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EPA /635/R-20/400Fa
www.epa.gov/iris
Toxicological Review of Ethyl Tertiary Butyl Ether
(CASRN 637-92-3]
August 2021
Integrated Risk Information System
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
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Toxicological Review of Ethyl Tertiary Butyl Ether
DISCLAIMER
This document has been reviewed by the U.S. Environmental Protection Agency, Office of
Research and Development, and approved for publication. Any mention of trade names, products,
or services does not imply an endorsement by the U.S. Government or the U.S. Environmental
Protection Agency. The EPA does not endorse any commercial products, services, or enterprises.
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Toxicological Review of Ethyl Tertiary Butyl Ether
CONTENTS
AUTHORS | CONTRIBUTORS | REVIEWERS ix
PREFACE xiii
PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS xvii
EXECUTIVE SUMMARY xxv
LITERATURE SEARCH STRATEGY | STUDY SELECTION AND EVALUATION xxxii
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 1-1
1.1.3. Description of Toxicokinetic Models 1-3
1.1.4. Related Chemicals That Provide Supporting Information 1-4
1.2. PRESENTATION AND SYNTHESIS OF EVIDENCE BY ORGAN/SYSTEM 1-4
1.2.1. Kidney Effects 1-4
1.2.2. Liver Effects 1-41
1.2.3. Reproductive Effects 1-63
1.2.4. Developmental Effects 1-95
1.2.5. Carcinogenicity (Other Than in the Kidney or Liver) 1-108
1.2.6. Other Toxicological Effects 1-117
1.3. INTEGRATION AND EVALUATION 1-118
1.3.1. Effects Other Than Cancer 1-118
1.3.2. Carcinogenicity 1-119
1.3.3. Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes 1-122
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
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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
2.4. INHALATION UNIT RISK FOR CANCER 2-20
2.4.1. Analysis of Carcinogenicity Data 2-21
2.4.2. Dose-Response Analysis—Adjustments and Extrapolation Methods 2-21
2.4.3. Inhalation Unit Risk Derivation 2-23
2.4.4. Uncertainties in the Derivation of the Inhalation Unit Risk 2-24
2.4.5. Previous IRIS Assessment 2-25
2.5. APPLICATION OF AGE-DEPENDENT ADJUSTMENT FACTORS 2-25
REFERENCES R-l
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TABLES
Table ES-1. Organ/system-specific RfDs and overall RfD for ETBE xxvii
Table ES-2. Organ/system-specific RfCs and overall RfCfor ETBE xxviii
Table LS-1. Details of the search strategy employed for ETBE xxxvi
Table LS-2. Summary of additional search strategies for ETBE xxxvi
Table LS-3. Inclusion-exclusion criteria xxxvii
Table LS-4. Considerations for evaluating experimental animal studies xl
Table LS-5. Summary of experimental animal evidence base xli
Table 1-1. Chemical identity and physicochemical properties of ETBE from EPA's CompTox
Chemicals Dashboard 1-2
Table 1-2. Evidence pertaining to kidney histopathology effects in animals following exposure to
ETBE 1-11
Table 1-3. Evidence pertaining to kidney biochemistry and urine effects in animals following
exposure to ETBE 1-14
Table 1-4. Evidence pertaining to kidney tumor effects in animals following exposure to ETBE 1-18
Table 1-5. Comparison of nephropathy and urothelial hyperplasia in individual male rats from
2-year oral exposure {JPEC, 2010, 1517477} 1-19
Table 1-6. Comparison of nephropathy and urothelial hyperplasia in individual male rats from
2-year inhalation exposure {JPEC, 2010, 1517421} 1-20
Table 1-7. Additional kidney effects potentially relevant to mode of action in animals exposed to
ETBE 1-26
Table 1-8. Summary of data informing whether the alpha 2u-globulin process is occurring in
male rats exposed to ETBE 1-29
Table 1-9. IARC criteria for an agent causing kidney tumors through an alpha 2u-globulin
associated response in male rats 1-36
Table 1-10. Evidence pertaining to liver weight effects in animals exposed to ETBE 1-44
Table 1-11. Evidence pertaining to liver histopathology effects in animals exposed to ETBE 1-46
Table 1-12. Evidence pertaining to liver biochemistry effects in animals exposed to ETBE 1-50
Table 1-13. Evidence pertaining to liver tumor effects in animals exposed to ETBE 1-55
Table 1-14. Evidence of key characteristics of carcinogens for ETBE 1-58
Table 1-15. Evidence pertaining to male reproductive effects in animals exposed to ETBE 1-66
Table 1-16. Evidence pertaining to female reproductive effects in animals exposed to ETBE 1-85
Table 1-17. Evidence pertaining to developmental effects in animals following exposure to ETBE 1-98
Table 1-18. Evidence pertaining to carcinogenic effects (in tissues other than liver or kidney) in
animals exposed to ETBE 1-111
Table 1-19. Supplemental evidence pertaining to ETBE promotion of mutagen-initiated tumors
in animals 1-113
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 2-7
Table 2-3. Organ/system-specific RfDs and overall RfD for ETBE 2-9
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 inhalation unit risk derivation 2-23
Table 2-8. Summary of uncertainties in the derivation of the inhalation unit risk for ETBE 2-24
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FIGURES
Figure LS-1. Summary of literature search and screening process for ETBE xxxv
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 metabolite internal blood concentration 1-9
Figure 1-3. Comparison of absolute kidney-weight change in male and female mice following
13-week inhalation exposure 1-10
Figure 1-4. Exposure-response array of kidney effects following oral exposure to ETBE 1-21
Figure 1-5. Exposure-response array of kidney effects following inhalation exposure to ETBE 1-22
Figure 1-6. Temporal pathogenesis of alpha 2u-globulin-associated nephropathy in male rats 1-25
Figure 1-7. ETBE oral exposure array of alpha 2u-globulin data in male rats 1-31
Figure 1-8. ETBE inhalation exposure array of alpha 2u-globulin data in male rats 1-32
Figure 1-9. Exposure-response array of noncancer liver effects following oral exposure to ETBE 1-53
Figure 1-10. Exposure-response array of noncancer liver effects following inhalation exposure to
ETBE 1-54
Figure 1-11. Exposure-response array of male reproductive effects following oral exposure to
ETBE 1-81
Figure 1-12. Exposure-response array of male reproductive effects following inhalation
exposure to ETBE 1-82
Figure 1-13. Exposure-response array of female reproductive effects following oral exposure to
ETBE 1-93
Figure 1-14. Exposure-response array of female reproductive effects following inhalation
exposure to ETBE 1-94
Figure 1-15. Exposure-response array of developmental effects following oral exposure to ETBE 1-107
Figure 1-16. Exposure-response array of carcinogenic effects following oral exposure to ETBE 1-116
Figure 1-17. Exposure-response array of carcinogenic effects following inhalation exposure to
ETBE 1-117
Figure 2-1. Oral candidate values with corresponding POD and composite UF 2-8
Figure 2-2. Inhalation candidate values with corresponding POD and composite UF 2-17
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ABBREVIATIONS
ADAF
age-dependent adjustment factor
HEC
human equivalent concentration
ADJ
adjusting experimental exposure
HED
human equivalent dose
concentrations to a value reflecting
HERO
Health and Environment Research on
continuous exposure duration
Online
ADME
absorption, distribution, metabolism,
HGPRT
hypoxanthine-guanine
excretion
phosphoribosyltransferase
AIC
Akaike's information criterion
HIBA
2-hydroxyisobutyrate
ALDH
aldehyde dehydrogenase
HT
heterogeneous
ALP
alkaline phosphatase
IARC
International Agency for Research on
ALT
alanine
Cancer
aminotransferase/transaminase
IRIS
Integrated Risk Information System
AST
aspartate
i.v.
intravenous
aminotransferase/transaminase
JPEC
Japan Petroleum Energy Center
atm
atmosphere
Km
Michaelis-Menten constant
ATSDR
Agency for Toxic Substances and
KO
knockout
Disease Registry
LD
lactation day
AUC
area-under-the-curve
LOAEL
lowest-observed-adverse-effect level
BBN
/V-butyl-/V-(hydroxybutyl]nitrosamine
MN
micronucleus, micronucleated
BMC
benchmark concentration
MNPCE
micronucleated polychromatic
BMCL
benchmark concentration lower
erythrocytes
confidence limit
MNU
iV-methyl-iV-nitrosourea
BMD
benchmark dose
MOA
mode of action
BMDL
benchmark dose lower confidence limit
MPD
2 -methyl-1,2 -propanediol
BMDS
Benchmark Dose Software
MTBE
methyl tert-butyl ether
BMR
benchmark response
MTD
maximum tolerated dose
BUN
blood urea nitrogen
N.D.
not detected
BW
body weight
No.
number
CAAC
Chemical Assessment Advisory
NOAEL
no-observed-adverse-effect level
Committee
NR
not reported
CAR
constitutive androstane receptor
NTP
National Toxicology Program
CASRN
Chemical Abstracts Service registry
OECD
Organisation for Economic
number
Co-operation and Development
Cmax
maximum concentration
ORD
Office of Research and Development
CPHEA
Center for Public Health and
OSF
oral slope factor
Environmental Assessment
PBPK
physiologically based pharmacokinetic
CPN
chronic progressive nephropathy
PCE
polychromatic erythrocyte
CSL
continuous simulation language
PND
postnatal day
CYP450
cytochrome P450
PNW
postnatal week
DAF
dosimetric adjustment factor
POD
point of departure
DEN
diethylnitrosamine
PPARa
peroxisome proliferator-activated
df
degrees of freedom
receptor a
DHPN
N- bis (2-hydroxypropyl]nitrosamine
PXR
pregnane X receptor
DMH
1,2-dimethylhydrazine dihydrochloride
QA
quality assurance
DNA
deoxyribonucleic acid
QSAR
quantitative structure-activity
EHEN
N- ethyl -iV-hy dr oxy ethylni tr o s amin e
relationship
EPA
Environmental Protection Agency
RD
relative deviation
ETBE
ethyl tertiary butyl ether
RfC
inhalation reference concentration
GD
gestation day
RfD
oral reference dose
GGT
y-glutamyl transferase
rho
Spearman's rank coefficient
GLP
good laboratory practice
RNA
ribonucleic acid
HBA
hydroxyisobutyric acid
S-D
Sprague-Dawley
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SAB
Science Advisory Board
UFc
composite uncertainty factor
SD
standard deviation
UFd
database deficiencies uncertainty factor
SE
standard error
UFh
human variation uncertainty factor
SRBC
sheep red blood cell
UFl
LOAEL-to-NOAEL uncertainty factor
SS IICA
Stoddard Solvent IICA
UFs
subchronic-to-chronic uncertainty
TBA
tert-butyl alcohol, tert-butanol
factor
TSCATS
Toxic Substances Control Act Test
USGS
U.S. Geological Survey
Submissions
Vmax
maximum substrate turnover velocity
TWA
time-weighted average
voc
volatile organic compounds
UF
uncertainty factor
WT
wild type
UFa
animal-to-human uncertainty factor
wt
weight
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Assessment Team
Kathleen Newhouse, M.S.
(Chemical Manager)
Andre Weaver, Ph.D.
(Co-Chemical Manager)
Janice Lee, Ph.D.
U.S. EPA
Office of Research and Development
Center for Public Health and Environmental Assessment
Keith Salazar, Ph.D.
Former Chemical Manager
EPA/ORD/NCEA
Currently with U.S. EPA, Office of Chemical Safety and
Pollution Prevention, Office of Pollution Prevention and
Toxics
Christopher Brinkerhoff, Ph.D.
Former ORISE Postdoctoral Fellow at U.S. EPA/ORD/NCEA
Currently with U.S. EPA, Office of the Administrator, Office of
Children's Health Protection
Contributors
Brandy Beverly, Ph.D.
Christine Cai, M.S.
Jeff Gift, Ph.D.
Karen Hogan, M.S.
Andrew Hotchkiss, Ph.D.
Channa Keshava, Ph.D.
Susan Makris, Ph.D.
Alan Sasso, Ph.D.
Paul Schlosser, Ph.D.
Erin Yost, Ph.D.
Vincent Cogliano, Ph.D.
Jason Fritz, Ph.D.
Charles Wood, Ph.D.
U.S. EPA
Office of Research and Development
Center for Public Health and Environmental Assessment
(previously with) U.S. EPA National Center for Environmental
Assessment
(previously with) U.S. EPA National Health and
Environmental Effects Research Lab
Production Team
Maureen Johnson
Dahnish Shams
Vicki Soto
U.S. EPA
Office of Research and Development
Center for Public Health and Environmental Assessment
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Contractor Support
Robyn Blain, Ph.D.
Ami Gordon, M.P.H.
Pam Ross, M.S.P.H.
ICF
9300 Lee Highway
Fairfax, VA
Executive Direction
Wayne E. Cascio, M.D. (CPHEA Director) U.S. EPA
Samantha Jones, Ph.D. (CPHEA Office of Research and Development
Associate Director) Center for Public Health and Environmental Assessment
Andrew Kraft, Ph.D. (CPAD Senior
Science Advisor)
Ravi Subramaniam, Ph.D. (CPAD
Branch Chief)
Kristina Thayer, Ph.D. (CPAD Director)
Paul White, Ph.D. (CPAD Senior
Science Advisor)
Internal Review Team
General Toxicology Workgroup U.S. EPA
Inhalation Workgroup Office of Research and Development
Neurotoxicity Workgroup Center for Public Health and Environmental Assessment
PharmacoKinetic Workgroup
Reproductive and Developmental
Toxicology Workgroup
Statistical Workgroup
Toxicity Pathways Workgroup
Executive Review Committee
Reviewers
This assessment was provided for review to scientists in EPA's Program and Regional Offices.
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, NY
Region 8, Denver, CO
This assessment was provided for review to other federal agencies and the Executive Office of the
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
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This assessment was released for public comment on September 1, 2016 for a 60-day public
comment period. The public comments are available on Regulations.gov. A summary and EPA's
disposition of the comments from the public is available in the revised external review draft
assessment on the IRIS website. Comments were received from the following entities:
American Chemistry Council
American Petroleum Institute
Exponent, Inc. on behalf of LyondellBasell
Japan Petroleum Energy Center
LyondellBasell
Tox Strategies on behalf of LyondellBasell
This assessment was peer reviewed by independent, expert scientists external to EPA convened by
EPA's Science Advisory Board (SAB). A peer-review meeting was held on August 15-17, 2017. The
report of the SAB's review of the EPA's Draft Toxicological Review of ETBE, dated February 27,
2019, is available on the SAB website. A summary and EPA's disposition of the comments received
from the SAB is included in Appendix E.
Janice E. Chambers, Ph.D. (chair)
Hugh A. Barton, Ph.D.
Janet Benson, Ph.D.
Trish Berger, Ph.D.
John Budroe, Ph.D.
James V. Bruckner, Ph.D.
Karen Chou, Ph.D.
Harvey Clewell, Ph.D.
Deborah Cory-Slechta, Ph.D.
Bevin Engelward, Ph.D.
Jeffrey Fisher, Ph.D.
William Michael Foster, Ph.D.
Alan Hoberman, Ph.D.
Tamarra James-Todd, Ph.D.
Lawrence Lash, Ph.D.
Marvin Meistrich, Ph.D.
Maria Morandi, Ph.D.
Isaac Pessah, Ph.D.
Lorenz Rhomberg, Ph.D.
Stephen M. Roberts, Ph.D.
Alan Stern, Ph.D.
Mississippi State University, MS
Pfizer, Inc., Groton, CT
Lovelace Biomedical, Albuquerque, NM
University of California, Davis, Davis, CA
Office of Environmental Health Hazard Assessment, Oakland, CA
University of Georgia, Athens, GA
Michigan State University, East Lansing, MI
Ramboll Environment and Health, Research Triangle Park, NC
University of Rochester, Rochester, NY
Massachusetts Institute of Technology, Cambridge, MA
U.S. Food and Drug Administration, Jefferson, AR
Independent Consultant, Durham, NC
Charles River Laboratories, Inc., Horsham, PA
Harvard University, Boston, MA
Wayne State University, Detroit, MI
Anderson Cancer Center, University of Texas, Houston, TX
Independent Consultant, Houston, TX
University of California, Davis, CA
Gradient, Cambridge, MA
University of Florida, Gainesville, FL
New Jersey Department of Environmental Protection/
University of Medicine and Dentistry of New Jersey-Robert
Wood Johnson Medical School, Trenton, NJ
The postexternal review draft of the assessment was provided for review to scientists in EPA's
Program and Regional Offices, and to other federal agencies and the Executive Office of the
President (EOP). A summary and EPA's disposition of major comments from the other federal
agencies and EOP is available on the IRIS website. Comments were submitted by:
EPA, Office of Air Quality Planning and Standards
EPA, Office of Water
EPA, Office of Children's Health Protection
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EPA, Office of Land and Emergency Management
EPA, Region 2
Department of Health and Human Services/The National Institute for Occupational Safety
and Health (NIOSH)
Executive Office of the President/Office of Management and Budget
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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 a noncancer oral reference dose (RfD), a noncancer inhalation 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.
ETBE was named in the Energy Policy Act of 2005 (Section 1505). This law directs the
Administrator to conduct an assessment of ETBE along with other alkylates, ethers, and heavy
alcohols. ETBE was nominated for IRIS assessment by the Office of Transportation and Air Quality
(OTAQ) within the Office of Air and Radiation (OAR). Assessment of ETBE is also relevant to the
Office of Land and Emergency Management/Office of Underground Storage Tanks (OLEM/OUST)
and the Office of Water (OW) due to concerns over groundwater contamination from leaking
storage containers.
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 of ETBE are likely 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 alpha 2u-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 fBorghoffetal.. 20161.
Before developing the IRIS assessment, EPA held a public meeting 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 oftert-Butyl
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Alcohol (tert-ButanoI) 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-2013-11111. In October 2016, a public science meeting was held to give the public an
opportunity to engage in early discussions on the draft IRIS Toxicological Review of ETBE and the
draft charge to the peer-review panel prior to release for external peer review. The complete set of
public comments, including the slides, is available on the docket at http: //www.regulations.gov
(Docket ID No. EPA-HQ-QRD-2009-02291.
Organ/system-specific reference values are calculated where feasible (in this case for
kidney toxicity only). 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 about 1990 to 2006, ETBE was periodically
added to gasoline at levels up to approximately 2 0%, but methyl tert-butyl ether (MTBE) and other
oxygenates were more commonly used. In 2006, use of ETBE and other ether fuel additives ceased
in the United States, and the use of ethanol as an additive increased dramatically fWeaver etal..
2010). ETBE is still registered with EPA for use as a fuel additive, but it is not used currently in the
United States. Several states have banned or limited the use of ether fuel additives, largely in
response to groundwater contamination concerns.
The United States is a major exporter of ETBE, producing 25% of the world's ETBE in 2012.
Worldwide consumption of ETBE 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 fUSDA. 20121.
Fate and Transport
ETBE is expected to be highly mobile in soil due to its high carbon-water partitioning
coefficient (HSDB. 2012). ETBE is not predicted to adsorb onto suspended particles and is unlikely
to undergo anaerobic biodegradation in water (HSDB. 2012). ETBE is estimated to have a half-life
of 2 days in air fHSDB. 20121.
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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 the water percolating
through soils. 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.
20061 measuring volatile organic compounds (VOCs) in general, show lower detection rates of
ETBE. The 2006 USGS study showed detections of ETBE 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, did not focus on 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. California maintains an
online database of measurements from contaminated sites (CalEPA. 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 of ETBE in other states was found using fewer and less standardized data.
Currently, only 13 states routinely analyze for ETBE at fuel-contaminated sites (NEIWPCC. 2003).
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 of ETBE include inhalation and, to a lesser extent, dermal
contact. ETBE inhalation exposure can occur owing to the chemical's volatility and release from
industrial processes and contaminated sites (HSDB. 2012).
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; Tiesiema and Baars f20091]. The results of this
assessment are presented in Appendix A of the Supplemental Information to this Toxicological
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Review. Of importance to recognize is that earlier assessments were prepared for different
purposes and used different methods. In addition, more recent studies are included and evaluated
in this IRIS assessment.
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PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS
Note: The Preamble summarizes the objectives at
systematic review procedures used in developing
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 ozone,
carbon monoxide, sulfur oxides, nitrogen
oxides, and lead).
Enhancements to the IRIS program are
improving its science, transparency, and
productivity. To improve the science, the IRIS
ope of the IRIS program, general principles and
assessments¦, and the overall development
program is adapting and implementing
principles of systematic review (i.e., using
explicit methods to identify, evaluate, and
synthesize study findings). To increase
transparency, the IRIS program discusses key
science issues with the scientific community
and the public as it begins an assessment
External peer review, independently managed
and in public, improves both science and
transparency. Increased productivity requires
that assessments be concise, focused on EPA's
needs, and completed without undue delay.
IRIS assessments follow EPA guidance2
and standardized practices of systematic
review. This Preamble summarizes and does
not change IRIS operating procedures or EPA
guidance.
Periodically, the IRIS program asks for
nomination of agents for future assessment or
reassessment. Selection depends on EPA's
priorities, relevance to public health, and
availability of pertinent studies. The IRIS
multiyear agenda3 lists upcoming
assessments. The IRIS program may also
assess other agents in anticipation of public
health needs.
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.
MRIS program website: http: IIwww.epa.gov /iris /.
2EPA guidance documents: http: IIwww.epa.gov/
iris/basic-information-about-integrated-risk-
information-system#guidance/.
3IRIS multiyear agenda: https: IIwww.epa.gov/
iris/iris-agenda.
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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 assessment
progresses and new agent-specific insights
and issues emerge.
3. Identifying and Selecting
Pertinent Studies
IRIS assessments conduct systematic
literature searches with criteria for inclusion
and exclusion. The objective is to retrieve the
pertinent primary studies (i.e., studies with
4Health and Environmental Research Online:
https://hero.epa.gov/hero/.
original data on health outcomes or their
mechanisms). PECO statements (Populations,
Exposures, Comparisons, Outcomes) govern
the literature searches and screening criteria.
"Populations" and animal species generally
have no restrictions. "Exposures" refers to the
agent and related chemicals identified during
scoping and problem formulation and may
consider route, duration, or timing of
exposure. "Comparisons" means studies that
allow comparison of effects across different
levels of exposure. "Outcomes" may become
more specific (e.g., from "toxicity" to
"developmental toxicity" to "hypospadias") as
an assessment progresses.
For studies of absorption, distribution,
metabolism, and elimination, the first
objective is to create an inventory of pertinent
studies. Subsequent sorting and analysis
facilitates characterization and quantification
of these processes.
Studies on mechanistic events can be
numerous and diverse. Here, too, the objective
is to create an inventory of studies for later
sorting to support analyses of related data.
The inventory also facilitates generation and
evaluation of hypothesized mechanistic
pathways.
The IRIS program posts initial protocols
for literature searches on its website and adds
search results to EPA's HERO database.4 Then
the IRIS program takes extra steps to ensure
identification of pertinent studies: by
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
5IRIS "stopping rules": https: //www.epa.gov/
sites/production/files/2 014-06/documents /
iris stoppingrules.pdf.
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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. 1998a. 1996. 1991b). As
subject-matter experts examine a group of
studies, additional agent-specific knowledge
or methodologic concerns may emerge and a
second pass become necessary.
Assessments use evidence tables to
summarize the design and results of pertinent
studies. If tables become too numerous or
unwieldy, they may focus on effects that are
more important or studies that are more
informative.
The IRIS program posts initial protocols
for study evaluation on its website, then
considers public input as it completes this
step.
5. Integrating the Evidence of
Causation for Each Health
Outcome
Synthesis within lines of evidence. For
each health outcome, IRIS assessments
synthesize the human evidence and the animal
evidence, augmenting each with informative
subsets of mechanistic data. Each synthesis
considers aspects of an association that may
suggest causation: consistency,
exposure-response relationship, strength of
association, temporal relationship, biological
plausibility, coherence, and "natural
experiments" in humans (U.S. EPA. 2005a.
19941.
Each synthesis seeks to reconcile
ostensible inconsistencies between studies,
taking into account differences in study
methods and quality. This leads to a
distinction between conflicting evidence
(unexplained positive and negative results in
similarly exposed human populations or in
similar animal models) and differing results
(mixed results attributable to differences
between human populations, animal models,
or exposure conditions) (U.S. EPA. 2005a).
Each synthesis of human evidence
explores alternative explanations (e.g., chance,
bias, or confounding) and determines whether
they may satisfactorily explain the 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. 2005al.
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
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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.
2005a).
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
information, conflicting evidence, or
negative results not sufficiently robust for
not likely.
Not likely to be carcinogenic to humans: robust
evidence to conclude that there is no basis
for concern. Examples include no effects in
well-conducted studies in both sexes of
multiple animal species, extensive
evidence showing that effects in animals
arise through modes-of-action that do not
operate in humans, or convincing evidence
that effects are not likely by a particular
exposure route or below a defined dose.
If there is credible evidence of
carcinogenicity, there is an evaluation of
mutagenicity, because this influences the
approach to dose-response assessment and
subsequent application of adjustment factors
for exposures early in life (U.S. EPA. 2005a. b).
6. Selecting Studies for Derivation of
Toxicity Values
The purpose of toxicity values (slope
factors, unit risks, reference doses, reference
concentrations; see section 7) is to estimate
exposure levels likely to be without
appreciable risk of adverse health effects. EPA
uses these values to support its actions to
protect human health.
The health outcomes considered for
derivation of toxicity values may depend on
the hazard descriptors. For example, IRIS
assessments generally derive cancer values
for agents that are carcinogenic or likely to be
carcinogenic, and sometimes for agents with
suggestive evidence (U.S. EPA. 2005a).
Derivation of toxicity values begins with a
new evaluation of studies, as some studies
used qualitatively for hazard identification
may not be useful quantitatively for
exposure-response assessment Quantitative
analyses require quantitative measures of
exposure and response. An assessment weighs
the merits of the human and animal studies, of
various animal models, and of different routes
and durations of exposure fU.S. EPA. 19941.
Study selection is not 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.
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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 (U.S. EPA. 2011.2006.
2005a. 19941.
For human studies, an assessment may
develop exposure-response models that
reflect the structure of the available data (U.S.
EPA. 2005al For animal studies, EPA has
developed a set of empirical ("curve-fitting")
models6 that can fit typical data sets (U.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
benchmark Dose Software:
http: II www, epa. go v /bmds /.
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.
2012, 2005a).
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 (U.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.
This is appropriate for agents with direct
mutagenic activity. It is also the default if there
is no established mode-of-action (U.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 (U.S. EPA. 2005a. b).
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
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risk of adverse health effects over a lifetime
(U.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. 2014.
20021.
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 factor may not be necessary for
reference values of shorter duration.
Adverse-effect level to no-observed-adverse-
effect level: For reference values based on
a lowest-observed-adverse-effect level, an
uncertainty factor reflects a level judged to
have no observable adverse effects.
Database deficiencies: If there is concern that
future studies may identify a more
sensitive effect, target organ, population,
or lifestage, a database uncertainty factor
reflects the nature of the database
deficiency.
8. Process for Developing and
Peer-Reviewing IRIS Assessments
The IRIS process (revised in 2009 and
enhanced in 2013) involves extensive public
engagement and multiple levels of scientific
review and comment IRIS program scientists
consider all comments. Materials released,
comments received from outside EPA, and
disposition of major comments (steps 3,4, and
6 below) become part of the public record.
Step 1: Draft development. As outlined in
section 2 of this Preamble, IRIS program
scientists specify the scope of an
assessment and formulate science issues
for discussion with the scientific
community and the public. Next, they
release initial protocols for the systematic
review procedures planned for use in the
assessment. IRIS program scientists then
develop a first draft, using structured
approaches to identify pertinent studies,
evaluate study methods and quality,
integrate the evidence of causation for
each health outcome, select studies for
derivation of toxicity values, and derive
toxicity values, as outlined in Preamble
sections 3-7.
Step 2: Agency review. Health scientists
across EPA review the draft assessment.
Step 3: Interagency science consultation.
Other federal agencies and the Executive
Office of the President review the draft
assessment.
Step 4: Public comment, followed by
external peer review. The public reviews
the draft assessment. IRIS program
scientists 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
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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
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
studies or inferred from mechanistic data.
These may warrant further analysis to
quantify differences in susceptibility. Another
subsection identifies biological considerations
for selecting health outcomes, studies, or data
sets, as outlined in Preamble section 6.
Section 2 includes a subsection for each
toxicity value. Each subsection discusses study
selection, methods of analysis, and derivation
of a toxicity value, as outlined in Preamble
sections 6 and 7.
Front matter. The Executive Summary
provides information historically included in
IRIS summaries on the IRIS program website.
Its structure reflects the needs and
expectations of EPA's program and regional
offices.
A section on systematic review methods
summarizes key elements of the protocols,
including methods to identify and evaluate
pertinent studies. The final protocols appear
as an appendix.
The Preface specifies the scope of an
assessment and its relation to prior
assessments. It discusses issues that arose
during assessment development and
emerging areas of concern.
This Preamble summarizes general
procedures for assessments begun after the
date below. The Preface identifies
assessment-specific approaches that differ
from these general procedures.
10.References
U.S. EPA (U.S. Environmental Protection
Agency). (1991b). Guidelines for
developmental toxicity risk
assessment (pp. 1-83). (EPA/600/FR-
91/001). Washington, DC: U.S.
Environmental Protection Agency,
Risk Assessment Forum.
http://cfpub.epa.gov/ncea/cfm/recor
display.cfm?deid=23162
U.S. EPA (U.S. Environmental Protection
Agency). (1994). Methods for
derivation of inhalation reference
concentrations and 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/reco
rdisplay.cfm?deid=71993&CFID=511
74829&CFTOKEN=25006317
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Toxicological Review of Ethyl Tertiary Butyl Ether
U.S. EPA (U.S. Environmental Protection
Agency). (1996). Guidelines for
reproductive toxicity risk assessment
[EPA Report], (EPA/630/R-96/009).
Washington, DC.
http: / /www. ep a. gov/r af/publicatio ns
/pdfs/REPR051.PDF
U.S. EPA (U.S. Environmental Protection
Agency). (1998b). Guidelines for
neurotoxicity risk assessment [EPA
Report], (EPA/630/R-95/001F).
Washington, DC.
http: / /www. ep a. gov/r af/publicatio ns
/pdfs/NEUROTOX.PDF
U.S. EPA (U.S. Environmental Protection
Agency). (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 (U.S. Environmental Protection
Agency). (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 (U.S. Environmental Protection
Agency). (2005b). Supplemental
guidance for assessing susceptibility
from early-life exposure to
carcinogens [EPA Report] (pp. 1125-
1133). (EPA/63 0/R-03/003F).
Washington, DC.
http://www.epa.gov/cancerguideline
s/guidelines- carcinogen-
supplementhtm
U.S. EPA (U.S. Environmental Protection
Agency). (2006). Approaches for the
application of physiologically based
pharmacokinetic (PBPK) models and
supporting data in risk assessment
(Final Report) [EPA Report],
(EPA/600/R-05/043F). Washington,
DC.
http://cfpub.epa.gov/ncea/cfm/recor
display.cfm?deid=l 57668
U.S. EPA (U.S. Environmental Protection
Agency). (2011). Recommended use of
body weight 3/4 as the default method
in derivation of the oral reference dose
(pp. 1-50). (EPA/100/R11/0001).
Washington, DC: U.S. Environmental
Protection Agency, Risk Assessment
Forum, Office of the Science Advisor.
https://www.epa.gov/risk/recomme
nded-use-body-weight-34-default-
method-derivation-oral-reference-
dose
U.S. EPA (U.S. Environmental Protection
Agency). (2012). Benchmark dose
technical guidance (pp. 1-99).
