBE
Consideration and
Impact on Cancer Risk Value
Decision
Justification and Discussion
Alternatives could 4^ or T*
Inhalation unit risk.
unknown, this overall approach is expected
to neither overestimate nor underestimate
human equivalent risks.
Dose-response modeling:
Alternatives could 4^ or T* slope
factor.
Used multistage dose-
response model to derive a
BMC and BMCL
No biologically based models for ETBE were
available. The multistage model has
biological support and is the model most
consistently used in EPA cancer assessments.
Low-dose extrapolation:
4/ cancer risk estimate would be
expected with the application of
nonlinear low-dose extrapolation.
Linear extrapolation of risk
in low-dose region was
used.
Linear low-dose extrapolation for agents
without a known MOA is supported (U.S.
EPA, 2005a).
Statistical uncertainty at POD:
4/ inhalation unit risk 1.4-fold if
BMC used as the POD rather than
BMCL.
BMCL (preferred approach
for calculating slope factor)
was used.
Limited size of bioassay results in sampling
variability; lower bound is 95% CI on
administered exposure at 10% extra risk of
liver tumors.
Sensitive subpopulations
1" inhalation unit risk to unknown
extent.
Individuals deficient in
ALDH2 are potentially
more sensitive.
Experiments showed enhanced liver toxicity
and genotoxicity in mice when ALDH2 was
absent. Human subpopulations deficient in
ALDH2 are known to be at enhanced risk of
ethanol-induced cancer mediated by
acetaldehyde, discussed in Section 1.3.3. No
chemical-specific data are available,
however, to determine the extent of
enhanced sensitivity due to ETBE-induced
carcinogenicity. Beyond ALDH deficiency, no
chemical-specific data are available to
determine the range of human
toxicodynamic variability or sensitivity,
including the susceptibility of children.
Because determination of a mutagenic MOA
has not been made, an age-specific
adjustment factor is not applied.
1 2.4.5. Previous IRIS Assessment: Inhalation Unit Risk
2 No previous cancer assessment for ETBE is available in IRIS.
3 2.5. APPLICATION OF AGE-DEPENDENT ADJUSTMENT FACTORS
4 As discussed in the Supplemental Guidance for Assessing Susceptibility from Early-Life
5 Exposure to Carcinogens fU.S. EPA. 2005bl. either default or chemical-specific age-dependent
6 adjustment factors (ADAFs) are recommended to account for early-life exposure to carcinogens
7 that act through a mutagenic MOA. Because chemical-specific lifestage susceptibility data for cancer
8 are not available, and because the MOA for ETBE carcinogenicity is not known (see Section 1.3.2),
9 application of ADAFs is not recommended.
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Borghoff. ST: Parkinson. H: Leavens. TL. (2010). Physiologically based pharmacokinetic rat model
for methyl tertiary-butyl ether; comparison of selected dose metrics following various
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Borghoff. ST: Prescott. IS: Tanszen. DB: Wong. BA: Everitt. II. (2001). alpha2u-Globulin nephropathy,
renal cell proliferation, and dosimetry of inhaled tert-butyl alcohol in male and female F-
344 rats. Toxicol Sci. 61: 176-186. http://dx.doi.Org/10.1093/toxsci/61.l.176.
Borghoff. ST: Ring. C: Banton. MI: Leavens. TL. (2016). Physiologically based pharmacokinetic model
for ethyl tertiary-butyl ether and tertiary-butyl alcohol in rats: Contribution of binding to
a2u-globulin in male rats and high-exposure nonlinear kinetics to toxicity and cancer
outcomes. J Appl Toxicol. http://dx.doi.org/10.10Q2 /iat3412.
Brennan. P. (2002). Gene-environment interaction and aetiology of cancer: what does it mean and
how can we measure it? Carcinogenesis. 23: 381-387.
Brennan. P: Lewis. S: Hashibe. M: Bell. DA: Boffetta. P: Bouchardv. C: Caporaso. N: Chen. C: Coutelle.