(EPA/100/R-12/001). Washington,
DC: U.S. Environmental Protection
Agency, Risk Assessment Forum.
U.S. EPA. (US. Environmental Protection
Agency). (2014). Guidance for
applying quantitative data to develop
data-derived extrapolation factors for
interspecies and intraspecies
extrapolation [EPA Report],
(EPA/100/R-14/002F). Washington,
DC: Risk Assessment Forum, Office of
the Science Advisor.
https://www.epa.gov/risk/guidance-
applying-quantitative-data-develop-
data-derived-extrapolation-factors-
interspecies-and
August 2016
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EXECUTIVE SUMMARY
Summary of Occurrence and Health Effects
Ethyl tertiary -butyl ether (ETBE) does not occur naturally; it is a man-made ether
oxygenate used primarily 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 through 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 ETBE is
manufactured. The magnitude of human exposure to ETBE depends on factors such
as the distribution of ETBE in groundwater and the extent of the contamination.
Animal studies demonstrate that exposure to ETBE is associated with noncancer
kidney effects following oral and inhalation exposure. Evidence is suggestive of
carcinogenic potential for ETBE based on liver tumors in rats following inhalation
exposure.
ES.l EFFECTS OTHER THAN CANCER OBSERVED FOLLOWING ORAL EXPOSURE
Kidney effects are identified in this assessment as a potential human hazard of ETBE
exposure. Although no human studies are available to evaluate the effects of ETBE, oral exposure
studies in animals have consistently reported increased kidney weight in male and female rats
accompanied by increased chronic progressive nephropathy (CPN), urothelial hyperplasia of the
renal pelvis (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 of the renal
pelvis, and increases in serum markers of kidney function. Additionally, effects were also observed
across routes of exposure, and sex (with the exception of urothelial hyperplasia of the renal pelvis
which was observed only in male rats).
The relevance of the kidney findings to humans was evaluated with respect to alpha
2u-globulin nephropathy, a disease process that occurs exclusively in the male rat kidney (Capen et
al.. 1999: U.S. EPA. 1991al. While ETBE binds to alpha 2u-globulin and meets some criteria of the
alpha 2u-globulin U.S. Environmental Protection Agency (EPA) and International Agency for
Research on Cancer (IARC) frameworks (Capen etal.. 1999: U.S. EPA. 1991a). it does not meet all.
With respect to male rats, U.S. EPA (1991a) noted that "[i]f a compound induces a2u-globulin
accumulation in hyaline droplets, the associated nephropathy in male rats is not an appropriate
endpoint to determine noncancer (systemic) effects potentially occurring in humans." However, as
alpha 2u-globulin nephropathy is strictly a male rat phenomenon, the dose-related kidney effects in
female rats are not confounded by alpha 2u-globulin nephropathy.
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It has been observed that chemicals that bind to alpha 2u-globulin also exacerbate the
incidence and/or severity of background CPN in male rats fFrazier etal.. 2012: Travlos etal.. 2011:
U.S. EPA. 1991al. While the etiology of CPN is unknown fNIEHS. 2019: Hard and Khan. 2004: Peter
etal.. 19861 and it has no known analogue in the aging human kidney fNIEHS. 2019: Hard etal..
20091. it cannot be ruled out that a chemical that exacerbates CPN in rats could also exacerbate
disease processes in the human kidney [e.g., chronic kidney disease, diabetic nephropathy,
glomerulonephritis, interstitial nephritis, etc.; NIEHS f20191]. Therefore, increased incidence of
kidney effects with ETBE exposure in the female rat (but not the male rat) are considered
appropriate for identifying a hazard to the kidney.
Evidence is suggestive that liver toxicity follows oral ETBE exposure. The strongest
supporting evidence is the increased liver weights and centrilobular hypertrophy in exposed male
and female rats consistently reported across oral-exposure studies. No additional histopathological
findings were observed, however, and only one serum marker potentially indicative of liver toxicity
(y-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 minimal and, except for organ-weight data, no
consistent dose-response relationships were observed. Mechanistic data suggest that ETBE
exposure leads to activation of several nuclear receptors, but there is inadequate evidence 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 of ETBE. Thus, even with the
consistently observed increases in rat liver weight and centrilobular hypertrophy, the evidence
remains suggestive that liver toxicity follows ETBE exposure because of the relatively small
magnitude of effects and inconsistent dose-response relationships.
Inadequate information exists to draw conclusions regarding reproductive, developmental,
or immune system effects. The ETBE database does include developmental, reproductive, and
multigenerational studies, which are generally null and do not appear to indicate an area of concern
(see hazard discussions in Sections 1.2.3 and 1.2.4). However, this body of evidence is not
sufficiently robust to conclude that ETBE is not likely to be a reproductive or developmental hazard.
Regarding immune system effects, the ETBE database contains no evidence of altered immune
function that correlate with modest T cell population reductions and altered splenic organ weights
(see Appendix B of the Supplemental Information); thus, the available immune data are inadequate
to draw conclusions as a human hazard of ETBE exposure.
ES.2 ORAL REFERENCE DOSE (RFD) FOR EFFECTS OTHER THAN CANCER
Kidney toxicity, represented by increased absolute kidney weight in female rats, was chosen
as the basis for the overall RfD (see Table ES-1). The chronic study by TPEC f2010al [with selected
data published as Suzuki etal. f20121] and the observed kidney effects were used to derive the RfD.
The endpoint of increased kidney weight was selected as the critical effect because it is a specific
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and sensitive indicator of kidney toxicity and was induced in a dose-responsive manner. The
increases in kidney weight seen in male and female rats are likely a result of the increases in the
severity of CPN seen with ETBE exposure. A characteristic feature of CPN in the rat is increasing
kidney size and weight fHard etal.. 20131. Benchmark dose (BMD) modeling was used to derive the
benchmark dose lower confidence limit corresponding to 10% extra risk (BMDLio) of
120 mg/kg-day. The BMDL was converted to a human equivalent dose (HED) of 28.8 mg/kg-day
using body weight3/4 scaling (U.S. EPA. 20111. and this value was used as the point of departure
(POD) for RfD derivation.
The overall RfD was calculated by dividing the POD for increased absolute kidney weight by
a composite uncertainty factor (UFc) of 30 to account for extrapolation from animals to humans (3)
and interindividual differences in human susceptibility (10).
Table ES-1. Organ/system-specific RfDs and overall RfD for ETBE
Hazard
Basis
Point of departure3
(mg/kg-d)
UFC
Chronic RfD
(mg/kg-d)
Study exposure
description
Confidence
Kidney
Increased absolute
kidney weight
28.8
30
lx 10°
Chronic
High
Overall RfD
Kidney
28.8
30
1 x 10°
Chronic
High
aHuman equivalent dose PODs were calculated using body weight scaling to the 3/4 power (BW3/4) (U.S. EPA,
2011).
ES.3 EFFECTS OTHER THAN CANCER OBSERVED FOLLOWING INHALATION EXPOSURE
Kidney effects are a potential human hazard of inhalation exposure to ETBE. Although no
human studies are available to evaluate the effects of exposure, studies in animals have observed
increases in kidney weight, altered kidney histopathology, as well as alterations in clinical
chemistry including serum cholesterol, BUN, and creatinine. While the histological lesion of
urothelial hyperplasia of the renal pelvis was a sensitive endpoint in male rats, it was not observed
in female rats or mice of either sex, whereas increases in kidney weight were observed in multiple
studies in rats of both sexes and in mice. Changes in kidney weight in female rats were dose-
dependent, consistent across multiple studies and are not confounded by alpha 2u-globulin
nephropathy, and therefore considered appropriate for identifying a hazard to the kidney.
ES.4 INHALATION REFERENCE CONCENTRATION (RFC) FOR EFFECTS OTHER THAN
CANCER
Kidney toxicity, represented by increased absolute kidney weight, was chosen as the basis
for the overall RfC (see Table ES-2). The chronic study by TPEC (2010b) [selected data published as
Saito etal. f20131] and the observed kidney effects were used to derive the RfC. The endpoint,
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increased absolute kidney weight, 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. The increases
in kidney weight are likely a result of the increases in the severity of CPN seen with ETBE exposure,
as CPN is characterized by cell proliferation and chronic inflammation that results in increased
kidney weight fMelnick et al.. 2012: Travlos etal.. 20111.
Benchmark concentration modeling of increased kidney weight (in female rats) was
attempted, but an adequate fit was not achieved. Therefore, a no-observed-adverse-effect level
(NOAEL) was used to derive the POD of 6,270 mg/m3. The NOAEL was adjusted for continuous
exposure and converted to a human equivalent concentration (HEC) of 1,110 mg/m3.
The overall RfC was calculated by dividing the POD for increased absolute kidney weight by
a UF of 30 to account for toxicodynamic differences between animals and humans (3) and
interindividual differences in human susceptibility (10).
Table ES-2. Organ/system-specific RfCs and overall RfC for ETBE
Hazard
Basis
Point of departure3
(mg/m3)
UF
Chronic RfC
(mg/m3)
Study exposure
description
Confidence
Kidney
Increased absolute
kidney weight
1,110
30
4x 101
Chronic
Medium
Overall RfC
Kidney
1,110
30
4x 101
Chronic
Medium
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.
ES.5 EVIDENCE OF HUMAN CARCINOGENICITY
Under EPA Cancer Guidelines (U.S. EPA. 2005a). the evidence of carcinogenic potential for
ETBE is suggestive for inhalation exposure but inadequate for oral exposure. ETBE induced liver
tumors in male (but not female) rats in a 2-year inhalation exposure study fSaito etal.. 2013: TPEC.
2010b). No significant effects were observed in two chronic oral studies in male and female rats
[one of high quality; see Section 1.2.5; TPEC (2010a): Maltoni etal. (1999)]. Data on tumorigenicity
in mice following ETBE exposure were not available. However, supplementary evidence from
two-stage initiation-promotion oral carcinogenesis bioassays indicate increased mutagen-initiated
liver tumors, as well as increased tumor incidence in the thyroid, colon, and urinary bladder.
ES.6 QUANTITATIVE ESTIMATE OF CARCINOGENIC RISK FROM ORAL EXPOSURE
A quantitative estimate of carcinogenic potential from oral exposure to ETBE was not
derived because an increase in tumors was not observed in the two available chronic oral cancer
bioassays (TPEC. 2010a: Maltoni et al.. 1999). A route-to-route extrapolation of cancer risk from the
inhalation-to-oral route was not carried out because there was no consistent dose-response
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relationship observed for liver tumors when compared across oral and inhalation studies on the
basis of physiologically based pharmacokinetic (PBPK) modeled internal dose.
ES.7 QUANTITATIVE ESTIMATE OF CARCINOGENIC RISK FROM INHALATION EXPOSURE
Although ETBE was considered to have suggestive evidence of carcinogenic potential, the
main study (Saito etal.. 2013: TPEC. 2010b) was conducted according to well-established guidelines
for examining potential carcinogenicity and was suitable for quantitative analyses. Thus, while
recognizing the uncertainty in the data and the suggestive nature of the weight of evidence, the
analysis may be useful for some purposes, such as providing a sense of the magnitude of potential
risks, ranking potential hazards, or setting research priorities fU.S. EPA. 2005a).
A quantitative estimate of carcinogenic potential from inhalation exposure to ETBE was
based on the increased incidence of hepatocellular adenomas and carcinomas in male F344 rats
following 2-year inhalation exposure (Saito 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 the
animals for up to 2 years, and included detailed reporting of methods and results.
An inhalation unit risk 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 benchmark concentration
lower confidence level corresponding to 10% extra risk (BMCLio), a human equivalent inhalation
unit risk was derived using inhalation unit risk = 0.1/BMCLio and calculated to be
8 x 10"5 per mg/m3.
ES.8 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; Brennan etal. f20041].
Evidence is strong in humans that this ALDH2 variant increases the internal dose of acetaldehyde
and the cancer risks from acetaldehyde, especially in the development of ethanol-related cancers
fEriksson. 2015: IARC. 20101. Several in vivo and in vitro genotoxicity assays in Aldh2 knockout
(KO) and heterozygous 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 (Weng etal.. 2019: Weng etal.. 2014: Weng etal.. 2013: Weng etal.. 2012: Weng
etal.. 20111. Inhalation ETBE exposure increased blood concentrations of acetaldehyde in Aldh2
KO mice compared with wild type fWeng etal.. 20131. Thus, exposure to ETBE in individuals with
the ALDH2*2 variant would be expected to increase the internal dose of acetaldehyde and
potentially increase risks associated with acetaldehyde produced by ETBE metabolism in the liver.
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Collectively, these data present evidence that people with diminished ALDH2 activity could
be considered a susceptible population that could be more sensitive to liver toxicity from ETBE
exposure.
ES.9 KEY ISSUES ADDRESSED IN ASSESSMENT
Dose-related kidney effects were observed in male and female rats. The human relevance of
these effects, particularly as they relate to alpha 2u-globulin nephropathy and the exacerbation of
chronic progressive nephropathy, is a key issue analyzed in this assessment. An evaluation of
whether ETBE causes alpha 2u-globulin-associated nephropathy was performed using the EPA and
IARC frameworks (Capenetal.. 1999: U.S. EPA. 1991a). ETBE induced an increase in hyaline
droplet accumulation and increased alpha 2u-globulin deposition in male rats; however, most of the
subsequent steps in the pathological sequence were not observed. U.S. EPA f!991al states that "[i]f
a compound induces a2u-globulin accumulation in hyaline droplets, the associated nephropathy in
male rats is not an appropriate endpointto determine noncancer (systemic) effects potentially
occurring in humans." However, because alpha 2u-globulin nephropathy is strictly a male rat
phenomenon, the dose-related kidney effects in female rats are not confounded by alpha
2u-globulin nephropathy. CPN also plays a role in exacerbating nephropathy in rats; however, the
mode of action (MOA) is unknown. Given that there is no definitive pathogenesis for CPN, it cannot
be fully ruled out that chemicals that exacerbate CPN in rats may have the potential to exacerbate
other disease processes in the human kidney (NIEHS. 2019). Dose-related changes in several
indicators of kidney toxicity were observed in male and female rats, including increased absolute
kidney weight, histological changes, and increased blood biomarkers (Saito etal.. 2013: Suzuki et
al.. 2012: TPEC. 2010a. b). These specific effects are considered relevant to humans, particularly the
endpoints observed in female rats, because they are not confounded by alpha 2u-globulin related
processes.
In addition, the human relevance of the observed liver tumors is discussed in the
assessment (see Sections 1.2.2 and 1.3.2). Briefly, a well-conducted inhalation study demonstrated
a significant, positive exposure-response for hepatocellular adenomas and carcinomas in male rats
(Saito etal.. 2013: TPEC. 2010b). While the majority of liver tumors occurred at the highest
exposure, statistical tests conducted by the study authors found a significant dose-response trend
by both the Peto (incidental tumor test) and the Cochran-Armitage tests. However, two chronic
oral exposure studies (one with unrelated mortality were negative for liver tumors flPEC. 2010a:
Maltoni et al.. 1999). The integration of relevant carcinogenic evidence is discussed in Section 1.3.2.
The potential MOA for the observed liver tumors was accessed in the Section 1.2.2. The
available evidence base for the nuclear hormone receptor MOAs (i.e., peroxisome
proliferator-activated receptor a [PPARa], pregnane X receptor [PXR], and the constitutive
androstane receptor [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
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genotoxicity in exposed mice, the available database was inadequate to establish
acetaldehyde-mediated mutagenicity as an MOA for ETBE-induced liver tumors. No other MOAs for
liver carcinogenesis were identified. Because an MOA for liver carcinogenicity could not be
established, in the absence of data to indicate otherwise, the rat liver tumors observed following
inhalation exposure 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
ethyl tertiary butyl ether (ETBE), and remaining references were sorted into categories for further
evaluation. The original chemical-specific search was conducted in four online scientific databases,
PubMed, Toxline, Web of Science, and Toxic Substances Control Act Test Submissions (TSCATS),
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 of ETBE 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 of ETBE. 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.
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• 715 references were identified as being not pertinent (not on topic) for evaluating
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 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 Health and Environmental Research Online (HERO) website at
https://hero.epa.gov/hero/index.cfm/project/page/project id/1376.
LS.l POSTPEER-REVIEW LITERATURE SEARCH UPDATE
A post-peer-review literature search update was conducted in PubMed, Toxline, TSCATS,
and Defense Technical Information Center (DTIC) for the period December 2016 to July 2019 using
a search strategy consistent with previous literature searches (see Table LS-1). The documentation
and results for the literature search and screen, including the specific references identified using
each search strategy and tags assigned to each reference based on the manual screen, can be found
on the HERO website on the ETBE project page at:
fhttps://hero.epa.gov/hero/index.cfm/proiect/page/project id/13761.
Consistent with the Integrated Risk Information System (IRIS) Stopping Rules
fhttps://www.epa.gov/sites/production/files/2014-06/documents/iris stoppingrules.pdFl.
manual screening of the literature search update focused on identifying new studies that might
change a major conclusion of the assessment. The last formal literature search was in 2019 while
the draft was in external peer review, after which the literature was monitored in PubMed through
January 2021. No animal bioassays or epidemiological studies were identified in the
post-peer-review literature searches that would change any major conclusions in the assessment.
LS.2 SELECTION OF STUDIES FOR INCLUSION IN EVIDENCE TABLES
To summarize the important information systematically from the primary health effects
studies in the ETBE evidence base, evidence tables were constructed in a standardized tabular
format as recommended by NRC (2011). 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, 33 were identified as "Sources of Health Effects Data" and considered for extraction into
evidence tables for the hazard identification in Section 1. An initial review of studies examining
neurotoxic endpoints did not find consistent effects to warrant a comprehensive hazard evaluation;
thus, the one subchronic study (Dorman et al.. 1997) that examined only neurotoxic endpoints
(functional observational battery, motor activity, and terminal neuropathology) was not included in
evidence tables. Data from the remaining 32 studies were extracted into evidence tables.
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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 of ETBE 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 Section 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 evidence
base 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, for example, in sections such as
"Mode-of-Action Analysis" to augment the discussion or are presented in appendices, if they
provide additional information.
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Database Searches
(see Table LS-1 for keywords and limits)
/
"\
PubMed
n = 148
V
/
Web of Science
n = 518
Toxline
(incl. TSCATS)
n = 105
TSCATS 2
n = 1
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
<
Secondary Literature and Sources of
Contextual Information (n = 29)
1 QSAR
7 Mixtures
14 Reviews/editorials
5 Other agency assessments
2 Odor threshold
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
ADME = absorption, distribution, metabolism, excretion; QSAR = quantitative structure-activity relationship.
Figure LS-1. Summary of literature search and screening process for ETBE.
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Table LS-1. Details of the search strategy employed for ETBE
Database
(search date)
Keywords
Limits
PubMed
(03/31/2014)
Updated (12/2016)
"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 (12/2016)
"ETBE" OR "ethyl tert-butyl ether" OR "2-ethoxy-2-methy!-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 only
(e.g., the search term
is reduced to its lexical
root)
Toxline (includes
TSCATS)
(03/31/2014)
Updated (12/2016)
"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
01/2004 to 7/2019
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). "Ethvl
tertiary-butyl ether: a toxicological review."
Critical Reviews in Toxicology 37(4): 287-312
3/2014
68 references
Review article: de Peyster (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
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Table LS-3. Inclusion-exclusion criteria
Inclusion criteria
Exclusion criteria/supplemental material3
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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Table LS-3. Inclusion-exclusion criteria (continued)
Inclusion criteria
Exclusion criteria/supplemental material3
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
aStudies that met this exclusion criterion were considered supplemental (i.e., not considered a primary source of
health effects data, but were retained as potential sources of contextual information).
LS.3 EVIDENCE BASE EVALUATION
For this draft assessment, 33 experimental animal studies were the primary sources of
health effects data; no studies were identified that evaluated humans exposed to ETBE (e.g., cohort
studies, case reports, ecological studies). The animal studies were evaluated by considering aspects
of design, conduct, or reporting that could affect the interpretation of results, overall contribution
to the synthesis of evidence, and determination of hazard potential as noted in various EPA
guidance documents fU.S. EPA. 2005a. 1998a. 1996.1991b). The objective was to identify the
stronger, more informative studies based on a uniform evaluation of quality characteristics across
studies of 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;
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• 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;
• Ascertainment of survival, vital signs, disease or effects, and cause of death; and
• 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 (see 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.
LS.3.1. Experimental Animal Studies
The 33 experimental animal studies, all of which were performed on rats, mice, and rabbits,
were associated with drinking water, gavage, or inhalation exposures to ETBE. Many of these
studies were conducted according to Organisation for Economic Co-operation and Development
Good Laboratory Practice (GLP) guidelines and used well-established methods, were well reported,
and evaluated an extensive range of endpoints and histopathological data. For the body of available
studies, detailed discussion of any identified methodological concerns precedes each endpoint
evaluated in the hazard identification section. Overall, the experimental animal studies of ETBE
involving repeated oral or inhalation exposure were considered acceptable in quality, and whether
yielding positive, negative, or null results, were considered in assessing the evidence for health
effects associated with chronic exposure to ETBE.
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Table LS-4. Considerations for evaluating 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., h/d, d/wk); 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,
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)
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Table LS-5. Summary of experimental animal evidence base
Study category
Study duration, species/strain, and administration method
Chronic
2-vr study in F344 rats (drinking water) JPEC (2010a);a Suzuki et al. (2012)
2-yr study in F344 rats (inhalation) JPEC (2010b);a Saito et al. (2013)
2-vr studv in S-D rats (gavage) Maltoni et al. (1999)
Subchronic
13-wk studv in F344 rats (inhalation) Medinskv et al. (1999); U.S. EPA (1997)
26-wk studv in S-D rats (gavage) JPEC (2008c);a Mivata et al. (2013)
Fuiiietal. (2010); JPEC (2008e)
13-wk studv in S-D rats (inhalation) JPEC (2008b)a
23-wk studv in F344 rats (gavage) Hagiwara et al. (2011); JPEC (2008d)
13-wk studv in CD-I mice (inhalation) Medinskv et al. (1999); Bond et al. (1996)
23-wk studv in Wistar rats (gavage) Hagiwara et al. (2015)
31-wk studv in F344/DuCrlCrli rats (drinking water) Hagiwara et al. (2013)
13-wk studv in C57BL/6 mice (inhalation) Weng et al. (2012)
Reproductive
Two-generation reproductive toxicity studv on S-D rats (gavage) Gaoua (2004b)a
One-generation reproductive toxicity studv on S-D rats (gavage) Fuiii et al. (2010); JPEC
(2008e)
2-wk studv on Simonsen albino rats (drinking water) Berger and Horner (2003)
9-wk studv on C57BL/6 mice (inhalation) Weng et al. (2014)
14-d studv on F344 rats (gavage) de Pevster et al. (2009)
Two-generation reproductive toxicity studv in S-D rats (gavage) Gaoua (2004b)a
Developmental
Developmental studv (GD 6-27) on New Zealand White rabbits (gavage) Asano et al. (2011);
JPEC (2008i)
Developmental studv (GD 5-19) on S-D rats (gavage) Aso et al. (2014); JPEC (2008h)
Developmental studv (GD 5-19) on S-D rats (gavage) Gaoua (2004b)a
Developmental studv (GD 5-19) on S-D rats (gavage) Gaoua (2004a)a
Pharmacokinetic
Single-dose studv on S-D rats (gavage) JPEC (2008g)
14-d studv on S-D rats (gavage) JPEC (2008f)
Single-dose studv on S-D rats (gavage) JPEC (2008g)a
14-d studv on S-D rats (gavage) JPEC (2008f)a
GD = gestation day; S-D = Sprague-Dawley.
aThe IRIS program had this study peer reviewed.
xli
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Toxicological Review of Ethyl Tertiary Butyl Ether
1.HAZARD IDENTIFICATION
1.1. OVERVIEW OF CHEMICAL PROPERTIES AND TOXICOKINETICS
1.1.1. Chemical Properties
Ethyl tertiary butyl ether (ETBE) is a liquid at a temperature range of -94 to 72.6°C. It is
soluble in ethanol, ethyl ether, and water fDrogos and Diaz. 20011. ETBE has a strong, highly
objectionable odor and taste at relatively low concentrations. The chemical is highly flammable and
reacts with strong oxidizing agents. ETBE is stable when stored at room temperature in tightly
closed containers (Drogos and Diaz. 20011. Information on the physiochemical properties for ETBE
is available at U.S. Environmental Protection Agency (EPA)'s CompTox Chemicals Dashboard
(https://comptox.epa.gOv/dashboard/l and is summarized in Table 1-1.
1.1.2. Toxicokinetics
ETBE is rapidly absorbed following exposure through oral and inhalation routes (see
Appendix Section B.l.l). Studies in experimental animals indicate that >90% of the compound is
absorbed after oral administration within 6-10 hours (TPEC. 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 of ETBE was estimated to be 26% in a
short-term exposure study in humans conducted during light physical activity fNihlen etal..
1998b).
ETBE and its metabolite, tert-butanol, are distributed throughout the body following oral,
inhalation, and intravenous (i.v.) exposures (TPEC. 2008d. e; Poet etal.. 1997: Faulkner et al.. 1989:
ARCO. 19831. Following exposure to ETBE in rats, ETBE was found in the kidney, liver, and blood.
Comparison of the ETBE distribution in rats and mice demonstrated that concentrations of ETBE in
the rat kidney and mouse liver are proportional to the blood concentration.
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-1. Chemical identity and physicochemical properties of ETBE from
EPA's CompTox Chemicals Dashboard
Characteristic or property
Value
Chemical structure
h3c
VCHs
CASRN
637-92-3
Synonyms
Ethyl t-butyl ether; 2-ethoxy-2-methylpropane; propane, 2-ethoxy-2-methyl; Ethyl
te/t-butyl ether; 2-Methyl-2-ethoxypropane
(see https://comptox.epa.gov/dashboard for additional svnonvms)
Molecular formula
C6Hi40
Molecular weight (g/mol)
102.177
Average experimental value3
Average predicted value3
Flash point (°C)
—
-10.9
Boiling point (°C)
72.4
74.3
Melting point (°C)
-94
-90.8
Log Kow
—
1.72
Density (g/cm3)
—
0.768
Henry's law constant
(atm-m3/mole)
1.64 x 10"3
1.66 x 10"3
Vapor pressure (mm Hg at
20°C)
124
124
atm = atmosphere; CASRN = Chemical Abstracts Service registry number.
aMedian values and ranges for physicochemical properties of ETBE are also provided on the CompTox Chemicals
Dashboard at https://comptox.epa.gov/dashboard/.
A general metabolic scheme for ETBE, illustrating the biotransformation in rats and
humans, is shown in Figure 1-1 (see Appendix Section B.1.3).
Human data on the excretion of ETBE was measured in several studies fAmberg et al.. 2000:
Nihlen etal.. 1998a. c). The half-life of ETBE in urine was biphasic with half-lives of 8 minutes and
8.6 hours flohanson etal.. 19951. These studies showed urinary excretion of ETBE to be less than
0.2% of the uptake or absorption of ETBE (Nihlen etal.. 1998a. c). Ambergetal. (2000) observed a
similar half-life of 1-6 hours after human exposure to ETBE of 170 mg/m3; however, the
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Toxicological Review of Ethyl Tertiary Butyl Ether
elimination of ETBE in rats was considerably faster than in humans, and only the metabolites of
ETBE were detectable in rat urine, not the parent compound itself. No studies were identified
which investigated lactational transfer of ETBE in humans or animal models.
A more detailed summary of ETBE toxicokinetics is provided in Appendix Section B.l.
CH,
CH3
ETBE
glucuronide-0 -
-CH,
CYP2A6
CYP3A4
CH3
t-butyl glucuronide
rats, humans
l-UC-
-CH,
ChU
OH
ETBE-hemi-acetal
CYP450
h3c-
acetaldehyde
\.
/s'
o^\
ch3
t-butanol
^ CH,
rats,
humans
A ~
CH3 oh
2-methyl-1,2-propanediol
f
H3C—^OH
CH,
t-butyl sulfate
H°_°
[O]
2-hydroxyisobutyric acid
h2c=o
formaldehyde
o
HoC
acetone
CH,
CYP450 = cytochrome P450.
Source: Adapted from Dekant et al. (2001), NSF International (2003), ATSDR (1996), Bernauer et al. (1998), Amberg
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 predicting the adsorption, distribution, metabolism and excretion of ETBE in rats
fBorghoffetal.. 2016: Salazar etal.. 20151. The previously available models have studied
tert-butanol as the primary metabolite after oral or inhalation exposure to methyl tertiary butyl
ether (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 alpha 2u-globulin (Borghoffetal..
2010: Leavens and Borghoff. 2009). A PBPK model for inhalation exposure of humans to ETBE has
also been reported (Nihlen and Tohanson. 19991.
All available PBPK models of ETBE and its principal metabolite tert-butanol were evaluated
for potential use in this assessment (for extrapolation from animals to humans and for
extrapolation between routes of exposure). Regarding the extrapolation from animals to humans,
the existing human PBPK model was not considered adequate (see Appendix Section B.1.7);
therefore, default methodologies were applied to extrapolate toxicologically equivalent exposures
from adult laboratory animals to adult humans. Regarding the extrapolation between routes of
exposure, extrapolation from inhalation to oral routes of exposure was not supported due to the
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Toxicological Review of Ethyl Tertiary Butyl Ether
lack of a consistent dose-response relationship for liver tumors combined across oral and
inhalation studies.
A detailed summary of the toxicokinetic models is provided in Appendix Section 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 the effects induced by these
metabolites can inform the evaluation of ETBE-induced effects. Some of the noncancer kidney
effects observed in ETBE have been attributed to tert-butanol fSalazar 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/transitional epithelial hyperplasia (in males
and females), and increased suppurative inflammation in females (NTP. 1997.1995).
Inhalation exposures to acetaldehyde have been determined to cause carcinomas of the
nasal mucosa in rats and carcinomas of the larynx in hamsters (IARC. 1999b). In addition,
acetaldehyde is known to be the key metabolite in the development of cancer of the esophagus and
aerodigestive tract associated with ethanol consumption (IARC. 2010).
MTBE is a compound structurally related to ETBE 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 (ATSDR. 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 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; IARC (1999c)]. Although some similar effects are seen
with tert-butanol and ETBE, the evidence from MTBE is confounded by its metabolite
formaldehyde, a known human carcinogen (as classified by IARC and National Toxicology Program
[NTP]).
1.2. PRESENTATION AND SYNTHESIS OF EVIDENCE BY ORGAN/SYSTEM
1.2.1. Kidney Effects
Synthesis of Effects in the Kidney
This section reviews the studies that investigated whether subchronic or chronic exposure
to ETBE can cause kidney toxicity in humans or animals. The evidence base examining kidney
effects following ETBE exposure contains no human data and 10 animal studies, predominantly in
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Toxicological Review of Ethyl Tertiary Butyl Ether
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 (see 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 Mivataetal. (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.
In an unpublished analysis by Cohen etal. f20111. a pathology working group reexamined
kidney histopathology from the TPEC f2010al and TPEC f2007al studies. Cohen etal. f20111
reported slightly different incidences and severity of CPN, but incidences of carcinomas did not
differ from the original study (Suzuki etal.. 2012: TPEC. 2010a). Histopathological interpretations
from both Cohen etal. (2011) and TPEC (2007b) are considered for hazard identification and are
presented in Table 1-2.
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 et
al. (2012): U.S. EPA (1991a): http://ntp.niehs.nih.gov/nnl/urinary/kidney/necp/index.htm]. CPN
is more severe in male rats than in females and is particularly common in the Sprague-Dawley (S-D)
and Fischer 344 (F344) strains. Dietary and hormonal factors play a role in modifying CPN,
although the etiology is unknown (see further discussion below).