C: Diehl. SR: Hayes. RB: Olshan. AF: Schwartz. SM: Sturgis. EM: Wei. 0: Zavras. AI:
Benhamou. S. (2004). Pooled analysis of alcohol dehydrogenase genotypes and head and
neck cancer: a HuGE review. Am J Epidemiol. 159: 1-16.
Cal/EPA (California Environmental Protection Agency). (2016). GeoTracker. Available online at
http://geotracker.waterboards.ca.gov/ (accessed
Cederbaum. AI: Cohen. G. (1980). Oxidative demethylation of t-butyl alcohol by rat liver
microsomes. Biochem Biophys Res Commun. 97: 730-736.
Chernoff. N: Rogers. TM: Turner. CT: Francis. BM. (1991). SIGNIFICANCE OF SUPERNUMERARY RIBS
IN RODENT DEVELOPMENTAL TOXICITY STUDIES - POSTNATAL PERSISTENCE IN RATS
AND MICE. Fundam Appl Toxicol. 17: 448-453. http://dx.doi.org/10.1016/0272-
0590C91190196-B.
Cohen. SM: Hard. GC: Regan. KS: Seelv. TC: Bruner. RH. (2011). Pathology working group review of
selected histopathologic changes in the kidneys of rats assigned to toxicology and
carcinogenicity studies of ethyl tertiary butyl ether (ETBE): Japan Bioassay Research Center
studies no.: 0065 and 0691 [Unpublished report] (pp. 1-30). Research Triangle Park, NC:
Research Pathology Associates under contract to Lyondell Chemical Company.
Craig. EA: Yan. Z: Zhao. 01. (2014). The relationship between chemical-induced kidney weight
increases and kidney histopathology in rats. J Appl Toxicol. 35: 729-736.
http://dx.doi.org/10.1002/iat3036.
de Pevster. A. (2010). Ethyl t-butyl ether: Review of reproductive and developmental toxicity
[Review], Birth Defects Res B Dev Reprod Toxicol. 89: 239-263.
http://dx.doi.org/10.1002/bdrb.20246.
de Pevster. A: Stanard. B: Westover. C. (2009). Effect ofETBE on reproductive steroids in male rats
and rat Leydig cell cultures. Toxicol Lett. 190: 74-80.
http://dx.doi.Org/10.1016/i.toxlet.2009.06.879.
This document is a draft for review purposes only and does not constitute Agency policy.
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Dekant. W: Bernauer. U: Rosner. E: Amberg. A. (2001). Toxicokinetics of ethers used as fuel
oxygenates [Review], Toxicol Lett. 124: 37-45. http://dx.doi.org/10.1016/s0378-
4274C00100284-8.
Doi. AM: Hill. G: Seelv. 1: Hailev. 1R: Kissling. G: Bucher. TR. (2007). a2u-Globulin Nephropathy and
Renal Tumors in National Toxicology Program Studies. Toxicol Pathol. 35: 533-540.
Dorman. DC: Struve. MF: Wong. BA: Morgan. KT: Tanszen. DB: Gross. EB: Bond. TA. (1997).
Neurotoxicological evaluation of ethyl tertiary-butyl ether following subchronic (90-day)
inhalation in the Fischer 344 rat J Appl Toxicol. 17: 235-242.
http://dx.doi.org/10.1002/fsici") 1099-12 63 f199707") 17:4<235::aid-iat435>3.0.co:2-4.
Drogos. PL: Diaz. AF. (2001). Oxygenates in Gasoline Appendix A: Physical properties of fuel
oxgenates and addititves. In ACS Symposium Series. Washington, DC: American Chemical
Society, http://dx.doi.org/10.1021/bk-2002-0799.ch018.
ECHA (European Chemicals Agency). (2016). Registered substances. 2-ethoxy-2-methylpropane. EC
number: 211-309-7. CAS number: 637-92-3. https://echa.europa.eu/registration-dossier/-
/registered-dossier/15520.
Elcombe. CR: Peffer. RC: Wolf. DC: Bailey. 1: Bars. R: Bell. D: Cattlev. RC: Ferguson. SS: Geter. D:
Goetz. A: Goodman. II: Hester. S: Tacobs. A: Omiecinski. CI: Schoenv. R: Xie. W: Lake. BG.
(2014). Mode of action and human relevance analysis for nuclear receptor-mediated liver
toxicity: A case study with phenobarbital as a model constitutive androstane receptor (CAR)
activator [Review], Crit Rev Toxicol. 44: 64-82.
http://dx.doi.org/10.3109/10408444.2013.835786.
Eriksson. CI. (2015). Genetic-epidemiological evidence for the role of acetaldehyde in cancers
related to alcohol drinking [Review], Adv Exp Med Biol. 815: 41-58.
http://dx.doi.org/10.1007/978-3-319-09614-8 3.
Faulkner. TP: Wiechart. ID: Hartman. DM: Hussain. AS. (1989). The effects of prenatal tertiary
butanol administration in CBA/J and C57BL/6J mice. Life Sci. 45: 1989-1995.
Fuiii. S: Yabe. K: Furukawa. M: Matsuura. M: Aovama. H. (2010). A one-generation reproductive
toxicity study of ethyl tertiary butyl ether in rats. Reprod Toxicol. 30: 414-421.
http://dx.doi.Org/10.1016/i.reprotox.2010.04.013.
Fukami. T: Nakaiima. M: Yoshida. R: Tsuchiva. Y: Fuiiki. Y: Katoh. M: Mcleod. HL: Yokoi. T. (2004). A
novel polymorphism of human CYP2A6 gene CYP2A6*17 has an amino acid substitution
(V365M) that decreases enzymatic activity in vitro and in vivo. Clin Pharmacol Ther. 76:
519-527. http://dx.doi.org/10.1016/i.clpt2004.08.014.
Gaoua. W. (2004a). Ethyl tertiary butyl ether (ETBE): Prenatal developmental toxicity study by the
oral route (gavage) in rats (pp. 1-280). (CIT Study No. 24860 RSR). unpublished study for
Totalfinaelf on behalf of the ETBE Producers' Consortium. An external peer review was
conducted by EPA in November 2008 to evaluate the accuracy of experimental procedures,
results, and interpretation and discussion of the findings presented. A report of this peer
review is available through EPA's IRIS Hotline, at (202) 566-1676 (phone), (202) 56-1749
(fax), or hotline.iris@epa.gov (e-mail address) and at www.epa.gov/iris.
Gaoua. W. (2004b). Ethyl tertiary butyl ether (ETBE): Two-generation study (reproduction and
fertility effects) by the oral route (gavage) in rats. (CIT Study No. 24859 RSR). unpublished
study for Totalfinaelf on behalf of the ETBE Producers' Consortium. An external peer review
was conducted by EPA in November 2008 to evaluate the accuracy of experimental
procedures, results, and interpretation and discussion of the findings presented. A report of
this peer review is available through EPA's IRIS Hotline, at (202) 566-1676 (phone), (202)
56-1749 (fax), or hotline.iris@epa.gov (e-mail address) and at www.epa.gov/iris.
This document is a draft for review purposes only and does not constitute Agency policy.
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Gehlhaus. MW. Ill: Gift. TS: Hogan. KA: Kopvlev. L: Schlosser. PM: Kadrv. A. -R. (2011). Approaches to
cancer assessment in EPA's Integrated Risk Information System [Review], Toxicol Appl
Pharmacol. 254: 170-180. http://dx.doi.Org/10.1016/i.taap.2010.10.019.
Guvton. KZ: Chiu. WA: Bateson. TF: linot. I: Scott. CS: Brown. RC: Caldwell. TC. (2009). A
reexamination of the PPAR-alpha activation mode of action as a basis for assessing human
cancer risks of environmental contaminants [Review], Environ Health Perspect. 117: 1664-
1672. http: //dx.doi.org/10.1289 /ehp.0900758.
Hagiwara. A: Doi. Y: Imai. N: Nakashima. H: Ono. T: Kawabe. M: Furukawa. F: Tamano. S: Nagano. K:
Fukushima. S. (2011). Medium-term multi-organ carcinogenesis bioassay of ethyl tertiary-
butyl ether in rats. Toxicology. 289: 160-166. http://dx.doi.Org/10.1016/i.tox.2011.08.007.