Kidney weight. Increases in kidney weight serve as a general indication of renal toxicity.
Kidney weights (see Figure 1-2) exhibited strong dose-related increases with estimates of the
primary ETBE metabolite, tert-butanol, in blood in male rats following oral exposures (Spearman's
rank coefficient [rho] = 0.86, p < 0.01) of 16 weeks or longer (Mivataetal.. 2013: Suzuki etal.. 2012:
Fuiii etal.. 2010: TPEC. 2010a. 2008c: Gaoua. 2004b). and following inhalation exposures
(rho = 0.71, p = 0.05) of 13 weeks or longer f Saito etal.. 2013: TPEC. 2010b. 2008b: Medinskv etal..
19991. Kidney weight also showed strong dose-related increases following inhalation exposure
(rho = 0.82, p = 0.01) and moderate dose-related increases following oral exposure (rho = 0.42,
p = 0.2). Short-term inhalation studies in rats also showed increased kidney weight (TPEC. 2008a).
In utero ETBE exposure induced greater increases in kidney weights in F1 male and female rats
compared with parental exposure in one unpublished study (independently peer reviewed via EPA
contract, see Table LS-5) but the magnitude of increases were comparable to those observed in
other adult oral studies fGaoua. 2004bl. A 13-week mouse inhalation study observed small
increases in kidney weight in both sexes, with a greater magnitude of effect in males, up to 10% as
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Toxicological Review of Ethyl Tertiary Butyl Ether
compared with 4-6% increases in females [see Figure 1-3 (Medinskv etal.. 1999: Bond etal..
1996)].
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 et
al.. 2013: Suzuki etal.. 2012: TPEC. 2010b. 2008b. c; Gaoua. 2004bl For ETBE, body weight in
exposed animals was consistently decreased at several doses relative to controls in the oral and
inhalation studies. Thus, use of relative organ-weight change may not be a reliable measure of the
kidney-weight increases for this assessment Additionally, a recent analysis indicates that
statistically significant increases in subchronic absolute, but not relative, kidney weights correlate
well with chronic renal histopathologyf Craig etal.. 20141. Therefore, absolute weight was used as
the more appropriate measure of kidney-weight change for determining ETBE hazard potential.
Absolute and relative kidney-weight data are presented in Appendix B of the Supplemental
Information.
Interpretation of 2-year kidney-weight data in male rats treated by inhalation is
complicated by increased mortality attributed to CPN which would be expected to bias the analysis
of kidney weight (toward the null) because organ weight was not assessed in animals that did not
survive to study termination f Saito 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.. 2011). thus animals severely affected
by CPN, including those that died due to CPN, would be expected to have enlarged kidneys.
Although mortality in female rats in the 2-year inhalation study was also significantly increased, the
study authors attributed these deaths to pituitary tumors, which would not be expected to bias
measurement of kidney weight fTPEC. 2010bl. Mortality of male and female rats in the 2-year
drinking water studies was not significantly different from controls flPEC. 2010a],
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 of ETBE exposure. However, the severity of CPN
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 et al.
f20111: TPEC f2007al: TPEC f2010bl]. 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 (Salazar etal.. 2015).
CPN is a common and well-established constellation of age-related lesions in the kidney of
rats, although the mode of action of CPN is not known. In addition, no known counterpart to CPN
has been identified in the aging human kidney. However, several individual lesions noted in CPN
(e.g., tubule atrophy, tubule dilation, thickening of tubular basement membranes,
glomerulosclerosis) also occur in the human kidney fLusco etal.. 2016: Zoiaetal.. 2015: Seelv and
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Toxicological Review of Ethyl Tertiary Butyl Ether
Brix. 2014a: Frazier etal.. 2012: Satirapoj etal.. 20121. Therefore, exacerbation of one or more of
these lesions following ETBE exposure may reflect some type of cell injury or inflammatory
process, which is relevant to the human kidney.
Increased incidence of urothelial hyperplasia (also known as transitional epithelial
hyperplasia) of the renal pelvis (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).
However, urothelial hyperplasia was not observed in female rats following 2 years of oral or
inhalation exposure (Saito etal.. 2013: Suzuki etal.. 2012: TPEC. 2010a. b). The histological finding
of urothelial hyperplasia represents in an increase in the layers of urothelium and is typically
associated with inflammation or neoplasia fSeelv and Brix. 2014b: Peterson et al.. 20091. The
increase in urothelial hyperplasia in male rats appeared to be dose related on an internal dose basis
across routes of exposure fSalazar etal.. 20151. Cohen etal. f20111. however, attributed this effect
to CPN rather than the "direct" result of ETBE treatment In addition, there is some confusion
regarding the terminology of "urothelial hyperplasia of the renal pelvis" as reported in the JPEC
studies (TPEC. 2010a. b). Specifically, based on the pathological description of this lesion, it may
represent proliferation of the papillary lining epithelium and not true "urothelial hyperplasia"
fNIEHS. 20191. Hyperplasia of the epithelial lining of the renal papilla has been associated with
advanced CPN fSeelv and Brix. 2014al.
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 (see Table 1-5, Table 1-6). Urothelial hyperplasia and
severity of CPN were weakly correlated (rho = 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 alpha
2u-globulin (Mivata etal.. 2013: TPEC. 2008c. e, f; Medinskv etal.. 19991. The significance of this
finding, 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. Chronic and subchronic studies of male rats found 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 (Mivata 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; this dose corresponded
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Toxicological Review of Ethyl Tertiary Butyl Ether
with an elevated severity of CPN in females (Saito etal.. 2013: TPEC. 2010bl. 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:
TPEC. 201 Obi.
Kidney tumors. No increase in kidney tumor incidence was observed following chronic
oral or inhalation exposure in either F344 rats (Saito etal.. 2013: Suzuki etal.. 2012: TPEC. 2010a. b)
or S-D rats [Maltoni et al. (19991: see Table 1-4]. However, animals in the Maltoni et al. (19991
study had extremely low survival in the control and treated groups (approximately 25-30%) after
104 weeks, which is much lower than typical for a 2-year study, ultimately limiting the ability of
this study to predict potential carcinogenicity (see discussion in Section 1.2.5).
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 of diethylnitrosamine (DEN), iV-methyl-iV-nitrosourea (MNU),
/V-butyl-/V-(4-hydroxybutyl)nitrosamine (BBN), 1,2-dimethylhydrazine dihydrochloride (DMH), and
iV-bis(2-hydroxypropyl)nitrosamine (DHPN) (DMBDD) in male F344 rats (Hagiwara etal.. 2011:
TPEC. 2008dl: however, a moderate, but statistically significant dose-response trend in the
incidence of renal tubular adenoma or carcinoma incidence was observed with 19 weeks of daily
gavage ETBE exposure following 2 weeks of the mutagen (/V-ethyl-/V-hydroxyethylnitrosamine
[EHEN]) administration in male Wistar rats (Hagiwara etal.,2015). In Hagiwara et al. (2011).
kidney tumors were not observed following 23 weeks of ETBE exposure without mutagen exposure
[n = 11). An ETBE-only exposure group was not evaluated in the later study in Wistar rats
(Hagiwaraetal.. 20151.
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Toxicological Review of Ethyl Tertiary Butyl Ether
Male rats
Female rats
rho= 0.75 (all)
rho= 0.86 (oral)
rho= 0.71 (inh.)
• ••
o
o
rho= 0.46 (all)
rho= 0.42 (oral)
rho= 0.82 (inh.)
o
•
o
09
O
o
•
0
. ° •
•
•
o
•
o*
o
•
% . '
om
•
0 20 40 60
tert-butanol blood concentration (mg/l)
0 20 40 60
tert-butanol blood concentration (mg/l)
• Oral exposure
O Inhalation exposure
rho = Spearman's rank coefficient.
Figure 1-2. Comparison of absolute kidney-weight change in male and female
rats across oral and inhalation exposure based on metabolite internal blood
concentration. 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 of Ethyl Tertiary Butyl Ether
Mouse inhalation exposure
-------
Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-2. Evidence pertaining to kidney histopathology effects in animals
following exposure to ETBE
Reference and study design
Results
Cohen et al. (2011)
Rat, F344/DuCrlCrlj
oral—drinking water
male (50/group): 0, 625, 2,500,
or 10,000 ppm (0, 28,121, or
542 mg/kg-d);a female
(50/group): 0, 625, 2,500, or
10,000 ppm (0, 46,171, or
560 mg/kg-d)a
reanalysis of histopathology
data from J PEC (2010a) studv,
for which animals were dosed
daily for 104 wk
Male
Female
Dose
(mg/kg-d)
Average
severity of
CPN
Incidence
of CPN
Dose
(mg/kg-d)
Average
severity of
CPN
Incidence
of CPN
0
2.08
49/50
0
1.14
45/50
28
-
-
46
0.98
41/50
121
-
-
171
1.2
46/50
542
2.72b
50/50
560
1.36
46/50
Cohen et al. (2011)
Rat, F344/DuCrlCrlj
oral—drinking water
male (10/group): 0, 250,1,600,
4,000, or 10,000 ppm
(0, 17, 40, 101, 259, or
626 mg/kg-d)a
reanalysis of histopathology
data from J PEC (2007a) (studv
Male
Dose (mg/kg-d)
Number of CPN foci/rat
Number of granular casts/rat
0
1.2
0
17
-
-
40
-
-
101
-
-
No. 0665) study, for which
animals were dosed daily for
13 wk
259
-
-
626
27.2
8.2
Miyata et al. (2013); JPEC
(2008c)
Rat, Crl:CD(SD)
oral—gavage
male (15/group): 0, 5, 25,100,
or 400 mg/kg-d; female
(15/group): 0, 5, 25,100, or
400 mg/kg-d
daily for 180 d
Male
Female
Dose
(mg/kg-d)
Incidence of papillary
mineralization
Dose
(mg/kg-d)
Incidence of papillary
mineralization
0
0/15
0
0/15
5
0/15
5
-
25
0/15
25
-
100
1/15
100
-
400
0/15
400
0/15
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-2. Evidence pertaining to kidney histopathology effects in animals
following exposure to ETBE (continued)
Reference and study design
Results
Saito et al. (2013); JPEC (2010b)
Rat, F344
inhalation—vapor
male (50/group): 0, 500,1,500,
or 5,000 ppm (0, 2,090, 6,270,
or 20,900 mg/m3);c female
(50/group): 0, 500,1,500, or
5,000 ppm (0, 2,090, 6,270, or
20,900 mg/m3)c
dynamic whole-body
inhalation; 6 h/d, 5 d/wk for
104 wk; generation method,
analytical concentration
reported
Male
Dose
(mg/m3)
Average
severity of
CPN as
calculated by
EPAd
Incidence of
CPN
Incidence of
papillary
mineralization
Incidence of
urothelial
hyperplasia of
the renal pelvis
0
2.4
49/50
0/50
2/50
2,090
2.6
50/50
0/50
5/50
6,270
2.7
49/49
1/49
16/49b
20,900
_Q
1
rn
50/50
6/50b
41/50b
Female
Dose (mg/m3)
Average severity of CPN
as calculated by EPAd
Incidence of CPN
0
2.4
49/50
2,090
2.6
50/50
6,270
2.7
49/49
20,900
_Q
1
rn
50/50
Atypical tubule hyperplasia not observed in males or females.
Papillary mineralization and urothelial hyperplasia of the renal pelvis not
observed in females.
Suzuki et al. (2012); JPEC
(2010a)
Rat, F344
oral—water
male (50/group): 0, 625, 2,500,
or 10,000 ppm (0, 28,121, or
542 mg/kg-d);a female
(50/group): 0, 625, 2,500, or
10,000 ppm (0, 46,171, or
560 mg/kg-d)a
daily for 104 wk
Male
Dose
(mg/kg-d)
Average
severity of
CPN
Average severity of
CPN as calculated by
EPAd
Incidence of
atypical tubule
hyperplasia
Incidence
of CPN
0
2.1
2.1
0/50
49/50
28
2.0
1.7
0/50
43/50
121
2.0
1.8
0/50
45/50
542
2.4b
2.3
1/50
48/50
Dose
(mg/kg-d)
Incidence of
papillary
necrosis
Incidence of
papillary
mineralization
Incidence of urothelial
hyperplasia of the renal
pelvis
0
0/50
0/50
0/50
28
1/50
0/50
0/50
121
0/50
16/50b
10/50b
542
2/50
42/50b
25/50b
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-2. Evidence pertaining to kidney histopathology effects in animals
following exposure to ETBE (continued)
Reference and study design
Results
Suzuki etal. (2012); JPEC
Female
(2010a) (continued)
Dose
(mg/kg-d)
Average
severity of
CPN
Average severity of
CPN as calculated by
EPAd
Incidence of
atypical tubule
hyperplasia
Incidence
of 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
1.5b
1.2
2/50
39/50
Dose
(mg/kg-d)
Incidence of
papillary necrosis
Incidence of
papillary
mineralization
Incidence of
urothelial hyperplasia
of 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 the study authors.
bResult is statistically significant (p < 0.05) based on analysis of data by the study authors.
c4.18 mg/m3 = 1 ppm.
dAverage severity calculated as (grade x number of affected animals) -f total number of animals exposed.
-For controls, no response relevant; for other doses, no quantitative response reported.
Note: Percent change compared to controls calculated as 100 x [(treated value - control value) 4 control value].
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-3. Evidence pertaining to kidney biochemistry and urine effects in
animals following exposure to ETBE
Reference and study design
Results
JPEC (2008b)
Rat, Crl:CD(SD)
inhalation—vapor
male (10/group): 0,150, 500,
1,500, or 5,000 ppm
(0, 627, 2,090, 6,270, or
20,900 mg/m3);a female
(10/group): 0, 150, 500, 1,500,
or 5,000 ppm (0, 627, 2,090,
6,270, or 20,900 mg/m3)a
dynamic whole-body chamber;
6 h/d, 5 d/wk for 13 wk;
generation method, analytical
concentration, and method
reported
Male
Dose (mg/m3)
BUN (%)
Cholesterol (%)
Creatinine (%)
0
-
-
-
627
-9
8
-13
2,090
-5
9
-6
6,270
4
26
-
20,900
4
15
-3
Dose (mg/m3)
Proteinuria severity13
Proteinuria incidence
Urinary casts
0
0.5
3/6
0/6
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
Dose (mg/m3)
BUN (%)
Cholesterol (%)
Creatinine (%)
0
-
-
-
627
-5
7
0
2,090
3
9
3
6,270
-8
11
-9
20,900
-4
21
-9
Dose (mg/m3)
Proteinuria severity13
Proteinuria incidence
Urinary 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
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-3. Evidence pertaining to kidney biochemistry and urine effects in
animals following exposure to ETBE (continued)
Reference and study design
Results
Mivata et al. (2013); JPEC
(2008c)
Rat, Crl:CD(SD)
oral—gavage
male (15/group): 0, 5, 25,100,
or 400 mg/kg-d; female
(15/group): 0, 5, 25,100, or
400 mg/kg-d
daily for approximately 26 wk
Male
Dose (mg/kg-d)
BUN (%)
Cholesterol (%)
Creatinine (%)
0
-
-
-
5
12
-
0
25
1
21
-10
100
4
12
-3
400
8
53°
0
Dose (mg/kg-d)
Proteinuria
incidence
Proteinuria severity13
Urinary 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 (mg/kg-d)
BUN (%)
Cholesterol (%)
Creatinine (%)
0
-
-
-
5
-5
-7
-19
25
-7
-7
-12
100
-1
-2
-16
400
4
3
-16
Dose (mg/kg-d)
Proteinuria
incidence
Proteinuria severity13
Urinary 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
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-3. Evidence pertaining to kidney biochemistry and urine effects in
animals following exposure to ETBE (continued)
Reference and study design
Results
Saito et al. (2013); JPEC (2010b)
Rat, F344
inhalation—vapor
male (50/group): 0, 500,1,500,
or 5,000 ppm (0, 2,090, 6,270,
or 20,900 mg/m3);a female
(50/group): 0, 500,1,500, or
5,000 ppm (0, 2,090, 6,270, or
20,900 mg/m3)a
dynamic whole-body
inhalation; 6 h/d, 5 d/wk for
104 wk; generation method,
analytical concentration, and
method reported
Response relative to control:
Male
Dose
(mg/m3)
BUN (%)
Cholesterol (%)
Creatinine
(%)
Proteinuria
incidence
Proteinuria
severity13
0
-
-
-
44/44
3.7
2,090
41°
10
14
38/38
3.5
6,270
45°
29°
29°
40/40
3.6
20,900
179°
52°
71°
31/31
3.6
Female
Dose
(mg/m3)
BUN (%)
Cholesterol (%)
Creatinine
(%)
Proteinuria
incidence
Proteinuria
severity13
0
-
-
-
33/38
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°
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-3. Evidence pertaining to kidney biochemistry and urine effects in
animals following exposure to ETBE (continued)
Reference and study design
Results
Suzuki etal. (2012); JPEC
(2010a)
Rat, F344
oral—drinking water
male (50/group): 0, 625, 2,500,
or 10,000 ppm (0, 28,121, or
542 mg/kg-d);d female
(50/group): 0, 625, 2,500, or
10,000 ppm (0, 46,171, or
560 mg/kg-d)d
daily for 104 wk
Response relative to control:
Male
Dose
(mg/kg-d)
BUN (%)
Cholesterol (%)
Creatinine
(%)
Proteinuria
incidence
Proteinuria
severity15
0
-
-
-
39/39
3.0
28
3
-11
0
37/37
3.1
121
20°
10
17
34/34
3.1
542
43°
31°
17
35/35
3.1
Female
Dose
(mg/kg-d)
BUN (%)
Cholesterol (%)
Creatinine
(%)
Proteinuria
incidence
Proteinuria
severity13
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+")] t- total number of animals in group.
cResult is statistically significant (p < 0.05) based on analysis of data by study authors.
Conversion performed by study authors.
-For controls, no response relevant; for other doses, no quantitative response reported.
Note: Percent change compared to controls calculated as 100 x [(treated value - control value) -f control value].
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-4. Evidence pertaining to kidney tumor effects in animals following
exposure to ETBE
Reference and study design
Results
Saito et al. (2013); JPEC (2010b)
Rat, F344
inhalation—vapor
male (50/group): 0, 500,1,500, or
5,000 ppm (0, 2,090, 6,270, or
20,900 mg/m3);a female (50/group): 0,
500,1,500, or 5,000 ppm (0, 2,090,
6,270, or 20,900 mg/m3)a
Male
Female
Dose
(mg/m3)
Renal cell
carcinoma
Dose (mg/m3)
Renal cell
carcinoma
0
0/50
0
0/50
2,090
1/50
2,090
0/50
6,270
0/49
6,270
0/50
20,900
0/50
20,900
0/50
Suzuki et al. (2012); JPEC (2010a)
Rat, F344
oral—drinking water
male (50/group): 0, 625, 2,500, or
10,000 ppm (0, 28,121, or
542 mg/kg-d);b female (50/group): 0,
625, 2,500, or 10,000 ppm (0, 46, 171, or
560 mg/kg-d)c
daily for 104 wk
Male
Female
Dose
(mg/kg-d)
Renal cell
carcinoma
Dose (mg/kg-d)
Renal cell
carcinoma
0
0/50
0
0/50
28
0/50
46
0/50
121
0/50
171
0/50
542
1/50
560
1/50
Hagiwara et al. (2011); JPEC (2008d)
Rat, F344
oral—gavage
male (12/group): 0 or 1,000 mg/kg-d
daily for 23 wk
Male
Dose (mg/kg-d)
Renal transitional cell
carcinoma
Renal tubular adenoma
or carcinoma
0
0/12
0/12
1,000
0/12
0/12
Initiation Promotion Studies
Hagiwara et al. (2011); JPEC (2008d)
Rat, F344
oral—gavage
male (30/group): 0, 300, or
1,000 mg/kg-d
daily for 23 wk following a 4-wk tumor
initiation by DMBDDd
Male
Dose (mg/kg-d)
Renal tubular adenoma
or carcinoma
Renal transitional cell
carcinoma
0
11/30
1/30
300
6/30
0/30
1,000
13/30
2/30
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-4. Evidence pertaining to kidney tumor effects in animals following
exposure to ETBE (continued)
Reference and study design
Results
Hagiwara et al. (2015)
Rat, Wistar
oral—gavage
male (30/group): 0,100, 300, 500, or
1,000 mg/kg-d
daily for 19 wk following a 2-wk tumor
initiation by
W-ethyl-W-hydroxyethylnitrosamine
(EHEN)
Male
Dose (mg/kg-d)
Renal tubular adenoma or carcinoma0
0
18/30
100
23/30
300
25/30
500
26/30
1,000
26/30
a4.18 mg/m3 = 1 ppm.
bConversion performed by study authors.
cAuthors report significant trend.
dDiethylnitrosamine (DEN), W-butyl-W-(4-hydroxybutyl)nitrosamine (BBN), W-methyl-W-nitrosourea (MNU),
1,2-dimethylhydrazine dihydrochloride (DMH), and W-bis(2-hydroxypropyl)nitrosamine (DHPN).
Table 1-5. Comparison of nephropathy and urothelial hyperplasia in
individual male rats from 2-year oral exposure (IPEC. 2010a)
Urothelial hyperplasia
CPN
None
Minimal
Mild
Moderate
Marked
None
15
21
105
23
l
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, rho= 0.36.
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-6. Comparison of nephropathy and urothelial hyperplasia in
individual male rats from 2-year inhalation exposure (IPEC. 2010b)
Urothelial hyperplasia
CPN
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, rho = 0.36.
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Toxicological Review of Ethyl Tertiary Butyl Ether
¦ = exposures at which the en dpoint was reported statistically significant by study authors
~ = exposures at which the endpointwasreported not statistically significant by study authors
Increased Absolute PO Male rat; lGwks (B |
Kidney VVeiglit
PO Female ral; 17wks (B)
PO Male rat; 18wks (C)
PO Female ral; 18wks (C)
F'l Male rat; reproductive (C)
F1 Female rat; reproductive (C)
Male rat; 26wks (D)
Female rat; 26wks (D)
M ale rat; 104 wks (E)
Female rat; 104wks(E)
S B-
H B-
~—a—E3
B—B-
Incidence of Chronic
Progressive
Nephropathy
Average Severity of
Chronic Progressive
Nephropathy
Male rat; 104wks (A)
Female rat; 104wks (A)
Male rat; 104 wks (E)
Female rat; 104wks (E)
Male rat; 104 wks (A)
Female rat; 104wks (A)
Male rat; 104 wks (E)
Female rat; 104wks(E)
Urothelial Hyperplasia
of the Renal Pelvis
Male rat; 104wks (E)
Female rat; 104wks(E)
10 100 1.000
Dose (mg/kg-day]
10.000
Sources: (A) Cohen et al., 2011 reanalysis ofJPEC, 2010a; (B) Fujii et al., 2010; JPEC, 2008d; (C) Gaoua,
2004b; (D) Miyata et al, 2013; JPEC, 2008b; (E) Suzuki et al., 2012; JPEC, 2010a
Figure 1-4. Exposure-response array of kidney effects following oral exposure
to ETBE.
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Toxicological Review of Ethyl Tertiary Butyl Ether
¦ = exposures at which the endpoint was reported statistically significant by study authors
~ = exposures at which the endpoint was reported not statistically significant bystudy authors
Increased Absolute
Kidney Weight
Male rat; 13wits (A)
Female rat; 13wks (A)
Male rat; 13 wits (C)
Female rat; 13 wits (C)
Male mouse; 13 wks (B)
Female mouse; 13 wks(B)
Male rat; 104wks (D)
Female rat; 104wks(D)
Incidence of Chronic
Male rat; 104 wits (D)
Progressive
Nephropathy
Female rat; 104wks(D)
Average Severity of
Male rat; 104wks (D)
Chronic Progressive
Nephropathy
Female rat; 104wks (D)
Urothelial Hyperplasia
Male rat; 104wks (D)
of the Renal Pelvis
Female rat; 104wks(O)
100 1,000 10,000 100,000
Exposure Concentration (mg/m3)
Sources: (A) ] PEC, 2008a; (B) Medinsky et ai, 1999; Bond et al, 1996a (C) Medinsky et al., 1999; Bond et
al., 1996b (D) Saito et al.,2013; JPEC, 2"oiOb
Figure 1-5. Exposure-response array of kidney effects following inhalation
exposure to ETBE.
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Toxicological Review of Ethyl Tertiary Butyl Ether
Mode-of-Action Analysis—Kidney Effects
Toxicokinetic considerations relevant to kidney toxicity
ETBE is metabolized by cytochrome P450 (CYP450) enzymes to an unstable hemiacetal that
decomposes spontaneously into tert-butanol and acetaldehyde (Bernauer etal.. 1998).
Acetaldehyde is further metabolized in the liver. The main circulating breakdown product of ETBE
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 more detail in the draft IRIS assessment of
tert-butanol) also are relevant to evaluating the kidney effects observed after ETBE exposure. In
particular, tert-butanol, like ETBE, has been reported to cause nephrotoxicity in rats, including
effects associated with alpha 2u-globulin nephropathy
(https://cfpub.epa.gov/ncea/iris drafts/recordisplay.cfm?deid=2 62086). Unlike ETBE, however,
increased renal tumors, in the absence of an initiator, were reported following chronic drinking
water exposure to tert-butanol.
Alpha 2u-globulin-associated renal tubule nephropathy
One disease process to consider when interpreting kidney effects in rats is related to the
accumulation of alpha 2u-globulin protein. Alpha 2u-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 alpha 2u-globulin commonly detected in male rat urine and associated with renal tubule
nephropathy is considered specific to that sex and species. Exposure to chemicals that induce alpha
2u-globulin accumulation can initiate a sequence of histopathological events leading to kidney
tumorigenesis. Because alpha 2u-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 alpha 2u-globulin plays a role is important The
role of alpha 2u-globulin accumulation in the development of renal tubule nephropathy and
carcinogenicity observed following ETBE exposure was evaluated using the U.S. EPA f!991bl Risk
Assessment Forum Technical panel report, Alpha 2u-GIobuIin: Association with Chemically Induced
Renal Toxicity and Neoplasia in the Male Rat, as well as the IARC alpha 2u-globulin criteria fCapen et
al.. 19991. These frameworks provide 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 alpha 2u-globulin accumulation. The IARC alpha 2u-globulin criteria
(Capen etal.. 1999) are discussed following the EPA criteria (U.S. EPA. 1991a).
The hypothesized sequence of alpha 2u-globulin renal tubule nephropathy, as described by
U.S. EPA fl991al. is as follows. Chemicals that induce alpha 2u-globulin accumulation do so
rapidly. Alpha 2u-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
1-23
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Toxicological Review of Ethyl Tertiary Butyl Ether
feature of the mature male rat kidney; they are particularly evident in the S2 (P2) segment of the
proximal tubule and contain alpha 2u-globulin fU.S. EPA. 1991al. Abnormal increases in hyaline
droplets have more than one etiology and can be associated with the accumulation of different
proteins. As hyaline droplet 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 (see Figure 1-6).
U.S. EPA fl991al identified two questions that must be addressed to determine the extent
to which alpha 2u-globulin-mediated processes induce renal tubule nephropathy and
carcinogenicity. The first question is whether the alpha 2u-globulin process is occurring in male
rats and influencing renal tubule tumor development The second question is whether the renal
effects in male rats exposed to ETBE are due solely to the alpha 2u-globulin process.
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 present in the hyaline droplets in treated male rats is alpha 2u-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 1-11, and are evaluated below.
1-24
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Toxicological Review of Ethyl Tertiary Butyl Ether
TBA = te/t-butanol.
Source: Adapted from Swenberg and Lehman-McKeeman (1999); U.S. EPA (1991a).
Figure 1-6. Temporal pathogenesis of alpha 2u-globulin-associated
nephropathy in male rats. Alpha 2u-globulin synthesized in the livers of male
rats is delivered to the kidney, where it can accumulate in hyaline droplets
and be retained by epithelial cells lining the S2 (P2) segment of the proximal
tubules. Renal pathogenesis following continued exposure and increasing
droplet accumulation can progress stepwise from increasing epithelial cell
damage, death, and dysfunction, leading to the formation of granular casts in
the corticomedullary junction, and linear mineralization of the renal papilla,
in parallel with carcinogenesis of the renal tubular epithelium.
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Toxicological Review of Ethyl Tertiary Butyl Ether
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, or
5,000 ppm (0, 627, 2,090, 6,270, or
20,900 mg/m3);a female (10/group): 0,150,
500,1,500, or 5,000 ppm (0, 627, 2,090,
6,270, or 20,900 mg/m3)a
dynamic whole-body chamber; 6 h/d, 5 d/wk
for 13 wk; generation method, analytical
concentration, and method reported
Male
Dose (mg/m3)
Incidence of hyaline droplets in the
proximal tube epithelium
0
0/10
627
3/10
2,090
8/10b
6,270
8/10b
20,900
8/10b
Unspecified representative samples reported as "weakly positive"
for alpha 2u-globulin in males; no hyaline droplets observed in
proximal tubule of females; hyaline droplets positive for alpha
2u-globulin not examined in females.
JPEC (2008c); Mivata et al. (2013)
Rat, Crl:CD(SD)
oral—gavage
male (15/group): 0, 5, 25,100, or
400 mg/kg-d; female (15/group): 0, 5, 25,
100, or 400 mg/kg-d
daily for 180 d
Male
Female
Dose
mg/kg-d
Incidence
of hyaline
droplets
Incidence of
hyaline droplets
positive for alpha
2u-globulin
Dose
(mg/kg-d)
Incidence
of hyaline
droplets
0
0/15
0/1
0
0/15
5
0/15
-
5
-
25
0/15
-
25
-
100
4/15
2/2
100
-
400
10/15b
1/1
400
0/15
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-7. Additional kidney effects potentially relevant to mode of action in
animals exposed to ETBE (continued)
Reference and study design
Results
Medinskv et al. (1999); U.S. EPA (1997)
Rat, F344
inhalation—vapor
male (48/group): 0, 500,1,750, or
5,000 ppm
(0, 2,090, 7,320, or 20,900 mg/m3);a female
(48/group): 0, 500,1,750, or 5,000 ppm
(0, 2,090, 7,320, or 20,900 mg/m3)a
dynamic whole-body chamber; 6 h/d, 5 d/wk
for 13 wk; generation method, analytical
concentration, and method reported
Male
Proximal tubule proliferation (%)
Dose
mg/m3
Hyaline droplet
severity
1 wk
4 wk
13 wk
0
1.8
-
-
-
2,090
3.0
39
24
137b
7,320
3.2
23
-14
27b
20,900
3.8
102b
175b
171b
Female
Proximal tubule proliferation (%)
Dose (mg/m3)
1 wk
4 wk
13 wk
0
-
-
-
2,090
60b
3
73
7,320
_Q
00
00
15
64
20,900
49b
_Q
1
m
47
Saito et al. (2013); JPEC (2010b)
Rat, F344
inhalation—vapor
male (50/group): 0, 500,1,500, or
5,000 ppm
(0, 2,090, 6,270, or 20,900 mg/m3);a female
(50/group): 0, 500,1,500, or 5,000 ppm
(0, 2,090, 6,270, or 20,900 mg/m3)a
dynamic whole-body inhalation; 6 h/d,
5 d/wk for 104 wk; generation method,
analytical concentration, and method
reported
Male
No hyaline droplets observed.
Female
No hyaline droplets observed.