Hagiwara. A; Doi. Y; Imai. N; Suguro, M; Kawabe. M; Furukawa. F; Tamano. S; Nagano. K; Fukushima.
S. (2015). Promotion of liver and kidney carcinogenesis by ethyl tertiary-butyl ether (ETBE)
in male Wistar rats. J Toxicol Pathol. 28: 189-195. http://doi.org/10.1293/tox.2015-0023.
Hagiwara. A; Imai. N; Doi. Y; Suguro. M; Kawabe. M; Furukawa. F; Nagano. K; Fukushima. S. (2013).
No Promoting Effect of Ethyl Tertiary-butyl Ether (ETBE) on Rat Urinary Bladder
Carcinogenesis Initiated with N-Butyl-N-(4-hydroxybutyl)nitrosamine. J Toxicol Pathol. 26:
351-357. http://dx.doi.org/10.1293/tox.2013-0027.
Hard. GC: Banton. MI: Bretzlaff. RS: Dekant. W: Fowles. 1. R.: Mallett. AK: Mcgregor. DB: Roberts. KM:
Sielken. RL: Valdez-Flores. C: Cohen. SM. (2013). Consideration of rat chronic progressive
nephropathy in regulatory evaluations for carcinogenicity. Toxicol Sci. 132: 268-275.
http: / /dx. doi. o r g/10.109 3 /toxsci /kfs305.
Hard. GC: Tohnson. KT: Cohen. SM. (2009). A comparison of rat chronic progressive nephropathy
with human renal disease-implications for human risk assessment [Review], CritRev
Toxicol. 39: 332-346. http://dx.doi.org/10.1080/10408440802368642.
Hattis. D: Goble. R: Russ. A: Chu. M: Ericson. I. (2004). Age-related differences in susceptibility to
carcinogenesis: A quantitative analysis of empirical animal bioassay data. Environ Health
Perspect 112: 1152-1158. http://dx.doi.org/10.1289/ehp.6871.
HSDB (Hazardous Substances Database). (2012). Ethyl tert-butyl ether - CASRN: 637-92-3.
Washington D.C.. https://toxnet.nlm.nih.gov/cgi-
bin/sis/search/a?dbs+hsdb:@term+@D0CN0+7867.
IARC (International Agency for Research on Cancer). (1999a). Acetaldehyde [IARC Monograph] (pp.
319-335). Lyon, France. http://monographs.iarc.fr/ENG/Monographs/vol71/mono71-
88.pdf.
IARC (International Agency for Research on Cancer). (1999b). IARC monographs on the evaluation
of carcinogenic risks to humans: Re-evaluation of some organic chemicals, hydrazine and
hydrogen peroxide [IARC Monograph], Lyon, France: World Health Organization.
IARC (International Agency for Research on Cancer). (1999c). Methyl tert-butyl ether (group 3) (pp.
339-383). Lyon, France.
IARC (International Agency for Research on Cancer). (1999d). Some chemicals that cause tumours
of the kidney or urinary bladder in rodents and some other substances: Methyl tert-butyl
ether. In IARC Monographs on the Evaluation of Carcinogenic Risks to Humans. Lyon,
France: World Health Organization.
IARC (International Agency for Research on Cancer). (2010). Alcohol consumption and ethyl
carbamate [IARC Monograph], Lyon, France.
http://monographs.iarc.fr/ENG/Monographs/vol96/mono96.pdf.
This document is a draft for review purposes only and does not constitute Agency policy.
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IARC (International Agency for Research on Cancer). (2012). Consumption of Alcoholic Beverages
[IARC Monograph], Lyon, France. http://monographs.iarc.fr/ENG/Monographs/vollOOE/.
IARC (International Agency for Research on Cancer). (2015). Preamble to the IARC monographs.
Lyon, France: International Agency for Research on Cancer, World Health Organization.
http://monographs.iarc.fr/ENG/Preamble/.