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-7. Additional kidney effects potentially relevant to mode of action in
animals exposed to ETBE (continued)
Reference and study design
Results
Suzuki et al. (2012); JPEC (2010a)
Rat, F344
oral—water
male (50/group): 0, 625, 2,500,10,000 ppm
(0, 28,121, 542 mg/kg-d);c female
(50/group): 0, 625, 2,500, 10,000 ppm (0, 46,
171, 560 mg/kg-d)c
daily for 104 wk
Male
No hyaline droplets observed.
Female
No hyaline droplets observed.
a4.18 mg/m3 = 1 ppm.
bResult is statistically significant (p < 0.05) based on analysis of data by the study authors.
Conversion performed by the study authors.
-For controls, no response relevant; for other doses, no quantitative response reported.
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-8. Summary of data informing whether the alpha 2u-globulin process
is 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
+
J PEC (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 alpha
2u-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)
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-8. Summary of data informing whether the alpha 2u-globulin
process is occurring in male rats exposed to ETBE (continued)
Criterion
Duration
Results
Reference
(d) Linear mineralization of
tubules in the renal papilla
13 wk
-
J PEC (2008b)
13 wk
-
Medinsky et al. (1999)
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)
+ = Statistically significant change reported in one or more treated groups.
(+) = Effect reported in one or more treated groups, but statistics not reported.
- = No statistically significant change reported in any of the treated groups.
aDroplet severity.
bUnspecified "representative samples" examined.
Three samples from highest two dose groups examined.
dLabeling index statistically significantly increased, but no hyperplasia reported.
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Toxicological Review of Ethyl Tertiary Butyl Ether
¦ = 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
9 = effect was observed but statistics not reported
+ = unspecified representative samples reported positive for a2u-globulin
Miyata et al, 2013; (PEC, 2008c - 26wks
Accumulation of
hyaline droplets
Suzuki et al, 2012; |PEC, 2010a - 104wks
~
B C
~
] ¦
¦a ~
a2u-gl°bulinin
hyaline Miyata et al, 2013; )PEC, 2008c - 26wks
droplets
H
1-
Cohen ct al, 2011 -13 wks
Granular Miyata et al., 2013; JPEC, 2008c - 26wks
casts/dilation
Suzuki et al, 2012; (PEC, 2010a • 104wks
~
0 E
~
•
1 o
-b a
Miyata et al, 2013; (PEC, 2008c - 26wks
Linear
papillary
mineralization
Suzuki et al, 2012; (PEC, 2010a • 104wks
G
B E
~
3 b
¦ ¦
Tubular Suzuki el aj 2012; JPEC, 2010a - 104wks
hyperplasia
~
¦b b
Renal
adenoma
Suzuki et al, 2012; |PEC, 2010a - 104wks
01"
carcinoma
~
¦B B
1 10 100 1,000
Dose (mg/kg-day)
Figure 1-7. ETBE oral exposure array of alpha 2u-globulin data in male rats.
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Toxicological Review of Ethyl Tertiary Butyl Ether
¦ = 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
Medinsky et al., 1999; Bond et al, 1996 - lwk
Medinsky et al., 1999; Bond ct al, 1996 - 4wk
Accumulation
of hyaline Medinsky et al., 1999; Bond et al., 1996 - 13wk
droplets
jPEC, 2008b -13wk
Saito et al., 2013; JPEC, 2010b - 104wk
~ B-
a2u-globulinin
hyaline
droplets
Medinsky et al„ 1999; Bond et al, 1996 - lwk
Medinsky et al., 1999; Bond et al, 1996 - 4wk
Medinsky et al., 1999; Bond et al, 1996 - 13wk
JPEC, 2008b-13 wk
-+ +-
Granular
casts/dilation
Medinsky ct al., 1999; Bond et al., 1996 ¦ 13wk
JPEC, 2008b -13wk
Saito et al., 2013; JPEC, 2010b - 104wk
~ B B-
B B-
-B
Linear
papillaiy
mineralization
Medinsky et al., 1999; Bond et al., 1996 - 13wk
JPEC, 2008b -13wk
Saito et al., 2013; JPEC, 2010b - 104wk
B B B-
B B-
Tubular
hyperplasia
Saito et at, 2013; JPEC, 2010b - 104wk
Renal adenoma
or carcinoma
Saito et al., 2013; JPEC, 2010b - 104wk
B B-
-B
100 1,000 10,000 100,000
Exposure Concentration (mg/m3)
Figure 1-8. ETBE inhalation exposure array of alpha 2u-globulin data in male
rats.
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Toxicological Review of Ethyl Tertiary Butyl Ether
Question One: is the alpha 2u-gIobuIin process occurring in male rats exposed to ETBE?7
(1) The first criterion to consider is whether hyaline droplets are larger and more
numerous in male rats. The accumulation of hyaline droplets was observed in all three subchronic
ETBE exposure studies, but was not observed in two chronic ETBE studies (see Table 1-7 and
Table 1-8). Failure to observe alpha 2u-globulin and increased droplet accumulation in the 2-year
studies is not unusual because alpha 2u-globulin naturally declines in males around 5 months of
age (U.S. EPA. 1991a). Accumulation of hyaline droplets in the proximal tubular epithelium of the
kidney was observed in male rats following 90-day inhalation exposure to 627, 2,090, 6,270, or
20,900 mg ETBE/m3 flPEC. 2008bl The increases atthe three highest concentrations were
statistically significant; however, none of the animals had hyaline droplet grades over 1 flPEC.
2008b). Severity grade of the hyaline droplets exhibited a dose-response after a 1-week exposure,
as indicated by scores of 1.2, 3.4, 4.0, and 4.6 at 0, 2,090, 7,320, and 20,900 mg ETBE/m3,
respectively, and 90 days of ETBE inhalation exposure increased the severity grades of hyaline
droplets from 1.8 in the control to 3.0, 3.2, and 3.8 (Medinskv et al.. 19991. In addition, the
incidence of hyaline droplets statistically significantly increased in a dose-related manner after
26 weeks of gavage exposure to 100 and 400 mg ETBE/kg-day fMivata etal.. 2013: TPEC. 2008cl.
These data indicate consistent evidence of hyaline droplets increasing both in a dose-responsive
manner and within the expected time frame. Therefore, the available data are sufficient to fulfill the
first criterion that hyaline droplets are increased in size and number in male rats.
(2) The second criterion to consider is whether the protein in the hyaline droplets in male
rats is alpha 2u-globulin. Immunohistological staining to ascertain the protein composition in the
hyaline droplets was performed only in ETBE exposure studies that observed accumulation of
hyaline droplets. At the two highest doses, Mivataetal. f20131: TPEC f2008cl identified hyaline
droplets as positive for alpha 2u-globulin in 2/2 and 1/1 animals that were tested for the presence
of alpha 2u-globulin. The other two studies also reported that unspecified samples were positive
for alpha 2u-globulin (TPEC. 2008b: Medinskv etal.. 1999). TPEC (2008b) reported that the samples
stained weakly positive for alpha 2u-globulin and that positive alpha 2u-globulin staining was
observed only in male rats. No statistical tests were performed on these results. The available
studies that tested for alpha 2u-globulin in hyaline droplets did not test a sufficient number of
samples within a dose group nor were enough dose groups tested for alpha 2u-globulin to perform
dose-response analysis. Therefore, the available data are minimally sufficient to fulfill the second
criterion for alpha 2u-globulin present in the hyaline droplets but do suggest that ETBE is a weak
inductor of alpha 2u-globulin.
(3) The third criterion considered is whether several (but not necessarily all) additional
steps in the histopathological sequence associated with alpha 2u-globulin nephropathy appear in
male rats in a manner consistent with the understanding of alpha 2u-globulin pathogenesis (refer
7If the chemical meets the criteria for Question One, then a second question is asked: are the renal effects in
male rats exposed to this chemical due solely to the alpha 2u-globulin process?
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Toxicological Review of Ethyl Tertiary Butyl Ether
to Table 1-8). Of the remaining five endpoints in the pathological sequence, only linear papillary
mineralization and granular casts were observed. Papillary mineralization typically appears at
chronic time points, occurring after exposures of 3 months up to 2 years fU.S. EPA. 1991al.
Statistically significant increases in the incidence of papillary mineralization was observed in both
2-year studies. In male rats, papillary mineralization increased in a dose-related manner following
oral ETBE exposure at concentrations of 0, 28,121, and 542 mg/kg-day, respectively (Suzuki etal..
2012: TPEC. 2010a) and 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 fMivata 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 papillary mineralization rates of only 0, 2,
and 12% [lowest to highest exposure concentration; Saito etal. f20131: TPEC f2010bl]. 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 fTPEC. 2008b. c; Medinskv et al.. 19991. This observation might not be inconsistent
with the hypothesized MOA because cell death and exfoliation has been observed to occur
as early as 5 days post exposure, peak at 3 weeks, and then decline to near background
levels by 4-5 weeks fKanerva et al.. 19871: this endpointwas not examined in any study
evaluating ETBE exposures less than 13 weeks. Thus, the lack of exfoliation observations
could be due to weak induction of alpha 2u-globulin and a lack of appropriately timed
examinations.
b) Granular cast formation was observed in one study. The TPEC (2007a) 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 et al.. 19991 despite using similar
exposure concentrations. Granular cast formation, however, might not occur with weak
inducers of alpha 2u-globulin (Short etal.. 1986). which is consistent with the weak staining
of alpha 2u-globulin, as discussed above (TPEC. 2008b).
c) Linear mineralization of tubules within the renal papilla was consistently observed in male
rats after 2 years fSaito etal.. 2013: Suzuki etal.. 20121. 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 less in females than in males. Observation of proliferation in
both sexes suggests that this effect is not male specific, and thus not solely due to alpha
2u-globulin. Furthermore, renal tubule hyperplasia was not observed in any 2-year study,
suggesting that ETBE does not induce sustained proliferation (Saito etal.. 2013: Suzuki et
al.. 2012). Renal tubule hyperplasia is the preneoplastic lesion associated with alpha
2u-globulin nephropathy in chronic exposures that leads to renal tubule tumors fU.S. EPA.
1991a).
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The progression of histopathological lesions for alpha 2u-globulin nephropathy is
predicated on the initial response of excessive hyaline droplet accumulation (containing alpha
2u-globulin) leading to cell necrosis and cytotoxicity, which in turn causes 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 alpha 2u-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 fMedinskv
etal.. 19991. and increased incidence was subsequently observed at 90 days flPEC. 2008bl or
26 weeks flPEC. 2008cl: alpha 2u-globulin was identified as the protein in these droplets (Borghoff
etal.. 2001: Williams and Borghoff. 20011. Observations of the subsequent linear mineralization of
tubules fall within the expected timeframe of the appearance of these lesions. Granular cast
formation was reported in one oral study (Cohen etal.. 20111 but not in three other oral and
inhalation studies (TPEC. 2008b. c; Medinskv et al.. 19991. which also could indicate weak alpha
2u-globulin induction. Neither alpha 2u-globulin-mediated regenerative cell proliferation nor
atypical renal tubule hyperplasia were observed. Lack of necrosis and exfoliation might be due to
the weak induction of alpha 2u-globulin and a lack of appropriately timed examinations.
Hyaline droplets were weakly induced in all male rats in the 13-week inhalation studies
(TPEC. 2008b: Medinskv etal.. 1999) but did not result in increased linear mineralization at the
corresponding doses. The lack of increased linear mineralization at low doses is also 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 time frame
of detectability. Furthermore, no explicit inconsistencies are present in the temporal appearance of
the histopathological lesions associated with the alpha 2u-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. Therefore, the number of histopathological steps
observed was insufficient to fulfill the third criterion.
Summary and conclusions for Question One, Is the alpha 2u-gIobuIin process occurring in male rats
exposed to ETBE?
The evidence suggests that ETBE causes hyaline droplets to increase in size and number.
The documentation of alpha 2u-globulin staining is poor and provides weak evidence of alpha
2u-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 chronic bioassays,
exposure routes, and durations ranging from 13 weeks to 2 years. Overall, the available data were
insufficient to conclude that the alpha 2u-globulin process is exclusively operative.
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Consideration of additional IARC (19991 criteria
An alpha 2u-globulin framework was published by IARC in 1999 fCapen etal.. 19991. See
Table 1-9 for the criteria laid out in the IARC consensus document.
Table 1-9. IARC criteria for an agent causing kidney tumors through an alpha
2u-globulin associated response in male rats
• Lack of genotoxic activity (agent and/or metabolite) based on an overall evaluation of in vitro and in vivo
data.
• Male rat specificity for nephropathy and renal tumorigenicity.
• Induction of the characteristic sequence of histopathological changes in shorter term studies, of which
protein droplet accumulation is obligatory.
• Identification of the protein accumulating in tubule cells as alpha 2u-globulin.
• Reversible binding of the chemical or metabolite to alpha 2u-globulin.
• Induction of sustained increased cell proliferation in the renal cortex.
• Similarities in dose-response relationship of the tumor outcome with the histopathological endpoints
(protein droplets, alpha 2u-globulin accumulation, cell proliferation).
The EPA and IARC criteria differ in a few minor ways. The EPA framework requires the
observation of several (but not necessarily all) additional steps in the histopathological sequence
associated with alpha 2u-globulin nephropathy, whereas IARC requires the "induction of the
characteristic sequence of histopathological changes in shorter term studies, of which protein
droplet accumulation is obligatory," but IARC doesn't specify which or how many of the additional
histopathological changes must be observed to consider this criteria met In addition, the IARC
criteria have specific criteria pertaining to lack of genotoxicity of parent compound/metabolite and
male rat specificity for nephropathy and renal tumorigenicity whereas the EPA framework
considers these data as supplemental information (see Part 4, XVIIB. Additional Information Useful
for the Analysis). These additional criteria required by IARC (Capen etal.. 19991 are discussed
below.
Lack of genotoxic action
As discussed in Appendix Section B.2.2, limited data are available to help inform the
genotoxic potential of ETBE. Most studies indicate that ETBE does not induce genotoxicity in the
systems tested, although several key types of assays are missing (e.g., studies investigating
chromosomal aberrations and deoxyribonucleic acid [DNA] adducts). While some data suggest the
major metabolite of ETBE, tert-butanol, could be genotoxic, the overall evidence is inadequate to
establish a conclusion. However, ETBE's metabolite acetaldehyde has induced sister chromatid
exchanges in Chinese hamster ovary cells, gene mutations in mouse lymphomas, and DNA strand
breaks in human lymphocytes flARC. 1999al. In addition, increased genotoxicity of ETBE is noted
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Toxicological Review of Ethyl Tertiary Butyl Ether
when tested in animals with polymorphisms in the acetaldehyde dehydrogenase gene ALDH2,
which decreases the ability to metabolize acetaldehyde fWeng etal.. 2019: Weng etal.. 2014: Weng
etal.. 2013: Wang etal.. 2012: Weng etal.. 2012: Weng etal.. 20111. Approximately 8% of the
world's population carries this variant f Gross etal.. 20151. Overall, the IARC criterion for lack of
genotoxic activity has been weakly met for nonsusceptible populations but not met for the subset of
the population that cannot efficiently detoxify acetaldehyde.
Male rat specificity for nephropathy
Evidence is limited for ETBE-mediated kidney effects in other species. Only one subchronic
study in wild-type (WT) mice evaluated kidney effects fMedinskv et al.. 1999: Bond etal.. 19961. No
chronic studies are available that evaluated kidney effects in mice.
Increased absolute kidney weight and increased severity of chronic nephropathy was noted
in both sexes of rats exposed chronically to ETBE through inhalation and in drinking water (TPEC.
2010a. b). Changes in clinical chemistry suggestive of kidney toxicity (e.g., increased BUN,
cholesterol, and protein urea) were also noted in both male and female rats. However, dose-related
increased incidence of urothelial hyperplasia of the renal pelvis was observed in male rats in
chronic oral and inhalation bioassays but was not found in female rats by either route of exposure.
In summary, while male rats appear to be more sensitive to ETBE-meditated kidney toxicity,
indications of nephropathy were also observed in female rats. Therefore, this criterion has not
been met
Comparison of ETBE and tert-butanol alpha 2u-gIobuIin data
Both EPA and IARC have accepted the biological plausibility of the alpha
2u-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 has demonstrated that several steps in the sequence of pathological
events are not required for tumor development. The study showed that several alpha
2u-globulin-inducing chemicals fail to induce many of the pathological sequences in the alpha
2u-globulin pathway (Doi etal.. 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 alpha 2u-globulin process. These
chemicals included 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 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 alpha 2u-globulin sequence, as demonstrated by
SS IICA, which induced some of the most severe nephropathy relative to the other chemicals, but
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did not significantly increase kidney tumors (Doi 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 are also potentially relevant to evaluating the MOA of ETBE. In particular,
the effects of tert-butanol on the alpha 2u-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 increased in a dose-response manner. ETBE exposure increased hyaline droplets
at lower internal concentrations 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 (Saito etal.. 2013: Suzuki etal.. 2012: TPEC. 2010b). The failure of ETBE to induce several
histopathological lesions in the alpha 2u-globulin pathological sequence, including renal tubule
hyperplasia and tumors, at similar internal tert-butanol concentrations (as predicted by the Salazar
PBPK model) as those that induced hyperplasia and tumorigenesis following exposure to
tert-butanol suggests remaining uncertainty regarding toxicokinetic or toxicodynamic differences,
or both, across the two chemicals.
Chronic progressive nephropathy (CPN)
Exacerbation of CPN has been proposed as another rat-specific mechanism of
nephrotoxicity that is not relevant to humans (Hard etal.. 2009). CPN is an age-related renal
disease that occurs in rats of both sexes fNTP. 2015: Seelv and Brix. 2014a: Hard etal.. 2013:
Melnick etal.. 2012: U.S. EPA. 1991a). CPN is more severe in males than in females and is
particularly common in the S-D and F344 strains. Dietary and hormonal factors play a role in
modifying CPN, though its etiology is 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.. 2013).
However, an analysis of 60 NTP carcinogenicity studies did not find an association between
chemically exacerbated CPN and renal tubule tumor induction in rats (Melnick et al.. 2012). ETBE
exposure increased CPN severity following 2-year inhalation and 13-week oral exposure but, unlike
tert-butanol, did not affect tubule hyperplasia or increase renal tubule tumor incidence.
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.. 2009). 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
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Toxicological Review of Ethyl Tertiary Butyl Ether
when CPN was increased as in the 2-year inhalation exposure study (Saito 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—alpha 2u-globulin-related hyaline droplet nephropathy fWebb etal.. 19901.
Additionally, renal effects of alpha 2u-globulin accumulation can exacerbate the effects associated
with CPN (Travlos etal.. 2011: U.S. EPA. 1991a).
CPN is often more severe in males than in females. While background severity of CPN in
controls was higher in male rats, increased severity of CPN was reported in both male and female
rats with ETBE exposure and was statistically significant at the highest exposure groups of both
sexes following chronic inhalation (see Table 1-2). 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 (but not female) rats in both the inhalation
and oral studies (Saito 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 (see Table 1-5
and Table 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 urothelial hyperplasia is not related to ETBE treatment, given
the robust dose-response relationship in male rats treated with ETBE.
Finally, because tert-butanol is a major metabolite of ETBE 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 using PBPK modeling 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;
however, contributing toxicity from the parent compound and other metabolites cannot be ruled
out
Overall Conclusion on Mode of Action (MOA) for Kidney Effects
ETBE increases alpha 2u-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 alpha 2u-globulin nephropathy; however, the observation of alpha 2u-globulin
accumulation in hyaline droplets with ETBE exposure makes the human relevance of the associated
nephropathy in male rat CPN uncertain, and the exacerbation of CPN could play a role in the
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Toxicological Review of Ethyl Tertiary Butyl Ether
observed nephropathy. Currently, the MOA for the observed ETBE-induced exacerbation of CPN is
unknown, especially in female rats and in male and female mice, which are not affected by alpha
2u-globulin nephropathy. Collectively, the evidence indicates other, unknown processes contribute
to the observed nephrotoxicity following ETBE exposure, particularly in female rats.
Integration of Kidney Effects
Kidney effects (increases in severity of nephropathy, increased kidney weight, alterations in
clinical chemistry, hyaline droplets, linear mineralization, and urothelial hyperplasia of the renal
pelvis) were observed across multiple studies, predominantly in rats; chronic bioassays found no
treatment-related increases in renal tumors. The available evidence indicates that multiple
processes induce the noncancer kidney effects.
Some endpoints in male rats (hyaline droplets, linear mineralization) are components of the
alpha 2u-globulin process. U.S. EPA (1991a) states that"[i]f a compound induces alpha 2u-globulin
accumulation in hyaline droplets, the associated nephropathy in male rats is not an appropriate
endpoint to determine noncancer (systemic) effects potentially occurring in humans." Therefore, in
the case of ETBE exposure, endpoints directly associated with alpha 2u-globulin processes were not
considered an indication of human health hazard for noncancer kidney toxicity. Because alpha
2u-globulin nephropathy is strictly a male rat phenomenon, dose-related kidney effects in female
rats and mice are not confounded by alpha 2u-globulin nephropathy.
It has been observed that chemicals that bind to alpha 2u-globulin can lead to increased
incidence and/or severity of CPN (Frazier etal.. 2012: Travlos etal.. 2011: U.S. EPA. 1991a). CPN is
a common and well-established constellation of age-related lesions in the kidney of male and
female rats, with no known counterpart to 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 (e.g., tubular degeneration, thickening of basement membranes, glomerular
sclerosis, etc.) also occur in the human kidney (Lusco etal.. 2016: Zojaetal.. 2015: Frazier etal..
2012: Abrass. 2000). Although CPN has no known analogue in the aging human kidney (NIEHS.
2019: Hard etal.. 2009). the etiology is unknown (NIEHS. 2019: Hard and Khan. 2004: Peter etal..
19861. Given that there is no definitive pathogenesis for CPN, it cannot be ruled out that a chemical
which exacerbates CPN in rats could also exacerbate existing disease processes in the human
kidney fNIEHS. 20191. Therefore, increased incidence of kidney effects with ETBE exposure in the
female rat (including increased kidney weight and increased severity of CPN) are considered
relevant to humans.
Several noncancer endpoints were concluded to result from ETBE exposure including
increased absolute kidney weight, histopathological changes, and alterations in clinical chemistry in
male and female rats, with the effects in males tending to be more severe than in females
(potentially due to confounding by alpha 2u-globulin processes). A PBPK model-based analysis
yielded consistent dose-response relationships between kidney weight, urothelial hyperplasia, and
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Toxicological Review of Ethyl Tertiary Butyl Ether
chronic progressive nephropathy (CPN) using tert-butanol blood concentration as the dose metric,
consistent with the hypothesis that tert-butanol mediates the noncancer kidney effects following
ETBE administration fSalazar etal.. 20151. Based on dose-related increases in these noncancer
endpoints in rats, kidney effects are a potential human hazard of ETBE 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 the 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 the 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. Increased liver weight was observed with ETBE exposure in male and female
rats treated for various durations orally or by inhalation. Several factors associated with the 2-year
organ-weight data could confound consideration for hazard identification. Proliferative lesions
(altered hepatocellular foci) were observed in rat livers, especially males, in both 2-year oral and
inhalation studies, which 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 (Saito etal.. 2013:
TPEC. 2010b). Organ-weight data obtained from studies of shorter duration, however, are less
complicated by these age-associated factors (e.g., tumors, mortality), and therefore, could be
appropriate for hazard identification.
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 1-10). 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..
2004): thus, only relative liver weight is considered in determining 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
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dose following oral exposures of 16 weeks or longer (Mivata etal.. 2013: Fujii etal.. 2010: TPEC.
2008c: Gaoua. 2004bl In utero exposure yielded similar effects on F1 liver weights, in terms of the
magnitude of percent change, from adult exposure f Gaoua. 2004bl 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
(TPEC. 2008b). Short-term studies observed similar effects on liver weight (TPEC. 2008a: White et
al„ 19951.
Liver histopathology. Centrilobular hypertrophy and acidophilic (eosinophilic) and
basophilic focal lesions were the only dose-related types of pathological lesions observed in the
liver. Centrilobular hypertrophy was inconsistently increased throughout the evidence base but
was also 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 1-11; Figure 1-9,
Figure 1-10). A 26-week gavage study (Mivata etal.. 2013: TPEC. 2008c) in rats and three 13-week
inhalation studies in mice and rats (Weng etal.. 2012: TPEC. 2008b: Medinskv etal.. 1999) showed a
statistically significant increase in centrilobular hypertrophy at the highest dose. In addition,
2-year oral and inhalation studies in rats reported increased liver weight in males and females.
Acidophilic (eosinophilic) and basophilic preneoplastic lesions were increased in male, but
not female, rats at the highest tested dose following a 2-year inhalation exposure to ETBE fSaito et
al.. 2013: TPEC. 2010b). Following 2-year drinking water exposure to ETBE, an increasing, but not
statistically significant incidence in basophilic preneoplastic lesions (14/50,18/50, 20/50, 22/50)
was observed in the liver of male rats, whereas 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 1-12; Figure 1-9, Figure 1-10). No enzyme levels were affected in studies of
exposure durations less than 2 years f Mivata etal.. 2013: TPEC. 2008bl. y-Glutamyl transpeptidase
(GGT) was significantly increased in male rats at one intermediate dose following oral exposure and
at the two highest doses following inhalation exposure in 2-year studies (TPEC. 2010a. b). GGT
levels were not significantly affected in female rats in any study. No consistent dose-related
changes were observed in aspartate aminotransferase (AST), alanine aminotransferase (ALT), or
alkaline phosphatase (ALP) liver enzymes following either oral or inhalation exposure of any
duration. 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 1-13. Liver
tumors were statistically significantly increased in male F344 rats at the high dose, but not in
females, following 2-year inhalation exposure (Saito etal.. 2013: TPEC. 2010b). The incidence of
combined adenomas and carcinomas of 0/50, 2/50,1/50, and 10/50 at 0, 2,090, 6,270, and
20,900 mg/m3, respectively, also resulted in a statistically significant, positive exposure-response
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Toxicological Review of Ethyl Tertiary Butyl Ether
trend (Peto's test p < 0.001). In this study, preneoplastic lesions, acidophilic (eosinophilic) and
basophilic foci, were also increased in male rats following a similar exposure pattern.
At the highest exposure dose in male rats, the dose at which the majority of liver tumors
were observed, a 25% reduction in body weight was seen, raising some question as to whether the
liver tumors observed in the highest exposure group in male rats are solely the result of excessive
toxicity rather than carcinogenicity of the tested agent. EPA's 2005 Cancer Guidelines discuss the
determination of an "excessively high dose" as compared to an "adequately high dose" and describe
the process as one of expert judgment which requires that "...adequate data demonstrate that the
effects are solely the result of excessive toxicity rather than carcinogenicity of the tested agent"
fU.S. EPA. 2005al. In the case of the 2-year inhalation study, the study authors did not report any
overt toxicity or altered toxicokinetics at the high dose flPEC. 2010b], In addition, the high-dose
female rats had a similar reduction in body weight (22%) with no liver tumors (or increase in
preneoplastic foci) observed.
No significant increase in tumors was observed following two chronic oral bioassays
(Suzuki etal.. 2012: TPEC. 2010a: Maltoni et al.. 1999). However, one bioassay (Maltoni etal.. 1999)
was confounded by extremely low survival in the controls (25-28% at 2 years), potentially due to
widespread respiratory infections (see discussion in Section 1.2.5). This extreme reduction in
survival likely affects this study's power to detect potential carcinogenicity. The other available
2-year oral cancer bioassay, which was well designed, conducted, and reported, did not show
significant increases in liver tumors; however, it did show an increased, but not statistically
significant, trend in basophilic preneoplastic lesions in the liver of male rats, with the incidence of
these lesions being less in female rats (Suzuki etal.. 2012: TPEC. 2010a).
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 of ETBE exposure via gavage fHagiwara etal.. 2015:
Hagiwara etal.. 2011). Liver tumors were not observed in male F344 rats exposed to ETBE for
23 weeks (n = 12) withoutprior mutagen exposure (Hagiwara etal.. 2011). Liver tumorigenesis
without prior mutagen exposure was not evaluated in Wistar rats (Hagiwara etal.. 2015).
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Table 1-10. Evidence pertaining to liver weight effects in animals exposed to
ETBE
Reference and study design
Results
Fuiiietal. (2010); JPEC (2008e)
Rat, S-D
oral—gavage
P0, male (24/group): 0,100, 300, or
1,000 mg/kg-d
daily for 16 wk beginning 10 wk prior to mating
P0, female (24/group): 0,100, 300, or
1,000 mg/kg-d
daily for 17 wk beginning 10 wk prior to mating
to LD 21
Response relative to control:
P0, Male
P0, Female
Dose
(mg/kg-d)
Relative
weight (%)
Dose
(mg/kg-d)
Relative
weight (%)
0
-
0
-
100
1
100
-1
300
2
300
3
1,000
21a
1,000
9a
Gaoua (2004b)
Rat, S-D
oral—gavage
P0, male (25/group): 0, 250, 500, or
1,000 mg/kg-d
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, or
1,000 mg/kg-d
daily for a total of 18 wk beginning 10 wk before
mating until PND 21
Fl, male (25/group): 0, 250, 500, or
1,000 mg/kg-d
F0 dams dosed daily through gestation and
lactation, then Fl doses beginning PND 22 until
weaning of the F2 pups
Fl, female (24-25/group): 0, 250, 500, or
1,000 mg/kg-d
F0 dams dosed daily through gestation and
lactation, then Fl dosed beginning PND 22 until
weaning of F2 pups
Response relative to control:
P0, Male
P0, Female
Dose
(mg/kg-d)
Relative
weight (%)
Dose
(mg/kg-d)
Relative
weight (%)
0
-
0
-
250
3
250
10
500
6
500
8
1,000
24a
1,000
4
Fl, Male
Fl, Female
Dose
(mg/kg-d)
Relative
weight (%)
Dose
(mg/kg-d)
Relative
weight (%)
0
-
0
-
250
0
250
3
500
ir
500
6
1,000
25a
1,000
9a
Hagiwara et al. (2011); JPEC (2008d)
Rat, F344
oral—gavage
male (12/group): 0 or 1,000 mg/kg-d
daily for 23 wk
Response relative to control:
Male
Dose
(mg/kg-d)
Relative weight (%)
0
-
1,000
27a
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-10. Evidence pertaining to liver-weight effects in animals exposed to
ETBE (continued)
Reference and study design
Results
JPEC (2008b)
Rat, Crl:CD(SD)
inhalation—vapor
male (NR): 0,150, 500,1,500, or 5,000 ppm (0,
627, 2,090, 6,270, or 20,900 mg/m3);b female
(NR): 0, 150, 500, 1,500, or 5,000 ppm (0, 627,
2,090, 6,270, or 20,900 mg/m3)
dynamic whole-body chamber; 6 h/d, 5 d/wk
for 13 wk; generation method, analytical
concentration, and method reported
Response relative to control:
Male
Female
Dose
(mg/m3)
Relative
weight (%)
Dose
(mg/m3)
Relative
weight (%)
0
-
0
-
627
5
627
4
2,090
5
2,090
-1
6,270
5
6,270
6
20,900
10
20,900
18a
JPEC (2008b)
Rat, Crl:CD(SD)
inhalation—vapor
male (6/group): 0 or 5,000 ppm (0 or
20,900 mg/m3);b female (6/group): 0 or
5,000 ppm (0 or 20,900 mg/m3)b
dynamic whole-body chamber; 6 h/d, 5 d/wk
for 13 wk followed by a 28-d recovery period;
generation method, analytical concentration,
and method reported
Response relative to control:
Male
Female
Dose
(mg/m3)
Relative
weight (%)
Dose
(mg/m3)
Relative
weight (%)
0
-
0
-
20,900
9a
20,900
7
Mivata et al. (2013); JPEC (2008c)
Rat, Crl:CD(SD)
oral—gavage
male (15/group): 0, 5, 25,100, or 400 mg/kg-d;
female (15/group): 0, 5, 25,100, or 400 mg/kg-d
daily for 26 wk
Response relative to control:
Male
Female
Dose
(mg/kg-d)
Relative
weight (%)
Dose
(mg/kg-d)
Relative
weight (%)
0
-
0
-
5
5
5
1
25
7
25
1
100
9
100
4
400
17a
400
12a
LD = lactation day; NR = not reported.
aResult is statistically significant (p < 0.05) based on analysis of data by the study authors.
Conversion performed by the study authors.
b4.18 mg/m3 = 1 ppm.
-For controls, no response relevant; for other doses, no quantitative response reported.
Note: Percent change compared with controls calculated as 100 x [(treated value - control value) -f control
value].
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Table 1-11. Evidence pertaining to liver histopathology effects in animals
exposed to ETBE
Reference and study design
Results
Gaoua (2004b)
Rat, S-D
oral—gavage
P0, male (25/group): 0, 250, 500, or
1,000 mg/kg-d
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, or
1,000 mg/kg-d
daily for a total of 18 wk beginning 10 wk before
mating until PND 21
P0, Male
P0, Female
Dose
(mg/kg-d)
Incidence of
centrilobular
hypertrophy
Dose
(mg/kg-d)
Incidence of
centrilobular
hypertrophy
0
0/25
0
0/25
250
0/25
250
0/25
500
0/25
500
0/25
1,000
3/25
1,000
0/25
JPEC (2008b)
Rat, Crl:CD(SD)
inhalation—vapor
male (NR): 0,150, 500,1,500, or 5,000 ppm (0,
627, 2,090, 6,270, or 20,900 mg/m3);a female
(NR): 0, 150, 500, 1,500, or 5,000 ppm (0, 627,
2,090, 6,270, or 20,900 mg/m3)
dynamic whole-body chamber; 6 h/d, 5 d/wk for
13 wk; generation method, analytical
concentration, and method reported
Male
Female
Dose
(mg/m3)
Incidence of
centrilobular
hypertrophy
Dose
(mg/m3)
Incidence of
centrilobular
hypertrophy
0
0/10
0
0/10
627
0/10
627
0/10
2,090
0/10
2,090
0/10
6,270
0/10
6,270
0/10
20,900
4/10b
20,900
6/10b
JPEC (2008b)
Rat, Crl:CD(SD)
inhalation—vapor
male (6/group): 0 or 5,000 ppm (0 or
20,900 mg/m3);a female (6/group): 0 or
5,000 ppm (0 or 20,900 mg/m3)a
dynamic whole-body chamber; 6 h/d, 5 d/wk for
13 wk followed by a 28-d recovery period;
generation method, analytical concentration, and
method reported
Male
Female
Dose
(mg/m3)
Incidence of
centrilobular
hypertrophy
Dose
(mg/m3)
Incidence of
centrilobular
hypertrophy
0
0/6
0
0/6
20,900
0/6
20,900
0/6
Medinskv et al. (1999); U.S. EPA (1997)
Rat, F344
inhalation—vapor
male (48/group): 0, 500,1,750, or 5,000 ppm (0,
2,090, 7,320, or 20,900 mg/m3);a female
(48/group): 0, 500,1,750, or 5,000 ppm (0, 2,090,
7,320, or 20,900 mg/m3);a
dynamic whole-body chamber; 6 h/d, 5 d/wk for
13 wk; generation method, analytical
concentration, and method reported
Male
Female
Dose
(mg/m3)
Incidence of
centrilobular
hypertrophy
Dose
(mg/m3)
Incidence of
centrilobular
hypertrophy
0
0/11
0
0/10
2,090
0/11
2,090
0/11
7,320
0/11
7,320
0/11
20,900
0/11
20,900
0/11
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Table 1-11. Evidence pertaining to liver histopathology effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Medinskv et al. (1999); Bond et al. (1996)
Mice, CD-I
inhalation—vapor
male (40/group): 0, 500,1,750, or 5,000 ppm (0,
2,090, 7,320, or 20,900 mg/m3);a female
(40/group): 0, 500,1,750, or 5,000 ppm (0, 2,090,
7,320, or 20,900 mg/m3)a
dynamic whole-body chamber; 6 h/d, 5 d/wk for
13 wk; generation method, analytical
concentration, and method reported
Male
Female
Dose
(mg/m3)
Incidence of
centrilobular
hypertrophy
Dose
(mg/m3)
Incidence of
centrilobular
hypertrophy
0
0/15
0
0/13
2,090
0/15
2,090
2/15
7,320
2/15
7,320
1/15
20,900
8/10b
20,900
9/14b
Mivata et al. (2013); JPEC (2008c)
Rat, Crl:CD(SD)
oral—gavage
male (15/group): 0, 5, 25,100, or 400 mg/kg-d;
female (15/group): 0, 5, 25,100, or 400 mg/kg-d
daily for 26 wk
Male
Female
Dose
(mg/kg-d)
Incidence of
centrilobular
hypertrophy
Dose
(mg/kg-d)
Incidence of
centrilobular
hypertrophy
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
6/15b
400
6/15b
Saito et al. (2013); JPEC (2010b)
Rat, F344
inhalation—vapor
male (50/group): 0, 500,1,500, or 5,000 ppm (0,
2,090, 6,270, or 20,900 mg/m3);a female
(50/group): 0, 500,1,500, or 5,000 ppm (0, 2,090,
6,270, or 20,900 mg/m3)a
dynamic whole-body inhalation; 6 h/d, 5 d/wk for
104 wk; generation method, analytical
concentration, and method reported
Male
Dose
(mg/m3)
Acidophilic
foci in liver
Basophilic
foci in liver
Bile duct
hyperplasia
Centrilobular
hypertrophy
0
31/50
18/50
48/50
0/50
2,090
28/50
10/50
44/50
0/50
6,270
36/49
13/49
46/49
0/49
20,900
39/50b
33/50b
41/50
0/50
Female
Dose
(mg/m3)
Acidophilic
foci in liver
Basophilic
foci in liver
Bile duct
hyperplasia
Centrilobular
hypertrophy
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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Table 1-11. Evidence pertaining to liver histopathology effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Suzuki et al. (2012); JPEC (2010a)
Rat, F344
oral—water
male (50/group): 0, 625, 2,500, or 10,000 ppm (0,
28,121, or 542 mg/kg-d);c female (50/group): 0,
625, 2,500, or 10,000 ppm (0, 46, 171, or
560 mg/kg-d)c
daily for 104 wk
Male
Dose
(mg/kg-d)
Acidophilic
foci in liver
Basophilic
foci in liver
Bile duct
hyperplasia
Centrilobular
hypertrophy
0
14/50
14/50
49/50
0/50
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
(mg/kg-d)
Acidophilic
foci in liver
Basophilic
foci in liver
Bile duct
hyperplasia
Centrilobular
hypertrophy
0
2/50
36/50
1/50
0/50
46
2/50
25/50b
4/50
0/50
171
1/50
31/50
4/50
0/50
560
0/50
30/50b
3/50
0/50
Weng et al. (2012)
Mice, C57BL/6
inhalation—vapor
male (5/group): 0, 500,1,750, or 5,000 ppm (0,
2,090, 7,320, or 20,900 mg/m3);a female
(5/group): 0, 500,1,750, or 5,000 ppm (0, 2,090,
7,320, or 20,900 mg/m3)a
dynamic whole-body chamber, 6 h/d, 5 d/wk for
13 wk; generation methods not reported, but
analytical methods (gas chromatograph) and
concentration reported
Male
Female
Dose
(mg/m3)
Incidence of
centrilobular
hypertrophy
Dose
(mg/m3)
Incidence of
centrilobular
hypertrophy
0
1/5
0
0/5
2,090
0/5
2,090
0/5
7,320
0/5
7,320
1/5
20,900
5/5b
20,900
5/5b
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Table 1-11. Evidence pertaining to liver histopathology effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Weng etal. (2012)
Mice, Aldh2 KO
inhalation—vapor
male (5/group): 0, 500,1,750, or 5,000 ppm (0,
2,090, 7,320, or 20,900 mg/m3);a female
(5/group): 0, 500,1,750, or 5,000 ppm (0, 2,090,
7,320, or 20,900 mg/m3)a
dynamic whole-body chamber, 6 h/d, 5 d/wk for
13 wk; generation methods were not reported,
but analytical methods (gas chromatograph) and
concentration reported
Male
Female
Dose
(mg/m3)
Incidence of
centrilobular
hypertrophy
Dose
(mg/m3)
Incidence of
centrilobular
hypertrophy
0
0/5
0
0/5
2,090
3/5
2,090
0/5
7,320
2/5
7,320
0/5
20,900
5/5b
20,900
4/5b
NR = not reported; PND = postnatal day.
a4.18 mg/m3 = 1 ppm.
bResult is statistically significant (p < 0.05) based on analysis of data by the study authors.
Conversion performed by study authors.
-For controls, no response relevant; for other doses, no quantitative response reported.
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Table 1-12. 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, or 5,000 ppm
(0, 627, 2,090, 6,270, or 20,900 mg/m3);a
female (NR): 0,150, 500,1,500, or 5,000 ppm
(0, 627, 2,090, 6,270, or 20,900 mg/m3)
dynamic whole-body chamber; 6 h/d, 5 d/wk
for 13 wk; generation method, analytical
concentration, and method reported
Response relative to control:
Male
Dose (mg/m3)
ALT (%)
ALP (%)
AST (%)
GGT (%)
0
-
-
-
-
627
9
13
3
11
2,090
0
12
1
0
6,270
5
-12
-7
11
20,900
12
-9
4
-100
Female
Dose (mg/m3)
ALT (%)
ALP (%)
AST (%)
GGT (%)
0
-
-
-
-
627
-
-3
2
25
2,090
11
-12
-95
12
6,270
-5
-7
12
25
20,900
26
5
0
25
Mivata et al. (2013); JPEC (2008c)
Rat, Crl:CD(SD)
oral—gavage
male (15/group): 0, 5, 25,100, or
400 mg/kg-d; female (15/group): 0, 5, 25,
100, or 400 mg/kg-d
daily for 180 d
Response relative to control:
Male
Dose
(mg/kg-d)
ALT (%)
ALP (%)
AST (%)
GGT (%)
0
-
-
-
-
5
10
2
1
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
1-50
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-12. Evidence pertaining to liver biochemistry effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Saito et al. (2013); JPEC (2010b)
Rat, F344
inhalation—vapor
male (50/group): 0, 500,1,500, or 5,000 ppm
(0, 2,090, 6,270, or 20,900 mg/m3);a female
(50/group): 0, 500,1,500, or 5,000 ppm (0,
2,090, 6,270, or 20,900 mg/m3)a
dynamic whole-body inhalation; 6 h/d,
5 d/wk for 104 wk; generation method,
analytical concentration, and method
reported
Response relative to control:
Male
Dose
(mg/m3)
ALT (%)
ALP (%)
AST (%)
GGT (%)
0
-
-
-
-
2,090
53
0
29
33
6,270
-3
-21b
-16
_Q
o
LO
20,900
24
-5
-2b
200b
Female
Dose
(mg/m3)
ALT (%)
ALP (%)
AST (%)
GGT (%)
0
-
-
-
-
2,090
2
12
22
50
6,270
-5
-4
10
0
20,900
4b
4
_Q
00
1
150
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Table 1-12. Evidence pertaining to liver biochemistry effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Suzuki et al. (2012); JPEC (2010a)
Rat, F344
oral—water
male (50/group): 0, 625, 2,500, or
10,000 ppm (0, 28, 121, 542 mg/kg-d);c
female (50/group): 0, 625, 2,500, or
10,000 ppm (0, 46, 171, 560 mg/kg-d);c
daily for 104 wk
Response relative to control:
Male
Dose
(mg/kg-d)
ALT (%)
ALP (%)
AST (%)
GGT (%)
0
-
-
-
-
28
-17
-5
-21
0
121
2
3
-3
43b
542
-4
0
-1
29
Female
Dose
(mg/kg-d)
ALT (%)
ALP (%)
AST (%)
GGT (%)
0
-
-
-
-
46
-10
-16
-19
0
171
-15
2
-1
0
560
-26
-15
-46b
33
NR = not reported.
a4.18 mg/m3 = 1 ppm.
bResult is statistically significant (p < 0.05) based on analysis of data by the study authors.
Conversion performed by the study authors.
-For controls, no response relevant; for other doses, no quantitative response reported.
Note: Percent change compared to controls calculated as 100 x [(treated value - control value) -f control value].
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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
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
B-
B-
-B-
B—B—U
~ ~ ~
~ ¦ ¦
B—B—¦
B-
B-
Centrilobular
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-
B-
B—~—~
~ ~ ~
B-
B-
B-
B-
-0
-a
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;104wks (E
Female rat; ALT, ALP, GGT;104wks (E
Female rat; i AST;104wks (E
B-
B-
-O
-B-
B-
-O
B-
B-
B-
B-
1 10 100 1,000 10,000
Dose (mg/kg-day)
Sources: (A) Fujii et al, 2010; JPEC,2008e [B] Gaoua, 2004b [C] 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.
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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
Relative Liver
Weight
Male rats; 13wks (A)
Female rats; 13wks (A)
Male rats; 13wks, 28d recovery (A)
Female rats; 13wks, 28d recovery (A)
B-
-B-
B-
-B-
-B-
~
Centrilobular
Hypertrophy
Male rats; 13wks (A
Female rats; 13wks (A
Male rats; 13wks, 28d recovery (A
Female rats; 13wks, 28d recovery (A
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
B-
B-
B-
B-
B-
B-
B-
B-
B-
B-
B-
B-
-B-
-B-
-B-
~
~
-B
-B
-B
-B
Male rats; ALT, AST, ALP, GGT; 13wks [A
Female rats; ALT, AST, ALP, GCT; 13wks (A
Male rats; ALT; 104wks [C
Male rats; IAST; 104wks [C
Male rats; IALP; 104wks (C
Male rats; TGGT; 104wks CC
Female rats; ALT, ALP, GGT; 104wks fC
Female rats; TAST; 104wks [C
Serum Liver
Enzymes
B-
B-
B-
B-
B-
B-
B-
B-
-B-
-B-
-B-
-B-
-B
-B
-a
100 1,000 10,000
Exposure Concentration (mg/m3)
100,000
Sources: [A) JPEC, 2008b [B] Medinsky et al„ 1999; Bond et al., 1996 (C) Saito et al„ 2013; JPEC, 2010b [D] Weng
et al., 2012
Figure 1-10. Exposure-response array of noncancer liver effects following
inhalation exposure to ETBE.
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Table 1-13. Evidence pertaining to liver tumor effects in animals exposed to
ETBE
Reference and study design
Results
Hepatocellular adenoma and carcinoma
Suzuki et al. (2012); JPEC (2010a)
Rat, F344
oral—drinking water
male (50/group): 0, 625, 2,500, or
10,000 ppm (0, 28,121, or 542 mg/kg-d);a
female (50/group): 0, 625, 2,500, or 10,000
ppm (0, 46,171, or 560 mg/kg-d)a
daily for 104 wk
Male
Dose
(mg/kg-d)
Adenoma
Carcinoma
Adenoma or
carcinoma
0
2/50
2/50
4/50
28
0/50
0/50
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
Maltoni et al. (1999)
Rat, S-D
oral—gavage in olive oil
male (60/group): 0, 250, or 1,000 mg/kg-d;
female (60/group): 0, 250, or 1,000 mg/kg-d
4 d/wk for 104 wk; observed until natural
death (depressed survival 25-28% seen in
controls at 104 wk)
Note: Tumor data not reanalyzed bv Malarkev
and Bucher (2011).
Male
Female
Dose
(mg/kg-d)
Adenoma or
carcinoma
Dose
(mg/kg-d)
Adenoma or
carcinoma
0
0/60
0
0/60
250
0/60
250
0/60
1,000
0/60
1,000
0/60
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Table 1-13. Evidence pertaining to liver tumor effects in animals exposed to
ETBE (continued)
Reference and study design
Results
Saito et al. (2013); JPEC (2010b)
Rat, F344
inhalation—vapor
male (50/group): 0, 500,1,500, or 5,000 ppm
(0, 2,090, 6,270, or 20,900 mg/m3);b female
(50/group): 0, 500,1,500, or 5,000 ppm (0,
2,090, 6,270, or 20,900 mg/m3)b
dynamic whole-body inhalation; 6 h/d,
5 d/wk for 104 wk; generation method,
analytical concentration, and method
reported
Male
Dose
(mg/m3)
Adenoma
Carcinoma
Adenoma or
carcinoma
0
0/50
0/50
0/50
2,090
2/50
0/50
2/50
6,270
1/50
0/50
1/50
20,900
9/50°
1/50
10/50°
Female
Dose
(mg/m3)
Adenoma
Carcinoma
Adenoma or
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
Initiation-promotion studies
Hagiwara et al. (2011); JPEC (2008d)
Rat, F344
oral—gavage in olive oil
male (30/group): 0, 300, or 1,000 mg/kg-d
daily for 23 wk following a 4-wk tumor
initiation by DMBDDd
Male
Dose (mg/kg-d)
Adenoma
Carcinoma
0
0/30
1/30
300
1/30
0/30
1,000
6/30°
0/30
0e
0/12
0/12
l,000e
0/12
0/12
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Table 1-13. Evidence pertaining to liver tumor effects in animals exposed to
ETBE (continued)
Reference and study design
Results
Hagiwara et al. (2015)
Rat, Wistar
oral—gavage in olive oil
male (30/group): 0,100, 300, 500, or
1,000 mg/kg-d
daily for 19 wk following 2-wk tumor
initiation by
W-ethyl-W-hydroxyethylnitrosamine (EHEN)
Male
Dose
(mg/kg-d)
Adenoma
Carcinoma
Adenoma or
carcinoma
0
4/30
0/30
4/30
100
5/30
2/30
7/30
300
8/30
0/30
8/30
500
8/30
3/30
10/30
1,000
15/30°
5/30°
17/30°
Conversion performed by the study authors.
b4.18 mg/m3 = 1 ppm.
cResult is statistically significant (p < 0.05) based on analysis of data by study authors.
dDMBDD = diethylnitrosamine (DEN), W-butyl-W-(4-hydroxybutyl)nitrosamine (BBN), W-methyl-W-nitrosourea
(MNU), 1,2-dimethylhydrazine dihydrochloride (DMH), and W-bis(2-hydroxypropyl)nitrosamine (DHPN).
eNo DMBDD initiation.
-For controls, no response relevant; for other doses, no quantitative response reported.
Mode-of-Action Analysis—Liver Effects
Key characteristics of carcinogens
Mechanistic information was grouped into 10 "key characteristics" useful for summarizing
and organizing the mechanistic data relevant to carcinogens (Smith etal.. 20161. The evidence
available for each characteristic is summarized in Table 1-14. Altogether, experimental evidence
informing several of the key characteristics of carcinogens was identified in the available literature.
ETBE was found to have the potential to form electrophilic metabolites, but it was concluded that
there was inadequate evidence that ETBE induces any of the remaining key characteristics.
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Table 1-14. Evidence of key characteristics of carcinogens for ETBE
Characteristic
Evidence
1. Is electrophilic or can be metabolically activated to
electrophiles
Metabolized to acetaldehyde in the livera b
2. Is genotoxic
Inadequate evidence to draw a conclusion from 12
studies examining micronucleus, DNA strand breaks,
chromosomal aberration, and gene mutation assaysbc
3. Alters DNA repair or causes genomic instability
No pertinent studies identified
4. Induces epigenetic alterations
No pertinent studies identified
5. Induces oxidative stress
Inadequate evidence to draw a conclusion from three
studies examining 8-OHdG, 8-hOGGl formationbd
6. Induces chronic inflammation
No pertinent studies identified
7. Is immunosuppressive
No pertinent studies identified
8. Modulates receptor-mediated effects
Inadequate evidence to draw a conclusion from two
studies examining PPAR, CAR, and PXR activation6
9. Causes immortalization
No pertinent studies identified
10. Alters cell proliferation, cell death, or nutrient
supply
Inadequate evidence to draw a conclusion from three
studies examining basophilic, acidophilic foci, and
cellular proliferation6
CAR = constitutive androstane receptor; PPAR = peroxisome proliferator-activated receptor; PXR = pregnane X
receptor.
aSee Supplemental Information Section B.1.3.
bSee Acetaldehyde-mediated liver toxicity and genotoxicity in this section.
cSee Supplemental Information Section B.2.2.
dSee Oxidative stress in this section.
eSee Receptor-mediated effects in this section.
Toxicokinetic considerations relevant to liver toxicity and tumors
ETBE is metabolized by CYP450 enzymes to an unstable hemiacetal that decomposes
spontaneously into tert-butanol and acetaldehyde fBernauer et al.. 19981. 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 more detail in the IRIS
assessment of tert-butanol) and acetaldehyde are relevant to evaluating the MOA for 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
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establish a conclusion. One study reported that tert-butanol might induce centrilobular
hypertrophy in mice after 2 weeks fBlanck et al.. 20101. but no related liver pathology was
observed in other repeat-exposure rodent studies, including both subchronic and 2-year bioassays.
Although Blanck etal. f 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 of ETBE appears unlikely.
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 fSetshedi etal.. 20101. Additional discussion on the potential role of acetaldehyde in the
liver carcinogenesis of ETBE 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 increased centrilobular hypertrophy, which is one possible indication of liver
enzyme induction, chronic-exposure-induced focal proliferative lesions (including basophilic and
acidophilic foci) could be more directly related to tumorigenesis. The centrilobular hypertrophy
was increased in rats of both sexes via both oral and inhalation exposure at subchronic time points;
but not at 2 years, although significantly increased liver weight was observed. However, liver
tumors were only observed in one sex (males) following one route of exposure (inhalation),
indicating that subchronic hypertrophy may not be 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, some of which
occur by activating 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 (Klaunigetal.. 2003). The sequence of key events hypothesized for
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.. 2014). 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 of ETBE reported that several key sequences in the PPARa, PXR, and CAR pathways
were affected fKakehashi et al.. 2 0131.
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PPAR
Limited evidence suggests that ETBE could activate PPAR-mediated events fKakehashi et
al.. 20131. For instance, messenger ribonucleic acid (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, such as AC0X1, CYP4A2, and ECH1, involved in lipid and xenobiotic
metabolism were upregulated in the liver after 2 weeks of exposure. 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 of ETBE. Cell proliferation was increased after 3 days
fKakehashi et al.. 20151. unchanged after 1 week, significantly decreased after 2 weeks fKakehashi
etal.. 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 (Saito etal.. 2013: TPEC. 2010bl.
PPARa-mediated genes were investigated in one study fKakehashi etal.. 20131. The high
dose of ETBE (2,000 mg/kg-day), which induced the most consistent changes in PPARa, Cyp4a,
Cypla, and Cyp3a in the 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 f Saito etal.. 2013:
TPEC. 2010b). Only Cyp2b genes associated with PPARa expression were affected at the low gavage
dose (300 mg/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 of ETBE exposure fKakehashi etal.. 20131. However, 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 whether PPARa gene expression changes in knockout (KO) mice.
Overall, these data are inadequate to conclude that ETBE induces liver tumors via a PPARa MOA.
CAR/PXR
Kakehashi etal. f20131 reported several CAR- and PXR-mediated events following ETBE
exposure. After 2 weeks of exposure at the high dose of ETBE, 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 in the 2-year
inhalation study fSaito etal.. 2013: TPEC. 2010bl. Histological evidence (preneoplastic foci)
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supporting increased liver cell proliferation is available following chronic, but not subchronic
exposures fSaito etal.. 2013: TPEC. 2010bl. Several data gaps, such as a lack of clonal expansion
and gap junction communication, were not evaluated. These data provide evidence that CAR and
PXR are activated at high concentrations in the liver following acute ETBE exposure. Furthermore,
the data do not provide enough information to determine dose-response relationships or temporal
associations, which are helpful for establishing an MOA. Altogether, these data are inadequate to
conclude that ETBE induces liver tumors via a CAR/PXR MOA.
In summary, several gaps in the receptor-mediated effects data (for PPARa, CAR, and PXR)
are noted such as evidence in only one species, lack of any studies in PPAR KO mice, lack of dose
response concordance between receptor-mediated gene changes and tumors, and lack of any
receptor-mediated data outside of the 1 and 2 week time points, which preclude establishing
temporal associations. Overall, these data are inadequate to conclude that ETBE induces liver
tumors via a PPARa or CAR/PXR MOA.
Acetaldehvde-mediated liver toxicity and genotoxicity
Another possible MOA for increased tumors could be 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 (IARC. 2012). Acetaldehyde administered directly has been
shown 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 et al.. 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 (IARC. 2012: Brennan et al.. 2004). IARC (2012) found that ALDH2 status was
associated with increased esophageal cancer. Although IARC f20121 found inconclusive evidence
for a contribution of ALDH2 to liver cancer, Eriksson (2015) 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 Aldh2 KO mice compared with WT, while females appeared to be less sensitive, similar 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
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males, but not in WT females. Measures of DNA damage were all statistically significantly increased
in both male and female Aldh2 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 Aldh2 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 on oxidative stress are limited to several that measured
oxidative DNA damage in leukocytes and hepatocytes in mice fWeng etal.. 20121 and one study in
the liver of rats (Kakehashi etal.. 2013). Male mice had increased levels of 8-OHdG in hepatocytes
after 13 weeks of inhalation exposure to the concentration of ETBE 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 gavage in rats f Kakehashi etal.. 20131 at a concentration
twofold 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 Aldh2 KO mice fWeng etal.. 20131 and Aldh2
heterozygous mice (Wengetal.. 2019). Overall, these data are inadequate to conclude that ETBE
induces liver tumors via oxidative stress.
Overall Conclusions on Mode of Action (MOA) 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 et al.. 2014: Klaunig et
al.. 20031. although this conclusion has been questioned fGuvton et al.. 20091. The database is
inadequate to determine whether 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 effects are not
associated with the observed rat liver tumorigenesis. Therefore, given the available data,
ETBE-induced liver tumors in male rats are relevant to human hazard identification and are
scientifically supported.
Evidence suggests that metabolism of ETBE to acetaldehyde could contribute to
ETBE-induced liver carcinogenesis. For instance, enhancement of ETBE-induced liver toxicity and
genotoxicity has been reported inv4/d/j2-deficient mice, which have an impaired ability to
metabolize acetaldehyde fWengetal.. 2013: Wengetal.. 20121. The database, however, is
inadequate to conclude that ETBE induces liver tumors via acetaldehyde-mediated mutagenic MOA.
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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, which show centrilobular hypertrophy, increased liver weight, and changes in
serum liver enzyme levels. Based on dose-related increases in relative liver weight, and increases
in hepatocellular hypertrophy in male and female rats, contrasted with the poor temporal
correlation of serum biomarkers and pathological lesions indicative of accumulating liver 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 can progress to malignancy, eventually forming carcinomas (Liau 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 AIdh2 enhances ETBE-induced liver
toxicity and genotoxicity, which is consistent with the observed genotoxicity being mediated by the
ETBE metabolite acetaldehyde, although the database is inadequate to conclude that ETBE induces
liver tumors via this 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 from ETBE exposure consists of data
from rats and mice but not humans. 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-gene ration oral study fGaoua.
2004b), 13- and 9-week inhalation studies fWengetal.. 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. (2012): TPEC (2010a):
inhalation: Saito etal. (2013): TPEC (2010b)]. a medium-term carcinogenicity study (2 3-week oral
exposure) (Hagiwara etal.. 2011: TPEC. 2008d). a 180-day oral study (Mivataetal.. 2013: TPEC.
2008c), a 90-day inhalation study flPEC. 2008bl. and a 13-week inhalation study fMedinskv etal..
19991. These studies were conducted in S-D rats, F344 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 1-15.
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The one- and two-generation reproductive toxicity studies found no effects on copulation,
fertility, or sperm parameters in adult male S-D rats exposed to ETBE by gavage at concentrations
up to 1,000 mg/kg-day for 10 weeks prior to mating fFuiii etal.. 2010: Gaoua. 2004bl. nor in F1
male offspring exposed during gestation, lactation, and postweaning in the diet fGaoua. 2004bl 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 F344 rats at oral
doses up to 542 mg/kg-day (Suzuki etal.. 2012: TPEC. 2010a) or at inhalation exposure
concentrations up to 20,900 mg/m3 fSaito etal.. 2013: TPEC. 2010bl: in the medium-term
carcinogenicity study in F344 rats fHagiwara etal.. 2011: TPEC. 2008dl: in the 180-day oral study in
S-D rats at doses up to 400 mg/kg-day fMivataetal.. 2013: TPEC. 2008cl: in the 90-day inhalation
study in S-D rats at doses up to 20,900 mg/m3 flPEC. 2008bl: or in the 14-day oral study in F344
rats at doses up to 1,800 mg/kg-day (de Pevster et al.. 2009). 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 (Gaoua. 2004b 1 at the
highest doses tested, which may have been attributable to reduced body-weight gain in these
groups.
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
F344 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 fMedinskv et al.. 19991. In C57BL/6 wild-type and Aldh2 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 than in wild type (Weng etal.. 2014). 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
Aldh2 KO mice (2,090 mg/m3) compared with wild type (7,320-20,090 mg/m3). Significantly
decreased epididymis weight was observed in Aldh2 KO mice but not wild-type mice.
Weng etal. (2014) also conducted a 9-week inhalation study using lower ETBE exposure
concentrations (209-2,090 mg/m3) and three mouse genotypes (wild type, Aldh2-/~, and AIdh2
heterozygous; n = 5/group). 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 \x\Aldh2-/~ and heterozygous
mice at exposure concentrations as low as 836 mg/m3 ETBE. Aldh2 heterozygous mice had
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significantly decreased relative testis and epididymis weight in the 20,090 mg/m3 exposure group.
However, for unknown reasons, several reproductive effects were noted to be more pronounced in
the heterozygous mice than in the Aldh2-/~ mice. Taken together, the results ofWengetal. T20141
indicated that populations with inactive Aldh2 variants may be more susceptible to male
reproductive toxicity following exposure to ETBE. However, these effects are considered to be
preliminary because of the small sample size [n = 5) in one species and in one study, and the
unconvincing magnitude of many of the statistically significant effects (including the observation
that the heterozygotes exhibited more robust changes than the knockouts).
Although testicular lesions were not found in the 14-day oral study in F344 rats fde Pevster
etal.. 20091. plasma estradiol levels in these animals were increased by up to 106% compared with
controls. Plasma testosterone in the 1,800 mg/kg-day dose group was decreased by 34% compared
with 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 S-D rat Leydig cells and found reduced testosterone production in
ETBE-treated cells compared with controls (data not shown in evidence table).
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Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE
Reference and study design
Results
Male fertility
Fuiiietal. (2010); JPEC (2008e)
Rat, S-D
oral—gavage
F0, male and female
(24/sex/group): 0,100, 300, or
1,000 mg/kg-d; dosed daily for
17 wk, from 10 wk premating
to LD 21
F0 generation—parent
Dose
(mg/kg-d)
Copulation
index(%)
Absolute
change from
control (%)
Fertility index
(%)
Absolute
change from
control (%)
0
100
-
87.5
-
100
91.7
-8.3
100
12.5
300
95.8
-4.2
95.7
8.2
1,000
100
0
91.7
4.2
Gaoua (2004b)
Rat, S-D
oral—gavage
F0, male and female
(25/sex/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily for 18 wk from
10 wk premating until weaning
of F1 pups
Fl, male and female
(24-25/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
F0 generation—parent
Dose
(mg/kg-d)
Male mating
index(%)
Absolute
change from
control (%)
Male fertility
index (%)
Absolute
change from
control (%)
0
100
-
92
-
250
100
0
84
-8
500
100
0
88
-4
1,000
100
0
100
8
Fl generation—offspring
Dose
(mg/kg-d)
Male mating
index3 (%)
Absolute
change from
control (%)
Male fertility
indexb (%)
Absolute
change from
control (%)
0
96
-
92
-
250
96
0
92
0
500
100
4
88
-4
1,000
96
0
96
4
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Testicular histopathology
Medinskv et al. (1999); U.S.