Tohanson. G: Nihlen. A: Lof. A. (1995). Toxicokinetics and acute effects of MTBE and ETBE in male
volunteers. Toxicol Lett. 82/83: 713-718. http://dx.doi.org/10.1016/0378-4274C95103589-
3.
TPEC (Japan Petroleum Energy Center). (2007a). 13-Week toxicity test of 2-ethoxy-2-
methylpropane in F344 rats (drinking water study) [Preliminary study for the
carcinogenicity test], (Study No. 0665). Kanagawa, Japan: Japan Industrial Safety and
Health.
TPEC (Japan Petroleum Energy Center). (2007b). Micronucleus test of ETBE using bone marrow of
rats of the "13-week toxicity study of 2-ethoxy-2-methylpropane in F344 rats (drinking
water study) [preliminary carcinogenicity study]". (Study Number: 7046). Japan Bioassay
Research Center, Japan Industrial Safety and Health Association.
TPEC (Japan Petroleum Energy Center). (2008a). [28-day ETBE repeated dose full-body inhalation
toxicity test in rats (preliminary test)]. (Study No. B061828). Japan: Mitsubishi Chemical
Safety Institute Ltd.
TPEC (Japan Petroleum Energy Center). (2008b). A 90-day repeated dose toxicity study ofETBE by
whole-body inhalation exposure in rats. (Study Number: B061829). Mitsubishi Chemical
Safety Institute Ltd. An external peer review was conducted by EPA in August 2012 to
evaluate the accuracy of experimental procedures, results, and interpretation and
discussion of the findings presented. A report of this peer review is available through EPA's
IRIS Hotline, at (202) 566-1676 (phone), (202) 56-1749 (fax), or hotline.iris@epa.gov (e-
mail address) and at www.epa.gov/iris.
TPEC (Japan Petroleum Energy Center). (2008c). A 180-Day repeated dose oral toxicity study of
ETBE in rats. (Study Number: D19-0002). Japan: Hita Laboratory, Chemicals Evaluation and
Research Institute (CERI). An external peer review was conducted by EPA in August 2012 to
evaluate the accuracy of experimental procedures, results, and interpretation and
discussion of the findings presented. A report of this peer review is available through EPA's
IRIS Hotline, at (202) 566-1676 (phone), (202) 56-1749 (fax), or hotline.iris@epa.gov (e-
mail address) and at www.epa.gov/iris.
TPEC (Japan Petroleum Energy Center). (2008d). Medium-term mutli-organ carcinogenesis bioassay
of 2-ethoxy-2-methylpropane (ETBE) in rats. (Study Number: 0635). Ichinomiya, Japan:
DIMS Institute of Medical Science.
TPEC (Japan Petroleum Energy Center). (2008e). A one-generation reproduction toxicity study of
ETBE in rats. (Study Number: SR07060). Safety Research Institute for Chemical Compounds.
TPEC (Japan Petroleum Energy Center). (2008f). Pharmacokinetic study in rats treated with [14c]
ETBE repeatedly for 14 days. (P070497). Japan: Kumamoto Laboratory, Mitsubishi
Chemical Safety Institute Ltd. An external peer review was conducted by EPA in August
2012 to evaluate the accuracy of experimental procedures, results, and interpretation and
discussion of the findings presented. A report of this peer review is available through EPA's
IRIS Hotline, at (202) 566-1676 (phone), (202) 56-1749 (fax), or hotline.iris@epa.gov (e-
mail address) and at www.epa.gov/iris.
TPEC (Japan Petroleum Energy Center). (2008g). Pharmacokinetic study in rats treated with single
dose of [14C] ETBE. (P070496). Japan: Kumamoto Laboratory, Mitsubishi Chemical Safety
This document is a draft for review purposes only and does not constitute Agency policy.
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Institute Ltd. An external peer review was conducted by EPA in August 2012 to evaluate the
accuracy of experimental procedures, results, and interpretation and discussion of the
findings presented. A report of this peer review is available through EPA's IRIS Hotline, at
(202) 566-1676 (phone), (202) 56-1749 (fax), or hotline.iris@epa.gov (e-mail address) and
at www.epa.gov/iris.
TPEC (Japan Petroleum Energy Center). (2008h). A prenatal developmental toxicity study ofETBE in
rats. (Study Code Number E09-0006). Hita Research Laboratories, Chemicals Evaluation
and Research Institute (CERI).
TPEC (Japan Petroleum Energy Center). (2008i). Study for effects on embryo-fetal development in
rabbits treated orally with ETBE. (Study No. R-965). Shizuoka, Japan: Kannami Laboratory,
Bozo Research Center Inc.
TPEC (Japan Petroleum Energy Center). (2010a). Carcinogenicity test of 2-Ethoxy-2-methylpropane
in rats (Drinking water study). (Study No: 0691). Japan Industrial Safety and Health
Association, Japan Bioassay Research Center. An external peer review was conducted by
EPA in August 2012 to evaluate the accuracy of experimental procedures, results, and
interpretation and discussion of the findings presented. A report of this peer review is
available through EPA's IRIS Hotline, at (202) 566-1676 (phone), (202) 56-1749 (fax), or
hotline.iris@epa.gov (e-mail address) and at www.epa.gov/iris.
TPEC (Japan Petroleum Energy Center). (2010b). Carcinogenicity test of 2-Ethoxy-2-methylpropane
in rats (Inhalation study). (Study No: 0686). Japan: Japan Industrial Safety and Health
Association.. An external peer review was conducted by EPA in August 2012 to evaluate the
accuracy of experimental procedures, results, and interpretation and discussion of the
findings presented. A report of this peer review is available through EPA's IRIS Hotline, at
(202) 566-1676 (phone), (202) 56-1749 (fax), or hotline.iris@epa.gov (e-mail address) and
at www.epa.gov/iris.
Kakehashi. A: Hagiwara. A: Imai. N: Nagano. K: Nishimaki. F: Banton. M: Wei. M: Fukushima. S:
Wanibuchi. H. (2013). Mode of action of ethyl tertiary-butyl ether hepatotumorigenicity in
the rat: evidence for a role of oxidative stress via activation of CAR, PXR and PPAR signaling
pathways. Toxicol Appl Pharmacol. 273: 390-400.
http://dx.doi.Org/10.1016/i.taap.2013.09.016.
Kakehashi. A: Hagiwara. A: Imai. N: Wei. M: Fukushima. S: Wanibuchi. H. (2015). Induction of cell
proliferation in the rat liver by the short-term administration of ethyl tertiary-butyl ether. J
Toxicol Pathol. 28: 27-32. http://dx.doi.org/10.1293/tox.2014-0056.
Kaneko. T: Wang. P. -Y: Sato. A. (2000). Partition coefficients for gasoline additives and their
metabolites. J Occup Health. 42: 86-87. http://dx.doi.org/10.1539/ioh.42.86.
Kanerva. RL: Ridder. GM: Stone. LC: Alden. CL. (1987). Characterization of spontaneous and decalin-
induced hyaline droplets in kidneys of adult male rats. Food Chem Toxicol. 25: 63-82.
Klaunig. IE: Babich. MA: Baetcke. KP: Cook. TC: Corton. TC: David. RM: Deluca. TG: Lai. DY: Mckee. RH:
Peters. TM: Roberts. RA: Fenner-Crisp. PA. (2003). PPARalpha agonist-induced rodent
tumors: Modes of action andhuman relevance [Review], Crit Rev Toxicol. 33: 655-780.
http://dx.doi.org/10.1080 /713608372.
Leavens. TL: Borghoff. ST. (2009). Physiologically based pharmacokinetic model of methyl tertiary
butyl ether and tertiary butyl alcohol dosimetry in male rats based on binding to alpha2u-
globulin. Toxicol Sci. 109: 321-335. http://dx.doi.org/10.1093/toxsci/kfp049.