EPA (1997)
Rat, F344
inhalation—vapor
male (48/group): 0, 500,1,750,
or 5,000 ppm
(0, 2,090, 7,320, or
20,900 mg/m3);a f
dynamic whole-body chamber;
6 h/d, 5 d/wk for 13 wk
Dose
(mg/m3)
Incidence of spermatocyte
degeneration
Incidence of sloughed
epithelium
0
11/11
7/11
2,090
11/11
3/11
7,320
11/11
3/11
20,900
10/11
7/11
Dose
(mg/m3)
Mean seminiferous tubules with
spermatocyte degeneration (%)
Absolute change from
control (%)
0
2.1
-
2,090
2.4
0
7,320
7.8°
6
20,900
12.7°
11
Dose
(mg/m3)
Mean seminiferous tubules with
luminal debris (%)
Absolute change from
control (%)
0
2.1
-
2,090
0.7
-1
7,320
2.8
1
20,900
1
-1
Weng et al. (2014)
Mice, C57BL/6
inhalation—vapor
male (5/group): 0, 500,1,750,
or 5,000 ppm (0, 2,090, 7,320,
or 20,900 mg/m3)a
dynamic whole-body
inhalation; 6 h/d, 5 d/wk for
13 wk; methods described in
Weng et al. (2012)
Wild-type mice; 13-wk exposure
Dose
(mg/m3)
Incidence of
"extremely slight"
atrophy
Incidence of
"slight" atrophy
Total incidence of
atrophy of
seminiferous tubules
0
1/5
0/5
1/5
2,090
0/5
0/5
0/5
7,320
2/5
0/5
2/5
20,900
3/5
0/5
3/5
Knockout mice (Aldh2-/-); 13-wk exposure
Dose
(mg/m3)
Incidence of
"extremely slight"
atrophy
Incidence of
"slight" atrophy
Total incidence of
atrophy of
seminiferous tubules
0
2/5
0/5
2/5
2,090
2/5
3/5
5/5
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
7,320
4/5
1/5
5/5
20,900
3/5
2/5
5/5
Sperm parameters
Gaoua (2004b)
Rat, S-D
oral—gavage
F0, male and female
(25/sex/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily for 18 wk from
10 wk premating until weaning
of F1 pups
Fl, male and female
(24-25/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
F0 males
Dose
(mg/kg-d)
Mean epididymal
spermatozoa
count (n) ± SD
% Change
from
control
Mean epididymal
sperm motility
(%) ± SD
Absolute
change from
control (%)
0
923 ± 200
-
99.7 ± 1.5
-
250
938 ± 205
2
100 ±0
0
500
935 ±159
1
98.6 ±4
-1
1,000
918 ±194
-1
97.6 ±6.6
-2
Dose
(mg/kg-d)
Mean epididymal
sperm with normal
morphology
(%) ± SD
Absolute
change
from
control (%)
Mean testicular
sperm heads
(106/gram of
testis) ± SD
% Change
from 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
Dose
(mg/kg-d)
Mean daily
testicular sperm
production
(106/gram of
testis)
% Change
from
control
n (epididymal
sperm count)
n (other
sperm
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
Fl Males
Dose
(mg/kg-d)
Mean epididymal
spermatozoa
count (n)
± SD
% Change
from
control
Mean epididymal
sperm motility
(%) ± SD
Absolute
change from
control (%)
0
725 ±150
-
84.6 ±34.1
-
250
673 ±197
-7
87.1 ±31.6
3
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Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
500
701 ± 97
-3
93.3 ±22
9
1,000
688 ±177
-5
88.3 ± 29.4
4
Gaoua (2004b) (continued)
Dose
(mg/kg-d)
Mean epididymal
sperm with normal
morphology
(%) ± SD
Absolute
change
from
control (%)
Mean testicular
sperm heads
(106/gram of
testis) ± SD
% Change
from 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
Dose
(mg/kg-d)
Mean daily
testicular sperm
production
(106/gram of
testis)
% change
from
control
n (epididymal
sperm count)
n (other
sperm
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
Weng et al. (2014)
Mice, C57BL/6
inhalation—vapor
male (5/group): 0, 500,1,750,
or 5,000 ppm (0, 2,090, 7,320,
or 20,900 mg/m3)a
dynamic whole-body
inhalation; 6 h/d, 5 d/wk for
13 wk; methods described in
Weng et al. (2012)
Wild-type mice; 13-wk exposure
Dose
(mg/m3)
Mean sperm head
numbers (testis)
(x 106/g) ± SD
% Change
from
control
Motile sperm
(epididymal) ± SE
Absolute
change from
control (%)
0
166.62 ±21.9
-
67.34 ±3.45
-
2,090
167.74 ± 28.02
1
69.64 ±3.45
2
7,320
167.78 ±25.52
1
62.73 ± 1.73
-5
20,900
150.94 ± 23.07
-9
58.13 ±2.30
-9
Dose
(mg/m3)
% Static sperm
(epididymal)
Absolute
change
from
control (%)
% Sperm with
rapid movement
(epididymal)
Absolute
change from
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
10
46.25° ±2.50
-9
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Weng et al. (2014) (continued)
Dose
(mg/m3)
Epididymal sperm
DNA breaks (tail
intensity in comet
assay)
% Change
from
control
Epididymal sperm
DNA damage
(measurement of
8-OHdG in comet
assay)
% Change
from 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-wk exposure
Dose
(mg/m3)
Mean sperm head
numbers (testis)
(x 106/g) ± SD
% Change
from
control
Motile sperm
(epididymal) ± SE
Absolute
change from
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
Dose
(mg/m3)
% Static sperm
(epididymal)
Absolute
change
from
control (%)
% Sperm with
rapid movement
(epididymal)
Absolute
change from
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
20
45.03° ± 3.97
-22
Dose
(mg/m3)
Epididymal sperm
DNA breaks (tail
intensity in comet
assay)
% Change
from
control
Epididymal sperm
DNA damage
(measurement of
8-OHdG in comet
assay)
% Change
from 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
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Weng et al. (2014)
Mice, C57BL/6
inhalation—vapor
male (NR): 0, 50, 200, or
500 ppm (0, 209, 836, or
2,090 mg/m3)a
dynamic whole-body
inhalation; 6 h/d, 5 d/wk for
9 wk; methods described in
Weng et al. (2012)
Wild-type mice; 9-wk exposure
Dose
(mg/m3)
Mean sperm head
numbers (testis)
(x 106/g) ± SD
% Change
from
control
Motile sperm
(epididymal) ± SE
Absolute
change from
control (%)
0
199.62 ±27.22
-
85.82 ±4.26
-
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
80.14 ± 1.42
-6
Dose
(mg/m3)
% Static sperm
(epididymal)
Absolute
change
from
control (%)
% Sperm with
rapid movement
(epididymal)
Absolute
change from
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
Dose
(mg/m3)
Epididymal sperm
DNA breaks (tail
intensity in comet
assay)
% Change
from
control
Epididymal sperm
DNA damage
(measurement of
8-OHdG in comet
assay)
% Change
from 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 (Aldh2-/-); 9-wk exposure
Dose
(mg/m3)
Mean sperm head
numbers (testis)
(x 106/g) ± SD
% Change
from
control
Motile sperm
(epididymal) ± SE
Absolute
change from
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
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Weng et al. (2014) (continued)
Dose
(mg/m3)
% Static sperm
(epididymal)
Absolute
change
from
control (%)
% Sperm with
rapid movement
(epididymal)
Absolute
change from
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
Dose
(mg/m3)
Epididymal sperm
DNA breaks (tail
intensity in comet
assay)
% Change
from
control
Epididymal sperm
DNA damage
(measurement of
8-OHdG in comet
assay)
% Change
from 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
Haplotype mice (Aldh2 heterozygous); 9-wk exposure
Dose
(mg/m3)
Mean sperm head
numbers (testis)
(x 106/g) ± SD
% Change
from
control
Motile sperm
(epididymal) ± SE
Absolute
change from
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
Dose
(mg/m3)
% Static sperm
(epididymal)
Absolute
change
from
control (%)
% Sperm with
rapid movement
(epididymal)
Absolute
change from
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
1-72
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Weng et al. (2014) (continued)
Dose
(mg/m3)
Epididymal sperm
DNA breaks (tail
intensity in comet
assay)
% Change
from
control
Epididymal sperm
DNA damage
(measurement of
8-OHdG in comet
assay)
% Change
from 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)
Rat, S-D
oral—gavage
F0, male and female
(24/sex/group): 0,100, 300, or
1,000 mg/kg-d; dosed daily for
17 wk, from 10 wk premating
to LD 21
F0 parents—absolute organ weights
Dose
(mg/kg-d)
Mean testis weight
(g) ± SD
% Change
from
control
Mean epididymis
weight (mg) ± SD
% Change
from control
0
3.47 ±0.31
-
1,371 ± 136
-
100
3.48 ±0.28
0
1,360 ± 83
-1
300
3.57 ±0.24
3
1,381 ± 73
1
1,000
3.57 ±0.31
3
1,349 ± 95
-2
Dose
(mg/kg-d)
Mean prostate
weight (mg)
± SD
% Change
from
control
Mean seminal
vesicle weight (g)
± SD
% Change
from 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
Dose
(mg/kg-d)
Mean testis: body
weight ratio (%)
± SD
Absolute
change
from
control (%)
Mean epididymis:
body weight ratio
(%) ± SD
Absolute
change from
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
11
1-73
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Fuiiietal. (2010); JPEC (2008e)
(continued)
Dose
(mg/kg-d)
Mean prostate:
body weight ratio
(%) ± SD
Absolute
change
from
control (%)
Mean seminal
vesicle:body
weight ratio (%)
± SD
Absolute
change from
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)
Rat, S-D
oral—gavage
F0, male and female
(25/sex/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily for 18 wk from
10 wk premating until weaning
of F1 pups
Fl, male and female
(24-25/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
F0 parents—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
1.78 ±0.116
-
1.76 ±0.105
-
250
1.73 ±0.181
-3
1.76 ±0.179
0
500
1.78 ±0.142
0
1.76 ±0.13
0
1,000
1.75 ±0.237
-2
1.79 ±0.126
2
Dose
(mg/kg-d)
Mean epididymis
weight (left) (g)
± SD
% Change
from
control
Mean epididymis
weight (right) (g)
± SD
% Change
from 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
Dose
(mg/kg-d)
Mean prostate
weight ± SD
% Change
from
control
Mean seminal
vesicle weight
± SD
%
Change
from
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
1-74
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Gaoua (2004b) (continued)
F0 parents—relative organ weights
Dose
(mg/kg-d)
Mean testis
weight:body
weight ratio (left)
(g) ± SD
Absolute
change
from
control (%)
Mean testis
weight:body
weight ratio
(right) (g) ± SD
Absolute
change from
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
Dose
(mg/kg-d)
Mean epididymis
weight (left):body
weight ratio (g)
± SD
Absolute
change
from
control (%)
Mean epididymis:
body weight ratio
(right) (%) ± SD
Absolute
change from
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
Dose
(mg/kg-d)
Mean prostate
weight:body
weight ratio ± SD
Absolute
change
from
control (%)
Mean seminal
vesicle:body
weight ratio ± SD
Absolute
change
from
control
(%)
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
1-75
-------
Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Gaoua (2004b) (continued)
Dose
(mg/kg-d)
Mean epididymis
weight (left) (g)
±SD
% Change
from
control
Mean epididymis
weight (right) (g)
± SD
% Change
from 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
F1 offspring—relative organ weights
Dose
(mg/kg-d)
Mean testis
weight:body
weight ratio (left)
(g) ± SD
Absolute
change
from
control (%)
Mean testis
weight:body
weight ratio
(right) (g) ± SD
Absolute
change from
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
0.00
0.30958 ±0.05
0.00
1-76
-------
Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Gaoua (2004b) (continued)
Dose
(mg/kg-d)
Mean epididymis
weight (left): body
weight ratio (g)
±SD
Absolute
change
from
control (%)
Mean epididymis
weight (right) (g)
± SD
Absolute
change from
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
Dose
(mg/kg-d)
Mean prostate
weight:body
weight ratio ± SD
Absolute
change
from
control (%)
Mean seminal
vesicle:body
weight ratio ± SD
Absolute
change
from
control
(%)
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)
Mice, C57BL/6
inhalation—vapor
male (5/group): 0, 500,1,750,
or 5,000 ppm (0, 2,090, 7,320,
or 20,900 mg/m3)a
dynamic whole-body
inhalation; 6 h/d, 5 d/wk for
13 wk; methods described in
Weng et al. (2012)
Wild type mice; 13-wk exposure
Dose
(mg/m3)
Mean testis:body
weight ratio (%)
± SD
Absolute
change
from
control (%)
Mean epididymis:
body weight ratio
(%) ± SD
Absolute
change from
control (%)
0
0.7 ±0.06
-
0.24 ±0.02
-
2,090
0.74 ± 0.04
0.04
0.26 ±0.02
0.02
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
Knockout mice (Aldh2-/-); 13-wk exposure
Dose
(mg/m3)
Mean testis:body
weight ratio (%)
± SD
Absolute
change
from
control (%)
Mean epididymis:
body weight ratio
(%) ± SD
Absolute
change from
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
1-77
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Weng et al. (2014)
Mice, C57BL/6
inhalation—vapor
male (NR): 0, 50, 200, or
500 ppm (209, 836, or
2,090 mg/m3)a
dynamic whole-body
inhalation; 6 h/d, 5 d/wk for 9
wk; methods described in
Weng et al. (2012)
Wild type mice; 9-wk exposure
Dose
(mg/m3)
Mean testis:body
weight ratio (%)
±SD
Absolute
change
from
control (%)
Mean epididymis:
body weight ratio
(%) ± SD
Absolute
change from
control (%)
0
0.8 ±0.12
-
0.26 ±0.03
-
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-wk exposure
Dose
(mg/m3)
Mean testis:body
weight ratio (%)
±SD
Absolute
change
from
control (%)
Mean epididymis:
body weight ratio
(%) ± SD
Absolute
change from
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 heterozygous); 9-wk exposure
Dose
(mg/m3)
Mean testis:body
weight ratio (%)
±SD
Absolute
change
from
control (%)
Mean epididymis:
body weight ratio
(%) ± SD
Absolute
change from
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
de Pevster et al. (2009)
Rat, F344
oral—gavage
P0, male (12/group): 0, 600,
1,200, or 1,800 mg/kg-d
daily for 14 d
Absolute organ weights
Dose
(mg/kg-d)
Mean testis weight
(g) ± SD
% Change
from
control
Mean epididymis
weight (mg) ± SD
% Change
from control
0
2.55 ±0.09
-
0.696 ±0.016
-
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
-3
0.663 ±0.029
-5
1-78
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
de Pevster et al. (2009)
(continued)
Dose
(mg/kg-d)
Mean prostate
weight (g)
±SD
% Change
from
control
Mean seminal
vesicle weight (g)
± SD
% Change
from 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
Dose (mg/kg-d)
Mean weight of combined
accessory sex 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
Dose
(mg/kg-d)
Mean testis: body
weight ratio (%)
±SD
Absolute
change
from
control (%)
Mean epididymis:
body weight ratio
(%) ± SD
Absolute
change from
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
Dose
(mg/kg-d)
Mean prostate:
body weight ratio
(%) ± SD
Absolute
change
from
control (%)
Mean seminal
vesicle:body wt.
ratio (%) ± SD
Absolute
change from
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
Dose (mg/kg-d)
Mean combined accessory sex
organs:body weight ratio (%)
± SD
Absolute change from
control (%)
0
0.668 ±0.018
-
600
0.691 ±0.026
0.02
1-79
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-15. Evidence pertaining to male reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
1,200
0.746 ±0.019
0.08
1,800
0.727 ±0.035
0.06
Medinskv et al. (1999); U.S.
EPA (1997)
Rat, F344
inhalation—vapor
male (48/group): 0, 500,1,750,
or 5,000 ppm
(0, 2,090, 7,320, or
20,900 mg/m3);a female
(48/group): 0, 500,1,750, or
5,000 ppm
(0, 2,090, 7,320, or
20,900 mg/m3)a
dynamic whole-body chamber;
6 h/d, 5 d/wk for 13 wk
Organ weights of F344 rats and CD-I mice were not altered by exposure to
ETBE.
Testosterone and estradiol
de Pevster et al. (2009)
Rat, F344
oral—gavage
P0, male (12/group): 0, 600,
1,200, or 1,800 mg/kg-d
daily for 14 d
Dose
(mg/kg-d)
n
Mean plasma testosterone
(ng/mL) ± SE
% Change from control
0
12
2.07 ±42
-
600
12
3.1 ±0.78
50
1,200
11
2.61 ±0.55
26
1,800
10
1.36 ±0.39
-34
Dose
(mg/kg-d)
n
Mean plasma estradiol
(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
LD = lactation day; NR = not reported; SD = standard deviation; SE = standard error.
a4.18 mg/m3 = 1 ppm.
Conversion performed by the study authors.
cResult is statistically significant (p < 0.05) based on analysis of data by the study authors.
Male mating index (%) = (no. males able to mate with at least one female -f total males) x 100.
Male fertility index (%) = (no. males with at least one pregnant partner 4 males that mated at least once) x 100
-For controls, no response relevant; for other doses, no quantitative response reported.
% change from control = [(treated group value - control value) 4 control value] x 100.
Absolute change from control (%) = control value (%) - treated group value (%).
1-80
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Toxicological Review of Ethyl Tertiary Butyl Ether
¦ = 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 Fertility
FO Male rat; copulation; one-gen repro (A) Q-
FO Male rat; copulation; two-gen repro (B)
F1 Male rat; copulation; two-gen repro (B)
F1 Male rat; epididymal sperm morphology; two-gen repro
(B)
F1 Male rat; testicular sperm heads; two-gen repro (B)
FO Male rat; daily sperm production; two-gen repro (!:')
FO Male rat; fertility; one-gen repro (A) | 13 B
~ ~ ~
~ ~ ~
FO Male rat; fertility; two-gen repro (B) -j ~—B—~
F1 Male rat; fertility; two-gen repro (B) ] ~—0—E)
Sperm FO Male rat; epididymal sperm count; one-gen repro (A) B B B
Parameters
FO Male rat; epididymal sperm count; two-gen repro (B) B—B—0
F1 Male rat; epididymal sperm count; two-gen repro (B) B—B—B
FO Male rat; epididymal sperm motility; one-gen repro (A) B B
FO Male rat; epididymal sperm motility; two-gen repro (B) j | B—B—B
F1 Male rat; epididymal sperm motility; two-gen repro (B) ~—B—B
FO Male rat; epididymal sperm morphology; one-gen repro _
(A) U
FO Male rat; epididymal sperm morphology; two-gen repro
(B)
~—B—B
~—B—B
FO Male rat; testicular sperm heads; one-gen repro (A) j [p-
FO Male rat; testicular sperm heads; two-gen repro (B) -| ~—B—B
~ BO
B—B—B
Fl Male rat; daily sperm production; two-gen repro (B) \ | B—B—B
_ _ _
Male rat; testosterone; 14 d (C) B—~ ~
Testosterone
and Estradiol
Male rat; estradiol; 14 d (C) B-
10 100 1,000 10,000
Dose (mg/kg-day)
Sources: (A] Fujii et al., 2010; JPEC, 2008e (B) Gaoua, 2004b (C) de Peyster et al., 2009
Figure 1-11. Exposure-response array of male reproductive 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
Sperm
Parameters
Testicular
Histopathology Male rat; spermatocyte degeneration; 13 wks
Male mouse; atrophy of seminiferous tubules; 13 wks
Male Aldh2-/- mouse; atrophy ofseminiferous tubules; 13 wks
Male mouse; atrophy of seminiferous tubules; 9 wks
Male Aldh2-/_ mouse; atrophy ofseminiferous tubules; 9 wks
Male Aldh2+/- mouse; atrophy ofseminiferous tubules; 9 wks
Male mouse; deer, sperm heads; 13 wks
Male Aldh2-/- mouse; deer, sperm heads; 13 wks
Male mouse; deer, sperm heads; 9 wks
Male Aldh2-/- mouse; deer, sperm heads; 9 wks
Male Aldh2+/- mouse; deer, sperm heads; 9 wks
Male mouse; deer, sperm motility; 13 wks
Male Aldh2-/- mouse; deer, sperm motility; 13 wks
Male mouse; deer, sperm motility; 9 wks
Male Aldh2-/- mouse; deer, sperm motility; 9 wks
Male Aldh2 +/- mouse; deer, sperm motility; 9 wks
Male mouse; incr. no. of static sperm; 13 wks
Male Aldh2-/- mouse; incr. no. of static sperm; 13 wks
Male mouse; incr. no. of static sperm; 9 wks
Male Aldh2-/- mouse; incr. no. of static sperm; 9 wks
Male Aldh2+/- mouse; incr. no. of static sperm; 9 wks
Male mouse; deer. no. of rapidly moving sperm; 13 wks
Male Aldh2-/- mouse; deer. no. of rapidly moving sperm; 13 wks
Male mouse; deer. no. of rapidly moving sperm; 9 wks
Male Aldh2-/- mouse; deer. no. of rapidly moving sperm; 9 wks
Male Aldh2+/- mouse; deer, no. of rapidly moving sperm; 9 wks
Organ
Weights
Male mouse; deer, testis wt. (relative); 13 wks
Male Aldh2-/- mouse; deer, testis wt. (relative); 13 wks
Male mouse; deer, testis wt. (relative); 9 wks
Male Aldh2-/- rnouse; deer, testis wt. (relative); 9 wks
Male Aldh2+/- mouse; deer, testis wt. (relative); 9 wks
Male mouse; deer, epididymis wt (relative); 13 wks
Male Aldh2 -/- mouse; deer, epididymis wt. (relative); 13 wks
Male mouse; deer, epididymis wt (relative); 9 wks
Male Aldh2-/- mouse; deer, epididymis wt. (relative); 9 wks
Male Aldh2+/- mouse; deer, epididymis wt. (relative); 9 wks
b-
13-
B-
B-
B-
B-
B-
B-
B-
B-
Q-
B-
B-
B-
&-
B-
B-
B-
B-
B-
B-
B-
Q-
B-
B-
¦
-B
B-
¦
-~
B-
B-
-~
B-
B-
-B
-B- 0
-BH B
-BH B
-a-
100 1,000 10,000 100,000
Exposure Concentration (mg/m ')
Sources: (A] Medinsky et al, 1999; Bond et al., 1996b (B] Weng et al., 2014
Figure 1-12. Exposure-response array of male reproductive effects following
inhalation exposure to ETBE.
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Mechanistic Evidence of Male Reproductive Effects
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 in wild-type animals. The 13-week and 9-week inhalation studies
conducted in rats and mice (Weng etal.. 2014: Medinskv et al.. 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 subchronic studies. For example, the 2-year inhalation carcinogenicity
study (Saito etal.. 2013: TPEC. 2010b) used the same rat strain, same route of exposure, and similar
range of exposure concentrations as Medinskv et al. (19991 and did not observe any dose-related
effects on testicular histopathology. Weng etal. f20141. however, found that Aldh2-/~ and
heterozygous 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 more susceptible to effects from ETBE exposure (discussed in
Section 1.3.3). The 14-day study by de Pevster et al. (2009) 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, although minimal effects were observed at otherwise toxic dose levels, the
available evidence is considered insufficient to identify male reproductive effects as a potential
human hazard of ETBE, so these effects are not carried forward as a hazard. While the Aldh2 KO
data suggest a potential sensitive subpopulation for male reproductive effects, these findings are
considered preliminary.
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.
i; Gaoua. 2004a). In addition, some evidence was obtained from two 90-day toxicity studies (TPEC.
2008b: Medinskv etal.. 1999: Bond etal.. 1996). 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 fBerger and Horner.
2003). These studies evaluated the effects of ETBE exposure on maternal body-weight change (Aso
etal.. 2014: Asano etal.. 2011: Fuiii etal.. 2010: TPEC. 2008e. h, i; Gaoua. 2004a. b), fertility, mating,
and pregnancy parameters (Fujii etal.. 2010: TPEC. 2008e: Gaoua. 2004b: Berger and Horner.
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2003). fecundity (Aso etal.. 2014: Asano etal.. 2011: Fujii etal.. 2010: TPEC. 2008e. h, i; Gaoua.
2004a. b), estrous cyclicity fFuiii etal.. 2010: TPEC. 2008e: Gaoua. 2004bl. and organ weights (Aso
etal.. 2014: Mivata etal.. 2013: Saito etal.. 2013: Suzuki etal.. 2012: Asano etal.. 2011: Fuiii etal..
2010: TPEC. 2010a. b, 2008b. c, e, h, i; Gaoua. 2004b: Medinskv et al.. 1999: Bond etal.. 19961.
ETBE-induced effects were examined in pregnant rats and rabbits and nonpregnant 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 1-16.
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 body weight (for the gravid uterus), was observed
following gestational exposure to 1,000 mg/kg-day ETBE from Gestation Day (GD) 5-19; however,
this effect was not observed in another developmental exposure study in which ETBE was
administered at the same dose and duration (Aso etal.. 2014: TPEC. 2008h). Further,
administration of ETBE during the premating through lactation periods in parental and F1
generations fFuiii etal.. 2010: TPEC. 2008e: Gaoua. 2004bl did not affect maternal body-weight
parameters in rats. Maternal body weight and corrected body-weight change during the entire
pregnancy (GD 0-28) were decreased in rabbits administered 1,000 mg/kg-day ETBE (Asano etal..
2011: TPEC. 2008i): 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) complicate the interpretation of this
effect ETBE did not affect indices of mating or fertility, and precoital times and gestation lengths
were similarly unaffected in rats in the parental fFuiii etal.. 2010: TPEC. 2008e: Gaoua. 2004bl and
the F1 generations f Gaoua. 2004bl. In addition, the number of corpora lutea in pregnant rats and
rabbits (Aso etal.. 2014: Asano etal.. 2011: TPEC. 2008h. i), the average estrous cycle length, and
the percent of females with normal estrous cycles (Fujii etal.. 2010: TPEC. 2008e) were not
significantly affected by ETBE when compared with control values. Further supporting these
findings, oocyte quality and fertilizability were shown to be unaffected by ETBE (Berger and
Horner. 20031. Litter size was evaluated by Fuiii etal. f20101. TPEC f2008el. Gaoua f2004bl. Aso et
al. f20141. TPEC f2008hl. Asano etal. (20111. and TPEC f2008il. 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 during gestation in rabbits (Asano etal.. 2011:
TPEC. 2008i) nor were ovary and uterine weights affected after exposure during premating through
lactation periods in rats fFuiii etal.. 2010: TPEC. 2008e). Consistent with these findings, ovary and
uterine weights in nonpregnant female rats were not affected by ETBE after 90-day inhalation
flPEC. 2008b: Medinskv etal.. 1999: Bond etal.. 19961.180-day oral fMivata etal.. 2013: TPEC.
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2008c). and 2-year oral (Suzuki etal.. 2012: TPEC. 2010a) studies. However, a 2-year inhalation
study in rats fSaito etal.. 2013: TPEC. 2010bl showed a significant increase in relative (but not
absolute) ovary weight at exposures of 1,500 and 5,000 ppm ETBE. It is possible the finding of
increased relative ovary weight was influenced by concurrent decreases in bodyweight (9-22%) at
these exposures.
Table 1-16. Evidence pertaining to female reproductive effects in animals
exposed to ETBE
Reference and study design
Results
Maternal body weight
Gaoua (2004a)
Rat, S-D
oral—gavage
PO, female (24/group): 0, 250,
500, or 1,000 mg/kg-d
dams exposed from GD 5 to
GD 19
Dose
(mg/kg-d)
Body wt change
± SD, GD 5-20
(g)
% Change
from control
Net body wt
change ± SD
(g)
% Change
from control
0
132 ± 22
-
61.8 ± 13
-
250
132 ± 12
-2
59.4 ±8.1
-4
500
134 ± 19
-1
60 ± 11.3
-3
1,000
120a± 15
-11
51.5a ± 10.3
-17
Gaoua (2004b)
Rat, S-D
oral—gavage
F0, male and female
(25/sex/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily for 18 wk from
10 wk premating until weaning
of F1 pups
Fl, male and female
(24-25/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Dose
(mg/kg-d)
F0:Body wt
change ± SD (g)
% Change
from control
Fl:Body wt
change ± SD
(g)
% Change
from control
0
132 ± 15
-
146 ± 21
-
250
134 ± 14
2
145 ± 15
-1
500
136 ± 25
3
141 ± 21
-3
1,000
136 ± 12
3
137 ± 12
-6
Aso et al. (2014); JPEC (2008h)
Rat, S-D
oral—gavage
female (24/group): 0,100, 300,
or 1,000 mg/kg-d
dams dosed daily from GD 5 to
GD 19
C-section GD 20
Dose
(mg/kg-d)
Body wt ± SD,
GD 5 (g)
Body wt ±
SD, GD 20 (g)
Body wt
change ± SD,
GD 5-20 (g)
% Change
from control
0
280.9 ± 16.7
394.4 ±26.9
113.5
-
100
273.4 ± 10.8
380.3 ± 23.9
106.9
-6
300
280 ± 13.4
389.8 ±25.9
109.8
-3
1,000
277.7 ± 15.9
382.4 ±27.1
104.7
-8
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Table 1-16. Evidence pertaining to female reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Fuiiietal. (2010); JPEC (2008e)
Rat, S-D
oral—gavage
F0, male and female
(24/sex/group): 0,100, 300, or
1,000 mg/kg-d; dosed daily for
17 wk, from 10 wk premating
to LD 21
Dose (mg/kg-d)
F0: Body wt change ± SD,
GD 5-20 (g)
% Change from control
0
124.9 ± 22
-
100
119.6 ±20.3
-4
300
135.2 ±21.5
8
1,000
140.2a ± 19.1
12
Asanoetal. (2011); JPEC
(2008i)
Rabbit, New Zealand White
oral—gavage
female (24/group): 0,100, 300,
or 1,000 mg/kg-d
dams dosed daily from GD 6 to
GD 27
C-section GD 28
Dose
(mg/kg-d)
Body wt
change ± SD,
GD 6-28 (kg)
% Change
from
control
Body wt
change ±
SD, GD 0-28
(kg)
Adjusted
body wt
change ±
SD (kg)
% Change
from
control
0
0.26 ±0.12
-
0.40 ±0.12
0.02 ±0.14
-
100
0.23 ±0.12
-12
0.35 ±0.12
-0.06 ±
0.12
-400
300
0.28 ±0.08
8
0.40 ± 0.08
0±0.1
-100
1,000
0.12 ±0.19
-54
0.25a ±0.21
-0.07 ±
0.19
-450
Fertility, mating, and pregnancy
Fuiiietal. (2010); JPEC (2008e)
Rat, S-D
oral—gavage
F0, male and female
(24/sex/group): 0,100, 300, or
1,000 mg/kg-d; dosed daily for
17 wk, from 10 wk premating
to LD 21
Dose (mg/kg-d)
Copulation indexb (%)
Fertility indexc (%)
0
100
87.5
100
95.8
100
300
100
95.8
1,000
100
91.7
Gaoua (2004b)
Rat, S-D
oral—gavage
F0, male and female
(25/sex/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily for 18 wk from
10 wk premating until weaning
of F1 pups
Fl, male and female
(24-25/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Dose
(mg/kg-d)
Pregnant/mated
females, F0
Fertility index, F0
(%)
Pregnant/
mated
females,
Fl
Fertility index,
Fl
(%)
0
23/25
92
22/25
88
250
21/25
84
22/24
92
500
22/25
88
22/25
88
1,000
25/25
100
22/23
96
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-16. Evidence pertaining to female reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Aso et al. (2014); JPEC (2008h)
Rat, S-D
oral—gavage
female (24/group): 0,100, 300,
or 1,000 mg/kg-d
dams dosed daily from GD 5 to
GD 19
C-section GD 20
Dose (mg/kg-d)
Mean no. corpora lutea ± SD
% Change from control
0
15.5 ± 1.54
-
100
14.1 ± 1.48
-9
300
14.4 ± 1.85
-7
1,000
14.6 ± 2.44
-6
Litter size
Fuiiietal. (2010); JPEC (2008e)
Rat, S-D
oral—gavage
F0, male and female
(24/sex/group): 0,100, 300, or
1,000 mg/kg-d; dosed daily for
17 wk, from 10 wk premating
to LD 21
Dose (mg/kg-d)
Mean no. pups delivered ± SD (kg)
% Change from control
0
11.8 ±3.2
-
100
10.4 ± 3.4
-12
300
12.1 ±2.3
3
1,000
13.0 ± 1.9
10
Gaoua (2004b)
Rat, S-D
oral—gavage
F0, male and female
(25/sex/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily for 18 wk from
10 wk premating until weaning
of F1 pups
Fl, male and female
(24-25/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Dose
(mg/kg-d)
Litter size at
birth, F0
% Change
from
control, F0
Pregnant/
mated females,
Fl
% Change
from control,
Fl
0
14.3
-
13.7
-
250
14.1
-1
13.7
0
500
14.9
4
13.7
0
1,000
14.2
-1
14
2
Aso et al. (2014); JPEC (2008h)
Rat, S-D
oral—gavage
female (24/group): 0,100, 300,
or 1,000 mg/kg-d
dams dosed daily from GD 5 to
GD 19
C-section GD 20
Dose
(mg/kg-d)
Mean no. live fetuses ± SD (kg)
% Change from control
0
13.6 ± 1.5
-
100
12.0 ±2.65
-12
300
12.6 ±2.58
-7
1,000
12.3 ±2.8
-10
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Table 1-16. Evidence pertaining to female reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Asanoetal. (2011); JPEC
(2008i)
Rabbit, New Zealand White
oral—gavage
female (24/group): 0,100, 300,
or 1,000 mg/kg-d
dams dosed daily from GD 6 to
GD 27
C-section GD 28
Dose
(mg/kg-d)
Mean no. live fetuses ± SD (kg)
% Change from control
0
7.8 ±3.1
-
100
7.9 ±3.2
1
300
8.4 ±2.0
8
1,000
6.9 ±3.2
-12
Gestation length
Fuiiietal. (2010); JPEC (2008e)
Rat, S-D
oral—gavage
F0, male and female
(24/sex/group): 0,100, 300, or
1,000 mg/kg-d; dosed daily for
17 wk, from 10 wk premating
to LD 21
Dose
(mg/kg-d)
Mean gestation length ± SD (d)
% Change from control
0
22.2 ±0.4
-
100
22.1 ±0.4
0
300
22.2 ±0.4
0
1,000
22.6 ±0.5
2
Gaoua (2004b)
Rat, S-D
oral—gavage
F0, male and female
(25/sex/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily for 18 wk from
10 wk premating until weaning
of F1 pups
Fl, male and female
(24-25/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Dose
(mg/kg-d)
Gestation length
(d), F0
% Change
from
control, F0
Gestation
length (d), Fl
% Change
from control,
Fl
0
21.7
-
21.5
-
250
21.5
-1
21.6
0
500
21.5
-1
21.6
0
1,000
21.8
0
21.6
0
Estrous cyclicity
Fuiiietal. (2010); JPEC (2008e)
Rat, S-D
oral—gavage
F0, male and female
(24/sex/group): 0,100, 300, or
1,000 mg/kg-d; dosed daily for
17 wk, from 10 wk premating
to LD 21
Dose
(mg/kg-d)
% Females w/normal estrous
cycles, F0
Mean estrous
cycle length
± SD
(d)
% Change
from control
0
91.7
4.03 ± 0.09
-
100
97.1
4.1 ±0.29
2
300
97.1
4.06 ±0.17
1
1,000
95.8
4.29 ±0.61
6
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Table 1-16. Evidence pertaining to female reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Organ weights
Fuiiietal. (2010); JPEC (2008e)
Rat, S-D
oral—gavage
F0, male and female
(24/sex/group): 0,100, 300, or
1,000 mg/kg-d; dosed daily for
17 wk, from 10 wk premating
to LD 21
Absolute weight
Dose
(mg/kg-d)
Mean ovary wt ±
SD (mg)
% Change
from control
Mean uterus
wt ± SD (mg)
% Change
from control
0
98.8 ± 14.9
-
468 ±68
-
100
92.5 ± 16.6
-6
513 ±151
10
300
95.3 ± 11.1
-4
523 ±157
12
1,000
100.9 ± 16.9
2
516 ±136
10
Relative weight
Dose
(mg/kg-d)
Mean ovary wt ±
SD (mg)
% Change
from control
Mean uterus
wt ± SD (mg)
% Change
from 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)
Rat, S-D
oral—gavage
F0, male and female
(19-25/sex/group): 0, 250, 500,
or 1,000 mg/kg-d
dosed daily for 18 wk from
10 wk premating until weaning
of F1 pups
Fl, male and female
(19-25/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Dose
(mg/kg-d)
Mean ovary wt.
± SD (g)
% Change
from control
Mean uterus
wt. ± SD (g)
% Change
from control
Absolute weight, F0
0
0.168 ±0.025
-
0.54 ± 0.096
-
250
0.167 ±0.027
-1
0.587 ±0.231
9
500
0.167 ±0.022
-1
0.483 ±0.102
-11
1,000
0.164 ±0.023
-2
0.576 ±0.218
7
Absolute weight, Fl
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
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Table 1-16. Evidence pertaining to female reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
Medinskv et al. (1999); U.S.
EPA (1997)
Rat, F344
inhalation—vapor
male (48/group): 0, 500,1,750,
or 5,000 ppm
(0, 2,090, 7,320, or
20,900 mg/m3);d female
(48/group): 0, 500,1,750, or
5,000 ppm
(0, 2,090, 7,320, or
20,900 mg/m3)d
dynamic whole-body chamber;
6 h/d, 5 d/wk for 13 wk
Dose (mg/m3)
Mean ovary wt ± SD (g)
% Change from control
0
0.085 ± 0.022
-
2,090
0.095 ±0.016
12
7,320
0.088 ±0.12
4
20,900
0.090 ±0.19
6
Asano et al. (2011); JPEC
(2008i)
Rabbit, New Zealand White
oral—gavage
female (24/group): 0,100, 300,
or 1,000 mg/kg-d
dams dosed daily from GD 6 to
GD 27
C-section GD 28
Dose
(mg/kg-d)
Gravid uterus wt ± SD (g)
% Change from control
0
383 ± 98
-
100
398 ±128
4
300
403 ±91
5
1,000
323 ±128
-16
Mivata et al. (2013); JPEC
(2008c)
Rat, S-D
oral—gavage
male (15/group): 0, 5, 25,100,
or 400 mg/kg-d; female
(15/group): 0, 5, 25,100, or
400 mg/kg-d
daily for 180 d
Dose
(mg/kg-d)
Mean absolute ovary wt ± SD
(mg)
% Change from control
0
70.0 ± 18.7
-
5
71.0 ±21.7
1
25
73.8 ± 16.6
5
100
67.7 ± 17.7
-3
400
76.6 ± 18.2
9
Dose
(mg/kg-d)
Mean relative ovary wt ± SD
(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
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-16. Evidence pertaining to female reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
JPEC (2008b)
% Change
% Change
Rat, S-D
Dose
Mean ovary
from
Mean uterus
from
inhalation—vapor
(mg/m3)
n
wt ± SD (mg)
control
wt ± SD (g)
control
male (10/group): 0,150, 500,
1,500, or 5,000 ppm
(0, 627, 2,090, 6,270, or
20,900 mg/m3);d female
(10/group): 0, 150, 500, 1,500,
or 5,000 ppm (0, 627, 2,090,
6,270, or 20,900 mg/m3)d
dynamic whole-body chamber;
Absolute weight, Day 92
0
10
91.47 ± 10.26
-
0.709 ±0.222
-
627
10
87.36 ± 15.83
0
0.819 ±0.38
16
2,090
10
84.92 ± 16.91
0
0.654 ±0.159
-8
6,270
10
78.39 ±9.83
0
0.712 ±0.198
0
6 h/d, 5 d/wk for 13 wk;
generation method, analytical
concentration, and method
20,900
10
91.94 ±21.84
0
0.702 ± 0.205
-1
Absolute weight, Day 120
reported
0
6
82.82 ± 17.89
-
0.965 ±0.332
-
627
-
-
-
-
-
2,090
-
-
-
-
-
6,270
-
-
-
-
-
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
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-16. Evidence pertaining to female reproductive effects in animals
exposed to ETBE (continued)
Reference and study design
Results
JPEC (2010a)
Rat, F344
oral—water
male (50/group): 0, 625, 2,500,
or 10,000 ppm (0, 28,121, or
542 mg/kg-d);e female
(50/group): 0, 625, 2,500, or
10,000 ppm (0, 46,171, or
560 mg/kg-d)e
daily for 104 wk
Dose
(mg/kg-d)
Mean ovary 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
LD = lactation day; PND = postnatal day; SD = standard deviation.
% change from control = [(control value - treated group value) -f control value] x 100.
Absolute change from control (%) = control value (%) - treated group value (%).
aResult is statistically significant (p < 0.05) based on analysis of data by study authors.
Population index (%) = (no. of rats with successful copulation 4 no. of rats paired) x 100.
fertility index (%) = (no. females pregnant or no. of males sired 4 no. of rats with successful copulation) x 100.
d4.18 mg/m3 = 1 ppm.
Conversion performed by the study authors.
-For controls, no response relevant; for other doses, no quantitative response 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
Pregnancy Outcomes
FO Female rat; copulation; one-gen repro (C)
FO Female rat; fertility; one-gen repro (C)
FO Female rat; fertility; two-gen repro (E)
F1 Female rat; fertility; two-gen repro (E)
Female rat; maternal body weight gain; GD 5-19 (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; GD 5-19 (B)
FO Female rat; maternal body weight gain; one-gen
repro (C)
Female rabbit; maternal body weight gain; GD 6-27
(A)
Female rabbit; gravid uterine weight; GD 6-27 (A)
FO Female rat; gestation length; one-gen repro (C)
FO Female rat; gestation length; two-gen repro (E)
F1 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 repro (E)
Female rat; litter size; GD 5-19 (B)
Female rabbit; litter size; GD 6-27 (A)
FO Female rat; estrous cydicity; one-gen repro (C)
*maternal weight gain significantly increased in this
study whereas other studies showed a significant
decrease
I—I I I I i
~ B -J
~ B f]
~ B-
~ B f]
~—B—e:
-B-
~ B E]
~ B E]
~ B E]
~ B E]
[) B-
-E]
100
1,000
Dose (mg/kg-day)
Sources: (A) Asano et. al„ 2011; JPEC, 2008h (B) Aso et al„ 2014; JPEC, 2008g(C) Fujii et al„ 2010; JPEC, 2008e
(D) Gaoua, 2004a (E) Gaoua, 2004b
Figure 1-13. Exposure-response array of female reproductive effects
following oral exposure to ETBE.
10,000
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Toxicological Review of Ethyl Tertiary Butyl Ether
¦ = 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)
Female rat; ovary wt. (absolute); 13 wks (A)
Female rat; ovary wt. (relative); 13 wks (A)
Female rat; uterine wt. (absolute); 13 wks (A)
Female rat; uterine wt. (relative); 13 wks (A)
Source: (A) JPEC 2008b (B) Medinsky et al., 1999; Bond et al., 1996b
Figure 1-14. Exposure-response array of female reproductive effects
following inhalation exposure to ETBE.
100 1,000 10,000 100,000
Exposure Concentration (mg/m3)
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Toxicological Review of Ethyl Tertiary Butyl Ether
Mechanistic Evidence of Female Reproductive Effects
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, but not absolute 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. Collectively, although minimal effects were
observed at otherwise toxic dose levels, the available evidence is considered insufficient to identify
female reproductive effects as a potential human hazard of ETBE, 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 S-D rats or New Zealand White
rabbits in which ETBE was administered via gavage. These consisted of three prenatal
developmental toxicity studies [two in rats: Aso etal. f2014-1: TPEC f2008hl: Gaoua f2004al and one
study in rabbits: Asano etal. (2011): TPEC (2008i)]. a one-generation reproductive toxicity study in
rats (Fujii etal.. 2010: TPEC. 2008e). and a two-generation reproductive toxicity study in rats
(Gaoua. 2004a). 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-day, the recommended limit dose for prenatal developmental toxicology studies
COECD. 2001: U.S. EPA. 1998b1.
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 1-17.
Evidence of the effects of ETBE treatment on pre- or postnatal survival was minimal. In the
developmental toxicity study in rats by Aso etal. (2014) and 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
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Toxicological Review of Ethyl Tertiary Butyl Ether
100 mg/kg-day and 21.2% 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 [Gaoua f2004al and Asano etal. f20111: TPEC f2008il. respectively]. Postnatal survival
was not affected by ETBE treatment in either the first or second generation of the reproductive
toxicity study by Gaoua f2004bl. 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 (Fuiii etal.. 2010: IPEC. 2008e], there was evidence of a nonsignificant
decrease (10.5% as compared with control) in the Postnatal Day (PND) 4 viability index at
1,000 mg/kg-day. Examination of the individual animal data indicated the majority of pup deaths
occurred from the total loss of three litters 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 in which 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 the control and ETBE-treated groups in the prenatal
developmental toxicity studies in rats and rabbits fAso etal.. 2014: Asano etal.. 2011: TPEC. 2008h.
i; Gaoua. 2004al. Similarly, pup weights from PND 0-21 were not affected by treatment in the
reproductive toxicity studies fFuiii etal.. 2010: TPEC. 2008e: Gaoua. 2004bl. Additionally, Fuiii etal.
(2010) and TPEC (2008e) observed no effects in the completion rate 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 (Fuiii etal.. 2010: TPEC. 2008e: Gaoua.
2004b); 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, that were observed in the adult F1 offspring of the two-generation
reproduction study (Gaoua. 2004b). The findings in the F1 adults were similar to those in the
parental (F0) adults, indicating an absence of lifestage-related susceptibility for these outcomes.
No evidence existed of treatment-related effects on postnatal morphological assessments.
These assessments consisted of anogenital distance measurements in F1 and F2 pups on PND 1
fGaoua. 2004bl and the age of sexual maturation in F1 pups (preputial separation in males and
vaginal opening in females) fFuiii etal.. 2010: TPEC. 2008e: Gaoua. 2004bl.
In the prenatal developmental toxicity studies with ETBE (Aso etal.. 2014: Asano etal..
2011: TPEC. 2008h. ij Gaoua. 2004a). there was sporadic evidence of treatment-related alterations
in fetal development at 1,000 mg/kg-day but no consistent pattern of effect.
In Aso etal. (2014). a >threefold increase in the number and percent of rat fetuses with
skeletal variations was noted at 1,000 mg/kg-day compared with control. Examination of the
individual litter data revealed that this increase was primarily due to a statistically significant
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Toxicological Review of Ethyl Tertiary Butyl Ether
>sixfold increase in the number of fetuses (and >threefold 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 et al. fl9911].
Nevertheless, the increase in the incidence of this finding is statistically significant compared with
the concurrent control, which is considered more relevant and preferable than using historical
controls, 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 4th metacarpal as compared with 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. (2011) and TPEC (2008i). 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 significance. Also, in Asano etal. f20111 and
TPEC f2008il. a 66% increase in the number of rabbit fetuses with skeletal variations at
1,000 mg/kg-day as compared with control was found to be primarily attributed to incidences of
unossified talus (in 12 fetuses, 6 litters).
In examining reproductive toxicity, Fuiii etal. (2010). TPEC (2008e). and Gaoua (2004b)
conducted a limited evaluation of postnatal functional neurological development in F1 male and
female offspring. No treatment-related effects were found in reflex ontogeny assessments, which
included surface righting reflex on PND 5 fFuiii etal.. 2010: TPEC. 2008e: Gaoua. 2004bl. negative
geotaxis on PND 8 fFuiii etal.. 2010: TPEC. 2008el. cliff avoidance on PND 11 f Gaoua. 2004bl. and
air righting reflex on PND 17 (Gaoua. 2004b) or PND 18 (Fujii etal.. 2010: TPEC. 2008e). Gaoua
(2004b) 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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Table 1-17. Evidence pertaining to developmental effects in animals following
exposure to ETBE
Reference and study design
Results
Prenatal survival
Aso et al. (2014); JPEC (2008h)
Rat, S-D
oral—gavage
female (24/group): 0,100, 300,
or 1,000 mg/kg-d
dams dosed daily from GD 5 to
GD 19
C-section GD 20
Dose
(mg/kg-d)
No.
litters
No.
preimplantation
loss
%
Change
from
control
%
Preimplantation
loss3
%
Change
from
control
0
21
22
-
6.6
-
100
22
25
13.6
9.1
37.9
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)
Rat, S-D
oral—gavage
female (24/group): 0, 250, 500,
or 1,000 mg/kg-d
dosed daily from GD 5 to GD 19
C-section GD 20
Dose
(mg/kg-d)
No.
litters
No. preimplantation loss
% Preimplantation lossa
0
21
48
17.8
250
19
36
14.9
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)
Rabbit, New Zealand White
oral—gavage
female (24/group): 0,100, 300,
or 1,000 mg/kg-d
dams dosed daily from GD 6 to
GD 27
C-section GD 28
Dose
(mg/kg-d)
No.
litters
% Preimplantation lossa
% Postimplantation lossb
0
22
19.6
11.0
100
22
15.3
11.3
300
20
10.7
7.0
1,000
23
22.9
8.7
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-17. Evidence pertaining to developmental effects in animals
following exposure to ETBE (continued)
Reference and study design
Results
Postnatal survival
Fuiiietal. (2010); JPEC (2008e)
Rat, S-D
oral—gavage
F0, male and female
(24/sex/group): 0,100, 300, or
1,000 mg/kg-d; dosed daily for
17 wk, from 10 wk premating to
LD 21
Dose
(mg/kg-d)
Viability index
PND0±SD
Viability index
PND 4 ± SD
% Change
from
control
(PND 4)
Total litter loss
(PND 0-4)c
0
98.9 ±3.7
97.4 ±4.7
-
0
100
97.9 ±5.6
96.7 ±8.1
-0.7
0
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)
Viability 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)
Rat, S-D
oral—gavage
F0, male and female
(25/sex/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily for 18 wk from
10 wk premating until weaning
of F1 pups
Fl, male and female
(24-25/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Dose
(mg/kg-d)
Viability index
PND 0
Viability
index
PND 4
Total litter loss
(PND 0-4)
Viability index
PND 21
Fl
0
100
97.6
0
94.6
250
100
92.9
1
91.7
500
100
82.3
0
96.1
1,000
100
97.7
1
99.5
F2
0
100
97.6
0
97.6
250
100
94.8
0
98.8
500
100
97.0
3
100
1,000
100
92.9
0
99.3
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-17. Evidence pertaining to developmental effects in animals
following exposure to ETBE (continued)
Reference and study design
Results
Prenatal growth
Aso et al. (2014); JPEC (2008h)
Rat, S-D
oral—gavage
female (24/group): 0,100, 300,
or 1,000 mg/kg-d
dams dosed daily from GD 5 to
GD 19
C-section GD 20
Dose
(mg/kg-d)
No. litters
Mean fetal weight ± SD
male (g)
Mean fetal weight ± SD
female (g)
0
21
4.1 ±0.3
3.89 ±0.25
100
22
4.14 ±0.33
3.92 ±0.23
300
20
4.23 ±0.22
4.01 ±0.22
1,000
22
4.14 ±0.34
3.91 ±0.39
Gaoua (2004b)
Rat, S-D
oral—gavage
F0, male and female
(25/sex/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily for 18 wk from
10 wk premating until weaning
of F1 pups
Fl, male and female
(24-25/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Dose
(mg/kg-d)
No. litters
Mean fetal weight ± SD
male (g)
Mean fetal weight ± SD
female (g)
0
21
3.92 ±0.58
3.77 ±0.5
250
19
4.03 ±0.32
3.82 ±0.33
500
20
3.94 ±0.35
3.75 ±0.32
1,000
22
3.91 ±0.33
3.66 ±0.39
Asano et al. (2011); JPEC (2008i)
Rabbit, New Zealand White
oral—gavage
female (24/group): 0,100, 300,
or 1,000 mg/kg-d
dams dosed daily from GD 6 to
GD 27
C-section GD 28
Dose
(mg/kg-d)
No. litters
Mean fetal weight ± SD
male (g)
Mean fetal weight ± SD
female (g)
0
22
33.5 ±4.1
31.5 ±3.7
100
22
33.4 ±6.2
31.5 ±4.8
300
20
33.9 ±2.5
32.0 ±3.6
1,000
23
32.3 ±6.5
30.1 ±6.0
1-100
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Toxicological Review of Ethyl Tertiary Butyl Ether
Table 1-17. Evidence pertaining to developmental effects in animals
following exposure to ETBE (continued)
Reference and study design
Results
Postnatal growth
Fuiiietal. (2010); JPEC (2008e)
Rat, S-D
oral—gavage
F0, male and female
(24/sex/group): 0,100, 300, or
1,000 mg/kg-d; dosed daily for
17 wk, from 10 wk premating to
LD 21
Dose
(mg/kg-d)
No.
litters
Mean ± SD
PND 0(g)
Mean ± SD
PND 4 precull (g)
Mean ± SD
PND 21 (g)
Fl—male pup weight
0
21
6.9 ±0.7
11.0 ±2.0
61.3 ±6.3
100
22
6.9 ±0.8
11.0 ± 1.8
61.0 ±7.0
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
Gaoua (2004b)
Rat, S-D
oral—gavage
F0, male and female
(25/sex/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily for 18 wk from
10 wk premating until weaning
of F1 pups
Fl, male and female
(24-25/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Dose
(mg/kg-d)
Mean ± SD
PND 1 (g)
Mean ± SD
PND 4 precull (g)
Mean ± SD
PND 21 (g)
Fl—male pup weight
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
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Table 1-17. Evidence pertaining to developmental effects in animals
following exposure to ETBE (continued)
Reference and study design
Results
Gaoua (2004b) (continued)
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, S-D
oral—gavage
female (24/group): 0,100, 300,
or 1,000 mg/kg-d
dams dosed daily from GD 5 to
GD 19
C-section GD 20
Dose
(mg/kg-d)
No. fetuses
(litters)d
No. fetuses
examined for
visceral
anomalies
No. fetuses
with visceral
malformations
No. fetuses
with visceral
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)
Dose
(mg/kg-d)
No. fetuses
examined
for skeletal
anomalies
No. fetuses with
skeletal
malformations
No. fetuses
with skeletal
variations
% Fetuses
(litters) with
skeletal
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)
Dose (mg/kg-d)
No. fetuses (litters) with
rudimentary lumbar rib
% Fetuses (litters) with
rudimentary lumbar rib
0
4(4)
2.9 (19.0)
100
0
0(0)
300
2(2)
1.7 (10.0)
1,000
25e (11)
19.le (50.0)
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Table 1-17. Evidence pertaining to developmental effects in animals
following exposure to ETBE (continued)
Reference and study design
Results
Gaoua (2004a)
Rat, S-D
oral—gavage
female (24/group): 0, 250, 500,
or 1,000 mg/kg-d
dosed daily from GD 5 to GD 19
C-section GD 20
Dose
(mg/kg-d)
No. fetuses
(litters)d
No. fetuses
with external
malformations
No. fetuses
examined for
visceral
anomalies
No. fetuses with
visceral
malformations
0
255 (21)
0
120
0
250
226 (19)
1(1)
109
0
500
246 (20)
0
116
0
1,000
258 (22)
0
122
1(1)
Dose
(mg/kg-d)
No. fetuses
with visceral
variations
No. fetuses
examined for
skeletal
anomalies
No. fetuses
with skeletal
malformations
No. fetuses with
skeletal
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)
Dose
(mg/kg-d)
No. fetuses with
unossified 4th
metacarpal
% Change from control
% Fetuses with
unossified 4th
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
43e (12)
37.2
31.6 (54.5)
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Table 1-17. Evidence pertaining to developmental effects in animals
following exposure to ETBE (continued)
Reference and study design
Results
Asano et al. (2011); JPEC (2008i)
Rabbit, New Zealand White
oral—gavage
female (24/group): 0,100, 300,
or 1,000 mg/kg-d
dams dosed daily from GD 6 to
GD 27
C-section GD 28
Dose
(mg/kg-d)
No.
fetuses
(litters)d
No. fetuses with
external
malformations
No. fetuses
with visceral
malformations
No. fetuses
with skeletal
malformations
0
171 (22)
0
1(1)
5(4)
100
174 (22)f
1(1)
1(1)
4(4)
300
167 (20)
0
1(1)
3(2)
1,000
159 (23)f
1(1)
3(2)
8(5)
Dose
(mg/kg-d)
No. fetuses with
skeletal variations
Absent right
atrioventricular
valve
% Change from 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)
Rat, S-D
oral—gavage
F0, male and female
(24/sex/group): 0,100, 300, or
1,000 mg/kg-d; dosed daily for
17 wk, from 10 wk premating to
LD 21
Dose
(mg/kg-d)
No.
litters
Male preputial
separation—age (d)
mean ± SD
Female vaginal
opening—age (d)
mean ± SD
F1
0
21
41.0 ± 1.7
31.2 ± 1.4
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
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Table 1-17. Evidence pertaining to developmental effects in animals
following exposure to ETBE (continued)
Reference and study design
Results
Gaoua (2004b)
Rat, S-D
oral—gavage
F0, male and female
(25/sex/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily for 18 wk from
10 wk premating until weaning
of F1 pups
Fl, male and female
(24-25/group): 0, 250, 500, or
1,000 mg/kg-d
dosed daily from PND 22 until
weaning of F2 pups
Dose
(mg/kg-d)
No.
litters
Anogenital
distance®—males (PND 1)
mean ± SD
Anogenital
distance®—females (PND 1)
mean ± SD
Fl
0
21
2.48 ±0.18
1.53 ±0.18
250
22
2.45 ±0.17
1.5 ±0.14
500
23
2.4 ±0.21
1.45 ±0.14
1,000
20
2.43 ±0.15
1.44 ± 0.2
F2
0
21
2.41 ±0.18
1.51 ±0.18
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
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Table 1-17. Evidence pertaining to developmental effects in animals
following exposure to ETBE (continued)
Reference and study design
Results
Gaoua (2004b) (continued)
Dose
(mg/kg-d)
No.
litters
Male preputial
separation—age
(d)—mean ± SD
Female vaginal
opening—age (d)—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
LD = lactation day; SD = standard deviation.
% change from control = ([control value - treated group value] -f control value) x 100.
Absolute change from control (%) = control value (%) - treated group value (%).
aPercent preimplantation loss = (no. preimplantation embryonic loss 4 no. corpora lutea) x 100.
bPercent postimplantation loss = (no. resorptions and dead fetuses 4 no. implantations) x 100.
cTwo 1,000 mg/kg-day 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.
eResult is statistically significant (p < 0.05) based on analysis of data by the study authors.
fNo. of fetuses examined for visceral and skeletal anomalies at 100 and 1,000 mg/kg-day were 173 and 158,
respectively, because fetuses with external malformations were excluded.
gAGD/cube root of body weight.
-For controls, no response relevant; for other doses, no quantitative response 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
Survival -
Prenatal
Rat; preimplantation loss; GD 5-19 (B)
Rat; preimplantation loss; GD 5-19 (D)
Rabbit; preimplantation loss; GD6-27 (A)
Rat; postimplantation loss; GD 5-19 (B)
Rat; postimplantation loss; GD 5-19 (D)
Rabbit; postimplantation loss; GD 6-27 (A)
~ ~ ~
-0
~ ~ ~
Survival -
Postnatal
Growth -
Prenatal
Growth -
Postnatal
Morphology-
Prenatal
Rat; decreased viability, PND 4 or PND 21; one-gen repro (C)
Rat; decreased viability, PND 4 or PND 21; two-gen repro (E)
Rat; decreased mean fetal wt.; GD 5-19 (B)
Rat; decreased mean fetal wt.; GD 5-19 (D)
Rabbit; decreased mean fetal wt.; GD 6-27 (A)
Rat; decreased mean pup wt.; one-gen repro (C)
Rat; decreased mean pup wt.; two-gen repro (E)
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: rudimentary lumbar ribs; GD 5-19 (D)
Rabbit; skeletal variation: 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 variation: unossified 4th metatarsal; GD 6-27
(A)
Rat; altered F1 or F2 anogenital distance [PND 1); one-gen
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
in
Rat; altered F1 reflex ontogeny, acoustic startle response,
Functional pupil constriction, motor activity; one-gen repro (C)
(Neurological) Rat; altered F1 reflex ontogeny, acoustic startle response.
Development pupil constriction, motor activity; two-gen repro (E)
-~
~ ~ ~
~—b—a
~-
~ ~ ~
~ ~ ~
~ ~ ~
&-B-
~ ~ ~
~ ~ ~
~ ~ ~
10 100
Dose (mg/kg-day)
1,000
10,000
Sources: (A) Asano et al. 2011; JPEC, 2008h (B) Aso et al. 2014; JPEC, 2008g(C] 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.
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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 of ETBE 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 fAso etal.. 2014: TPEC. 2008hl and unossified 4th metatarsal fGaoua.
2004b) in two rat studies and unossified talus and absent right atrioventricular valve in a rabbit
study (Asano etal.. 2011: TPEC. 2008i). 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 of ETBE, 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 animal data only: three 2-year
studies (two oral, one inhalation), and two "initiation, promotion" cancer bioassays performed in
rats (Hagiwara etal.. 2013: Saito etal.. 2013: Suzuki etal.. 2012: Hagiwara etal.. 2011: Malarkev
and Bucher. 2011: TPEC. 2010a. b; Maltoni etal.. 19991 (see Table 1-18, Table 1-19; Figure 1-16,
Figure 1-17). Interpretation of the study results reported by Maltoni et al. f 19991 is complicated by
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the nonstandard histopathological diagnoses used and the greater than expected mortality in
treated groups and controls compared with other laboratories. Survival was reduced at the highest
exposure in males and females after 72 weeks (data not shown) and after 104 weeks. Survival in
the controls was approximately 25% in males and 28% in females—much lower survival rates than
expected for a 2-year study fMaltoni etal.. 19991. The high mortality in this study was potentially
due to chronic respiratory infections in the controls and treated groups (Malarkev and Bucher.
20111. ultimately limiting the ability of this study to predict potential carcinogenicity.
In response to the concerns regarding high mortality and the use of nonstandard
histopathological diagnoses, a pathology working group sponsored by EPA and the 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 et al. (1999) 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 S-D rats in the high-dose group fMaltoni 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
carcinomas were observed in uterine/vaginal tissue, but no significant trend was observed
(Malarkev and Bucher. 2011).