Li. 0: Kobavashi. M: Inagaki. H: Hirata. Y: Hirata. K: Shimizu. T: Wang. R. -S: Suda. M: Kawamoto. T:
Nakaiima. T: Kawada. T. (2011). Effects of subchronic inhalation exposure to ethyl tertiary
This document is a draft for review purposes only and does not constitute Agency policy.
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butyl ether on splenocytes in mice. Int J Immunopathol Pharmacol. 24: 837-847.
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Liau. SS: Oureshi. MS: Praseedom. R: Huguet. E. (2013). Molecular pathogenesis of hepatic
adenomas and its implications for surgical management [Review], 17: 1869-1882.
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Malarkev. DE: Bucher. TR. (2011). Summary report of the National Toxicology Program and
Environmental Protection Agency-sponsored review of pathology materials from selected
Ramazzini Institute rodent cancer bioassays [NTP], Research Triangle Park: National
T oxicology Program.
http://ntp.niehs.nih.gov/ntp/about ntp/partnerships/international/summarypwg report
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project of carcinogenicity bioassays on gasoline oxygenated additives: plan and first report
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Available online at
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37: 287-312. http://dx.doi.org/10.1080/10408440601177723.
Medinskv. MA: Wolf. DC: Cattlev. RC: Wong. B: Tanszen. DB: Farris. GM: Wright. GA: Bond. TA. (1999).
Effects of a thirteen-week inhalation exposure to ethyl tertiary butyl ether on Fischer-344
rats and CD-I mice. Toxicol Sci. 51: 108-118. http://dx.doi.Org/10.1093/toxsci/51.l.108.
Melnick. R: Burns. K: Ward. I: Huff. I. (2012). Chemically exacerbated chronic progressive
nephropathy not associated with renal tubule tumor induction in rats: An evaluation based
on 60 carcinogenicity studies by the National Toxicology Program. Toxicol Sci. 128: 346-
356. http://dx.doi.org/10.1093/toxsci/kfsl56.
Mivata. K: Koga. T: Aso. S: Hoshuvama. S: Aiimi. S: Furukawa. K. (2013). A subchronic (180-day) oral
toxicity study of ethyl tertiary-butyl ether, a bioethanol, in rats. Drug Chem Toxicol.
http: //dx.doi.org/10.3109 /01480545.2013.851690.
Montgomery. CR. (1994). n-Octanol/water partition co-efficient (Kow) for ethyl tertiary butyl ether
(ETBE) at 25 degrees C +/-1 degree C. Montgomery, CR.
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on a Survey of State Experiences with MtBE and Other Oxygenate Contamination at LUST
Sites. Lowell, Massachusetts: NEIWPCC.
http://www.neiwpcc.org/neiwpcc docs/2003mtbesum.pdf.
Nihlen. A: Tohanson. G. (1999). Physiologically based toxicokinetic modeling of inhaled ethyl
tertiary-butyl ether in humans. Toxicol Sci. 51: 184-194.
http://dx.doi.Org/10.1093/toxsci/51.2.184.
Nihlen. A: Lof. A: Tohanson. G. (1995). Liquid/air partition coefficients of methyl and ethyl t-butyl
ethers, t-amyl methyl ether, and t-butyl alcohol. J Expo Anal Environ Epidemiol. 5: 573-582.
This document is a draft for review purposes only and does not constitute Agency policy.
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Nihlen. A: Lof. A: Tohanson. G. (1998a). Controlled ethyl tert-butyl ether (ETBE) exposure of male
volunteers: I Toxicokenetics. Toxicol Sci. 46: 1-10.
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Nihlen. A: Lof. A: Tohanson. G. (1998b). Controlled ethyl tert-butyl ether (ETBE) exposure of male
volunteers: II. Acute effects. Toxicol Sci. 46: 143-150.
http://dx.doi.org/10.1006/toxs.1998.2517.
Nihlen. A: Lof. A: Tohanson. G. (1998c). Experimental exposure to methyl tertiary-butyl ether: I
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This document is a draft for review purposes only and does not constitute Agency policy.
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This document is a draft for review purposes only and does not constitute Agency policy.
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This document is a draft for review purposes only and does not constitute Agency policy.
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