Several initiation-promotion studies have been conducted with ETBE (Hagiwara etal..
2015: Hagiwara etal.. 2013: Hagiwara etal.. 2011). While chronic cancer bioassays are considered
key data for the evaluation of carcinogenicity, other types of studies, such as initiation-promotion
studies, are considered supplemental lines of information that can aid in the interpretation of more
standard toxicological evidence (e.g., rodent chronic bioassays), especially regarding potential
modes of action (U.S. EPA. 2005a). 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 etal..2011):
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. 2011): comparability with Japan Petroleum Energy Center (JPEC) controls unknown]. While
increased tumors were observed with ETBE following administration of tumor initiators, it is
important to recognize the limitations of these experimental protocols. Such limitations include
experimental manipulation of the carcinogenic process, a generally less than chronic exposure
duration, and smaller groups of animals. Male F344 rats (n = 12), exposed to ETBE via gavage for
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23 weeks (in the absence of DMBDD treatment), did not show significant tumor development in any
organs evaluated fHagiwara etal.. 20111. Increased tumorigenesis in these tissues was not
reported following 2 years of drinking water or inhalation exposure to ETBE alone in male or
female F344 rats fSaito etal.. 2013: Suzuki etal.. 2012: TPEC. 2010bl.
Mechanistic Evidence
The available mechanistic evidence was previously discussed in the context of kidney and
liver tumors (Sections 1.1.1 and 1.1.2). Aside from (predominantly negative) genotoxicity testing
results, which are 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 (DMBDD) increased the incidence of thyroid
adenomas or carcinomas, colon adenomas or carcinomas, forestomach papillomas, and urinary
bladder carcinomas in male rats fHagiwar a et al.. 2011: TPEC. 2008dl. Confidence in the data
demonstrating an increase in the incidence of schwannomas is reduced because of the lack of a
dose-response in S-D rats and lack of a similar effect reported in F344 rats from two other
well-conducted 2-year studies, or in F344 or Wistar rats from the two-stage subchronic cancer
bioassays. The hazard and dose-response conclusions regarding these carcinomas and adenomas
associated with ETBE exposure are further discussed as part of the overall weight of evidence for
carcinogenicity in Section 1.3.2.
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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
Results
Thyroid adenomas/adenocarcinomas
JPEC (2010a); Suzuki etal.
(2012)
Rat, F344
oral—water
male (50/group): 0, 625, 2,500,
or 10,000 ppm (0, 28,121, or
542 mg/kg-d);a female
(50/group): 0, 625, 2,500, or
10,000 ppm (0, 46,171, or
560 mg/kg-d)a
daily for 104 wk
Male
Female
Dose
(mg/kg-d)
Thyroid
follicular
adeno-
carcinoma
Thyroid
follicular
adenoma
Dose
(mg/kg-d)
Thyroid
follicular
adeno-
carcinoma
Thyroid
follicular
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, F344
inhalation—vapor
male (50/group): 0, 500,1,500,
or 5,000 ppm (0, 2,090, 6,270,
or 20,900 mg/m3);b female
(50/group): 0, 500,1,500, or
5,000 ppm (0, 2,090, 6,270, or
20,900 mg/m3)b
dynamic whole-body
inhalation; 6 h/d, 5 d/wk for
104 wk; generation method,
analytical concentration, and
method reported
Male
Female
Dose
(mg/m3)
Thyroid
follicular
adeno-
carcinoma
Thyroid
follicular
adenoma
Dose
(mg/m3)
Thyroid
follicular
adeno-
carcinoma
Thyroid
follicular
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, S-D
oral—gavage
male (60/group): 0, 250, or
1,000 mg/kg-d; female
(60/group): 0, 250, or
1,000 mg/kg-d
4 d/wk for 104 wk; observed
until natural death
Note: Tumor data not
reanalyzed bv Malarkev and
Bucher (2011)
Male
Female
Dose
(mg/kg-d)
Thyroid adenocarcinoma
Dose
(mg/kg-d)
Thyroid
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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Table 1-18. Evidence pertaining to carcinogenic effects (in tissues other than
liver or kidney) in animals exposed to ETBE (continued)
Reference and study design
Results
Endometrial/uterine carcinogenic effects
JPEC (2010a);Suzuki etal.
(2012)
Rat, F344
oral—water
female (50/group): 0, 625,
2,500, or 10,000 ppm (0, 46,
171, or 560 mg/kg-d)a
daily for 104 wk
Female
Dose
(mg/kg-d)
Endometrial stromal
sarcoma
Uterine
adenocarcinoma
Uterine fibroma
0
6/50
1/50
1/50
46
9/50
0/50
0/50
171
3/50
2/50
0/50
560
7/50
2/50
0/50
JPEC (2010b);Saito et al. (2013)
Rat, F344
inhalation—vapor
female (50/group): 0, 500,
1,500, or 5,000 ppm (0, 2,090,
6,270, or 20,900 mg/m3)b
dynamic whole-body
inhalation; 6 h/d, 5 d/wk for
104 wk; generation method,
analytical concentration, and
method reported
Female
Dose
(mg/m3)
Endometrial stromal sarcoma
Uterine
adenocarcinoma
0
2/50
2/50
2,090
2/50
3/50
6,270
3/50
1/50
20,900
2/50
4/50
Malarkev and Bucher (2011);
Maltoni et al. (1999)
Rat, S-D
oral—gavage
female (60/group): 0, 250, or
1,000 mg/kg-d
reanalvsis of data from Maltoni
et al. (1999) for which animals
were dosed 4 d/wk for 104 wk
Female
Dose
(mg/kg-d)
Carcinoma
of the
uterus/
vagina
Uterine leio-
myoma
Uterine leio-
myosarcoma
Schwannoma
of the
uterus/
vagina
Uterine
carci-
noma
0
0/60
0/60
1/60
0/60
0/60
250
1/60
0/60
0/60
7/60
1/60
1,000
0/60
3/60
0/60
2/60
0/60
Conversion performed by the study authors.
b4.18 mg/m3 = 1 ppm.
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Table 1-19. Supplemental evidence pertaining to ETBE promotion of
mutagen-initiated tumors in animals
Reference and dosing protocol
Results by endpoint
Colon adenoma or carcinoma
Hagiwara et al. (2011); JPEC (2008d)
Rat, F344
oral—gavage
male (30/group): 0, 300, or 1,000 mg/kg-d
daily for 23 wk following a 4-wk tumor initiation by
DMBDDa
Male
Dose (mg/kg-d)
Response (incidence)
0
25/30
300
21/30
1,000
28/30b
0C
0/12
1,000°
0/12
Forestomach papillomas or hyperplasia
Hagiwara et al. (2011); JPEC (2008d)
Rat, F344
oral—gavage
male (30/group): 0, 300, or 1,000 mg/kg-d
daily for 23 wk following a 4-wk tumor initiation by
DMBDDa
Male
Dose (mg/kg-d)
Response (incidence)
0
0/30
300
6/30b
1,000
6/30b
0C
0/12
1,000°
0/12
Thyroid gland adenoma or carcinoma
Hagiwara et al. (2011); JPEC (2008d)
Rat, F344
oral—gavage
male (30/group): 0, 300, or 1,000 mg/kg-d
daily for 23 wk following a 4-wk tumor initiation by
DMBDDa
Male
Dose (mg/kg-d)
Response (incidence)
0
8/30
300
17/30b
1,000
20/30b
0C
0/12
1,000°
0/12
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Table 1-19. Supplemental evidence pertaining to ETBE promotion of
mutagen-initiated tumors in animals (continued)
Reference and dosing protocol
Results by endpoint
Urinary bladder carcinoma
Hagiwara et al. (2013)
Rat, F344/DuCrlCrlj
oral—water
male (30/group): 0,100, 300, 500, or
1,000 mg/kg-d
daily for 31 wk beginning 1 wk after a 4-wk
exposure to BBN
Male
Dose (mg/kg-d)
Response (incidence)
0
5/30
100
7/30
300
6/30
500
14/30b
1,000
9/26
Urinary bladder papilloma
Hagiwara et al. (2013)
Rat, F344/DuCrlCrlj
oral—water
male (30/group): 0,100, 300, 500, or
1,000 mg/kg-d
daily for 31 wk beginning 1 wk after a 4-wk
exposure to W-butyl-W-(4-hydroxybutyl) (BBN)
Male
Dose (mg/kg-d)
Response (incidence)
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, or
1,000 mg/kg-d
daily for 31 wk beginning 1 wk after a 4-wk
exposure to W-butyl-W-(4-hydroxybutyl) (BBN)
Male
Dose (mg/kg-d)
Response (incidence)
0
24/30
100
18/30
300
20/30
500
25/30
1,000
21/26
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Table 1-19. Supplemental evidence pertaining to ETBE promotion of
mutagen-initiated tumors in animals (continued)
Reference and dosing protocol
Results by endpoint
Urinary bladder papillomatosis
Hagiwara et al. (2011); JPEC (2008d)
Rat, F344
oral—gavage
male (12/group): 0 or 1,000 mg/kg-d
daily for 23 wk following a 4-wk tumor initiation by
DMBDD3
Male
Dose (mg/kg-d)
Response (incidence)
0
0/30
300
0/30
1,000
10/30b
0C
0/12
1,000°
2/12
aDiethylnitrosamine (DEN), W-butyl-W-(4-hydroxybutyl)nitrosamine (BBN), W-methyl-W-nitrosourea (MNU),
1,2-dimethylhydrazine dihydrochloride (DMH), and W-bis(2-hydroxypropyl)nitrosamine (DHPN).
Statistically significant (p < 0.05) based on analysis of data conducted by the study authors.
cNo DMBDD initiation.
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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 (C)
Female rat uterine malignancies; 104wks (C)
Male rat all tissues;104wks (D)
Female rat all tissucs;104wks (D)
Male rat all tissues;23wks without DMBDD initiation (A)
~ E]
~ ~
a ~
-B ~
Male rat colon;23wks following 4wk initiation with DMBDD
(A)
Male rat forestomach or hyperplasia;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 papi)loma;31wks following 4wk
initiation with BBN (B)
Male rat urinary bladder papillamatosis;23wks following
4wk initiation with DMBDD (A)
C3 B—¦ E]
C3 B—B Q
10 100 1,000
Dose (mg/kg-day)
10,000
Sources: (A) Hagiwara etal., 2011; JPEC 2008d [B) Hagiwara et al., 2013 [C] Malarkey and Bucher, 2011
preanalysis 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] J
100
Source: [A] Saito et al, 2013; JPEC, 2010b
1,000 10,000 100,000
Exposure Concentration (mg/m3)
Figure 1-17. Exposure-response array of carcinogenic effects following
inhalation exposure to ETBE.
1.2.6. Other Toxicological Effects
The evidence base for other effects includes 11 rodent studies, some of which reported
decreased body weight, increased adrenal weights, altered spleen weights, and increased mortality
(see Supplemental Information, Appendix B). The available subchronic or chronic studies used
inhalation, gavage, or drinking water routes of exposure for 90 days or more. Shorter duration,
multiple-exposure studies that examined immunological endpoints were also 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.
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1.3. INTEGRATION AND EVALUATION
1.3.1. Effects Other Than Cancer
Kidney effects were identified as a potential human hazard of ETBE exposure based on
several affected endpoints in male and female rats, including kidney-weight increases, urothelial
hyperplasia (in male rats only), and—to a lesser extent—exacerbated severity of 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 the ETBE metabolite tert-butanol (e.g., CPN and
transitional epithelial hyperplasia) and a related compound, MTBE [e.g., CPN and mineralization;
ATSDR fl9961]. Changes in kidney parameters were consistently observed, but the magnitude of
change was generally moderate, although males had greater severity of effects than females. While
the ETBE metabolite tert-butanol meets the criteria for alpha 2u-globulin nephropathy
(https: //cfpub.epa.gov/ncea/iris drafts/recordisplay.cfm?deid=2 62086). ETBE, in contrast, binds
to alpha 2u-globulin and meets some but not all the criteria in the EPA and IARC alpha 2u-globulin
frameworks (Capen etal.. 1999: U.S. EPA. 1991a): see Section 1.2.1. U.S. EPA (1991a) noted that
"[i]f a compound induces a2u-globulin accumulation in hyaline droplets, the associated
nephropathy in male rats is not an appropriate endpointto determine noncancer (systemic) effects
potentially occurring in humans" (Section XVIII, page 89). However, because alpha 2u-globulin
nephropathy is strictly a male rat phenomenon, dose-related kidney effects in female rats are not
confounded by alpha 2u-globulin nephropathy.
It has been observed that chemicals that bind to alpha 2u-globulin also exacerbate the
incidence and/or severity of background CPN in male rats (Frazier etal.. 2012: Travlos etal.. 2011:
U.S. EPA. 1991a 1. CPN has no known analogue in the aging human kidney fNIEHS. 2019: Hard etal..
20091 and its etiology is unknown fNIEHS. 2019: Frazier etal.. 2012: Hard and Khan. 2004: Peter et
al.. 1986). However, many of the same lesions observed in CPN (e.g., thickening of tubule basement
membranes, tubule atrophy, tubule dilation, and glomerular sclerosis) are also observed in chronic
kidney disease in humans (Lusco etal.. 2016: Zoia etal.. 2015: Frazier etal.. 2012: Abrass. 2000).
Because the mode of action is unknown, it cannot be ruled out that a chemical that exacerbates CPN
in rats could also exacerbate existing disease processes in the human kidney (NIEHS. 2019).
Therefore, increased incidence of kidney effects with ETBE exposure, particularly in the female rat
(including increased kidney weight and increased severity of CPN), are considered relevant to
humans and were carried forward for dose-response analysis. Kidney effects in male rats were also
modeled and presented for comparison.
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
both sexes after 13-week inhalation and 26-week oral exposures. No other histopathological
findings were observed, and only one serum marker of liver toxicity (GGT) was elevated, although
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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. Because of the lack of data supporting the adversity of the
liver-weight increases with ETBE exposure (e.g., alterations in histology or clinical chemistry), liver
effects were not 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, female reproductive, developmental, or other toxic effects as human hazards of ETBE
exposure. While the ETBE database includes developmental, reproductive and multigenerational
studies, which are generally null and do not appear to indicate an area of concern (see hazard
discussions in Sections 1.2.3 and 1.2.4), this body of evidence is not sufficiently robust to conclude
that ETBE is not likely to be a reproductive or developmental hazard.
1.3.2. Carcinogenicity
Summary of Evidence
In F344 rats, administration of ETBE via inhalation increased hepatocellular adenomas in
males in an exposure-dependent manner, as indicated by a significant positive trend (p <0.001
with Peto's test). Hepatocellular tumors were not increased in female rats (Saito etal.. 2013). A
significantly increased incidence of hepatocellular adenomas or carcinomas (only one carcinoma
observed) was observed at the highest dose tested in males, and three hepatocellular adenomas
were observed at the two lower concentrations. Significant increases in preneoplastic foci
(basophilic and eosinophilic foci) were also observed in male rats f Saito etal.. 20131. Following
2-year gavage or drinking water exposure, the incidence of liver tumors was not increased in S-D or
F344 rats of either sex (Suzuki etal.. 2012: Maltoni etal.. 1999). although an apparent, but
statistically nonsignificant increase in preneoplastic foci (basophilic) was observed in F344 male
rats f Suzuki etal.. 20121. Regarding the 2-year gavage study by Maltoni et al. f!9991. depressed
survival (25-28% of male and female control rats survived to Week 104) may have confounded the
ability to detect carcinogenicity.
Saito etal. (2013) compared the doses achieved in the 2-year drinking water bioassay with
those for the 2-year inhalation bioassay (Suzuki etal.. 2012: TPEC. 2010a) and calculated the highest
intake in male rats to be 542 mg/kg-day as compared to the highest inhalation intake estimate of
3,015 mg/kg-day (after adjusting for 6 hours/day, 5 days a week exposure, minute volume of
561 mL/min, and 100% ETBE absorption). However, the ranges of the internal dose metrics of
ETBE expressed as either metabolized ETBE or metabolized tert-butanol (one of the two primary
breakdown products of ETBE) as computed by toxicokinetic modeling were similar for the oral and
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inhalation exposures of ETBE and the oral exposure of tert-butanol in these three studies (see the
range of the internal dose variables in Figure B-3 of the Supplemental Information). Nonetheless,
the incidence of liver tumors was not consistently correlated with any internal dose measure, and
the lack of a consistent dose-response relationship using any of these internal metrics suggests that
differences in liver tumor responses between oral and inhalation exposures are not likely due to
pharmacokinetic factors alone, and additional factors besides internal dose are necessary to explain
the induction of liver tumors. Statistically significant increases in liver tumor incidence were
observed in the livers of male F344 and Wistar rats in supplemental initiation-promotion studies
after 19-23 weeks of ETBE exposure via gavage, butonly following an initial 2-4 week mutagen
exposure fHagiwara etal.. 2015: Hagiwara et al.. 20111 which also produced increases in colon,
thyroid, forestomach, and urinary bladder tumorigenesis in male F344 rats f Hagiwara etal.. 2013:
Hagiwara etal.. 20111. No studies have evaluated chronic ETBE exposure in mice via any route.
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. However, as
discussed in Section 1.2.2, an MOA for liver carcinogenesis could not be established, and in the
absence of information to indicate otherwise fU.S. EPA. 2005bl. the liver tumors induced by ETBE
following inhalation exposure are considered relevant to human hazard identification.
The cancer guidelines indicate that information on metabolites can help inform the weight
of evidence for carcinogenicity. ETBE is primarily metabolized into acetaldehyde and tert-butanol.
Regarding the ETBE metabolite tert-butanol, drinking water exposure in F344 rats did not cause an
increase in liver tumors, but 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 in males fNTP. 19951. Regarding the ETBE metabolite
acetaldehyde, flARC. 1999bl has concluded that inhalation exposure causes carcinomas of the
nasal mucosa in rats and carcinomas of the larynx in hamsters. IARC classifies acetaldehyde as
possibly carcinogenic to humans (Group 2B) based on sufficient evidence in experimental animals.
Acetaldehyde produced in the liver as a result of ethanol ingestion and metabolism has been
suggested to be a contributor to liver toxicity and cancer (Setshedi etal.. 2010). IARC has classified
acetaldehyde associated with the consumption of alcoholic beverages to be carcinogenic to humans
[Group 1; IARC f!999al] and has concluded that it is the key alcohol metabolite in causing cancer of
the esophagus and aerodigestive tract flARC. 20101.
Integration of evidence
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; it 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
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species. The results for ETBE raise a concern for cancer; however, the effects were limited to
tumors in one tissue (liver), primarily at the high dose, by one route of exposure (inhalation) in
male rats (females were negative and no ETBE bioassays are available in mice).
Regarding the animal database for carcinogenicity, EPA considers chronic bioassays as key
evidence (i.e., the three chronic cancer bioassays, one inhalation and two oral), and other types of
studies, including initiation promotion studies, as supplemental lines of information which can aid
in the interpretation of more standard toxicological evidence, especially regarding potential modes
of action (U.S. EPA. 2005a). Across three initiation-promotion studies, orally administered ETBE
enhanced tumorigenesis in multiple tissues in male rats preexposed to mutagens, including kidney,
liver, forestomach, thyroid, colon, and urinary bladder. Furthermore, no MOA has been identified
for ETBE that could explain the potentiation of mutagen-induced carcinogenesis in the forestomach,
thyroid, colon, and urinary bladder. This suggests that the available database is insufficient to
determine the 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 of ETBE
2-year rodent oral bioassays alone.
The carcinogenicity of ETBE appears to be route dependent, therefore multiple cancer
descriptors are used in accordance with the cancer guidelines fU.S. EPA. 2005al. The evidence of
carcinogenic potential for ETBE is determined to be suggestive for exposure via the inhalation route
and inadequate for exposure via the oral route. This weight-of-evidence descriptor is based
primarily on a positive carcinogenic response following inhalation exposure in the liver in a single
animal study, along with significant increases in preneoplastic liver lesions and mechanistic data
(i.e., the metabolism of ETBE to the genotoxic compound acetaldehyde in the liver) and no
increased liver tumors detected in two chronic oral studies in S-D or F344 rats fSuzuki etal.. 2012:
TPEC. 2010a: Maltoni etal.. 19991.
Biological considerations for dose-response analysis
This section addresses the cancer hazards to bring forward to Section 2 for dose-response
analysis. The observed liver tumors in male rats following inhalation exposure are deemed
relevant for estimating human cancer hazard. The Saito etal. f20131 inhalation study observed a
positive exposure-response trend in the incidence of hepatocellular tumors in male rats (although
the majority of tumors were observed at the highest dose). This study was considered suitable for
dose-response analysis, because it is part of a well-designed GLP study 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. Although decreased body-weight gain and survival were noted in the
high-dose males and females, the study authors did not detect changes to the animals' general
condition (e.g., abnormal behavior or clinical signs) associated with ETBE. Similar decreases in
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body weight were observed in male (75% of control) and female animals (78% of control),
although significantly increased liver tumors were only observed in male rats. Some concern exists
that decreased survival in ETBE treated dose groups (primarily from CPN in the male rats) could
bias the cancer findings toward the null. This study was deemed appropriate for dose-response as
there is no clear indication that overt toxicity or altered toxicokinetics (as discussed in the Cancer
Guidelines) are responsible for the significantly increased incidence of liver tumors in male rats.
The results from MOA analysis can potentially inform dose-response analysis and
extrapolation approaches (U.S. EPA. 2005a). but no clear MOA was identified for ETBE. 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, which are 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 (U.S. EPA. 2005a).
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 (Brennan. 2002). 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 Aldh2 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 (Wengetal.. 2019: Weng etal..
2014: Wengetal.. 2013: Wengetal.. 2012: Weng etal.. 2011). Studies in Aldh2 KO mice observed
elevated blood concentrations of acetaldehyde following ETBE exposure compared with wild-type
mice (Weng etal.. 2013). increased alterations to sperm and male reproductive tissue (Weng etal..
20141. and increased incidence of centrilobular hypertrophy fWeng etal.. 2013: Weng etal.. 20121.
Similar effects on genotoxicity and liver histopathology have also been observed in AIdh2
heterozygous animals (Wengetal.. 2019). 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 (Saito etal.. 2013: TPEC. 2010b).
No MOA information exists to account for the sex discrepancies in genotoxic effects. Finally,
IARC f!999al and IARC f20121 identified acetaldehyde produced as a result of ethanol metabolism
as contributing to human carcinogenesis in the upper aerodigestive tract and esophagus following
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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. It is also plausible that
individuals with noncoding region variants in aldh2 (which could potentially affect gene
expression), as well as individuals with other variants in alcohol metabolism may also be
disproportionately affected by ETBE exposure. 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.
Regarding lifestages particularly susceptible to ETBE exposure, while certain lifestages
(e.g., development) are generally thought to have heightened vulnerability to most chemical
exposures, no specific data was identified to support heightened periods of susceptibility to ETBE
exposure in the available database of developmental and reproductive studies. Note that the
majority of developmental and reproductive studies were performed by the oral route, so
differential susceptibility from inhalation exposure cannot be ruled out, but in general, the effects
observed from ETBE exposure (except for liver tumors) appeared to be consistent across routes of
exposure (e.g., kidney effects, liver hypertrophy).
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2. DOSE-RESPONSE ANALYSIS
2.1. ORAL REFERENCE DOSE FOR EFFECTS OTHER THAN CANCER
The oral 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 (UFs) 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 f20021 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 when the reported effects are determined to be associated with exposure. No
human occupational or epidemiological studies of oral exposure to ethyl tertiary butyl ether
(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. 2002). 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 in males), and altered
serum biomarkers (cholesterol, creatinine, blood urea nitrogen [BUN]) in rats. The strongest and
most consistent findings across oral exposure routes and durations were for absolute
kidney-weight changes and urothelial hyperplasia in male rats; thus, only these endpoints were
analyzed for dose-response. Kidney effects, such as urothelial hyperplasia, observed after chronic
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exposure 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: Hagiwara et al.. 2011: Fuiii etal.. 2010:
TPEC. 2010b. 2008b. c; Gaoua. 2004b: Medinskv et al.. 19991. 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 et al. (20111 used only one dose group, so 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 in male rats were not considered suitable because of the
prevalence of age-associated confounders (e.g., chronic progressive nephropathy [CPN]-related
early deaths in male rats). Therefore, the urothelial hyperplasia data (observed in male rats only)
and absolute kidney weight (female rats, only) were the only endpoints from the 2-year studies
[(TPEC. 2010a): with selected data published as Suzuki etal. (2012)]. Absolute kidney weight was
the only endpoint from the 13- to 26-week studies considered for dose-response analysis. These
data and the absolute kidney weights from the remaining studies, TPEC f2008cl [with selected data
published as Mivataetal. f20131], Gaoua f2004bl. and Fuiii etal. f20101. are discussed further
below.
In the 2-year drinking water study (Suzuki etal.. 2012: TPEC. 2010a). 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 reached statistical significance
at 121 and 542 mg/kg-day.
In the TPEC f2008cl 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 (S-D)
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 (Postnatal Day [PND] 21) for 18 weeks. Absolute kidney
weight were significantly increased in all dose groups in F0 males (but not in F0 females) and 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, so F0 female kidney weights were not modeled. Absolute kidney weights were also
increased in F1 animals, but the F1 animals appeared to be less affected than the F0 animals.
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In the Fujii etal. (2010) 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.
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 the U.S. Environmental
Protection Agency's (EPA's) Benchmark Dose Software (BMDS) were applied. Consistent with
EPA's Benchmark Dose Technical Guidance Document (U.S. EPA. 20121. the benchmark dose (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 a BMR of 10% extra risk on incidences of urothelial
hyperplasia data were 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); which 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 (U.S. EPA. 2011). 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 3/4 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
human equivalent doses (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.
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Table 2-1. Summary of derivation of points of departure following oral
exposure for up to 2 years
Endpoint and reference
Species/
sex
Model3
BMR
BMD
mg/kg-d
BMDL
mg/kg-d
PODADJb
mg/kg-d
PODHEdc
mg/kg-d
Kidney
Increased urothelial hyperplasia;
2-yr
Suzuki et al. (2012); JPEC (2010a)
F344 rat/M
Quantal-
Linear
10% ER
79.3
60.5
60.5
14.5
Increased absolute kidney
weight; 2-yr
Suzuki et al. (2012); JPEC (2010a)
F344 rat/M
NOAELde 121 mg/kg-d
5% 1" in kidney weight
121
29.0
Increased absolute kidney
weight; 2-yr
Suzuki et al. (2012); JPEC (2010a)
F344 rat/F
Exponential
(M4)
10%
RD
204
120
120
28.8
Increased absolute kidney
weight; 26-wk
JPEC (2008c); Mivata et al. (2013)
S-D rat/M
Linear
10%
RD
176
115
115
27.6
Increased absolute kidney
weight; 26-wk
JPEC (2008c); Mivata et al. (2013)
S-D rat/F
Exponential
(M4)
10%
RD
224
57
57
13.7
Increased absolute kidney weight
(F0 generation); 18-wk
Gaoua(2004b)
S-D rat/M
Hill
10%
RD
244
94
94
22.6
Increased absolute kidney weight
(F0 generation); 16-wk
Fuiii et al. (2010)
S-D rat/M
Hill
10%
RD
435
139
139
33.4
Increased absolute kidney weight
(F0 generation); 17-wk
Fuiii et al. (2010)
S-D rat/F
Polynomial
(2 degree)
10%
RD
1,094
905
905
217
ER = extra risk, F = female; M = male; RD = relative deviation; TWA = time-weighted average.
aFor modeling details, see Appendix C of the Supplemental Information.
bFor studies in which animals were not dosed daily, administered doses were adjusted to calculate the TWA daily
doses prior to BMD modeling. This adjustment, however, was not required for the studies evaluated.
CHED PODs were calculated using BW3/4 scaling (U.S. EPA, 2011).
dNOAEL was used because of the lack of suitable model fit (see Appendix C).
e18% increase in kidney weight at LOAEL.
Consistent with EPA guidance (U.S. EPA. 2011). the PODs estimated based on effects in adult
animals, are converted to HEDs using a standard dosimetric adjustment factor (DAF) derived as
follows:
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DAF
(BWa1/4 -H BWhl/4)
(2-1)
where
BWa
BWh
animal body weight
human body weight
Using a standard BWa of 0.25 kg for rats and a BWh of 70 kg for humans fU.S. EPA. 19881.
the resulting DAF for rats is 0.24. Applying the DAF to the POD identified for effects in adult rats
yields a PODhed as follows (see Table 2-1):
Table 2-1 summarizes the sequence of calculations leading to the derivation of a
human-equivalent POD for each data set discussed above.
2.1.3. Derivation of Candidate Values
Consistent with EPA's A Review of the Reference Dose and Reference Concentration Processes
(U.S. EPA. 2002: Section 4.4.5). five possible areas of uncertainty and variability were considered
when determining the application of UF values to the PODs presented in Table 2-1. 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 oral exposure to ETBE fU.S. EPA. 20021.
An interspecies uncertainty factor, UFa, of 3 (10°5 = 3.16, rounded to 3) was applied to PODs
that used BW3/4 scaling to extrapolate oral doses from laboratory animals to humans. Although
BW3/4 scaling addresses some aspects of cross-species extrapolation of toxicokinetic and
toxicodynamic processes, some residual uncertainty remains. In the absence of chemical-specific
data to quantify this uncertainty, EPA's BW3/4 guidance (U.S. EPA. 2011) recommends using an
uncertainty factor of 3. For PODs that did not use BW3/4 such as early-life effects, a UFa of 10 was
applied CIJ.S. EPA. 20111.
A subchronic-to-chronic uncertainty factor, UFs, differs depending on the exposure
duration. For studies of 16- to 2 6-week duration, the magnitude of change observed in kidney
weights was similar to the effect observed at 104 weeks. This suggests a maximum effect could
have been reached by 16-26 weeks. The 104-week kidney data, however, are confounded due to
age-associated factors, so this comparison might not be completely reliable. Additionally, some but
not all markers of kidney toxicity appear more severely affected by ETBE at 2 years compared with
observations at 16-26 weeks [e.g., histopathology, BUN; Suzuki etal. f20121: TPEC f2010al]. Thus,
a UFs of 3 was applied for studies of 16- to 26-week duration to account for this uncertainty, and a
UFs of 1 was applied to 2-year studies.
PODhed
Duration-adjusted laboratory animal dose (mg/kg-day) x DAF (2-2)
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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 in absolute kidney weight and a
10% extra risk of urothelial hyperplasia were selected assuming that they represent minimal
biologically significant response levels.
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 fMivataetal.. 20131. prenatal developmental toxicity studies in rats and rabbits (Aso
etal.. 2014: Asano etal.. 20111. and both single- and multigeneration reproductive studies and
developmental studies in rats fFuiii etal.. 2010: Gaoua. 2004a. b). As discussed in Appendix B of
the Supplemental Information, the ETBE data set does not indicate immunotoxicity fBanton etal..
2011: Li etal.. 20111. Additionally, the available mouse study showed 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 the available evidence does not raise concern that additional studies
would likely lead to identification of a more sensitive endpoint or a lower POD. Furthermore, the
effects observed in inhalation studies support the noncancer effects observed in the oral studies.
Therefore, an uncertainty factor for the database, UFd, of 1 was applied.
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.
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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-yr
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; 2-yr
Suzuki et al. (2012); JPEC (2010a)
29.0
NOAEL
3
10
1
1
1
30
lx 10°
Increased absolute kidney
weight; female rat; 2-yr
Suzuki et al. (2012); JPEC (2010a)
28.8
BMDLio
3
10
1
1
1
30
lx 10°
Increased absolute kidney
weight; male rat; 26-wk
JPEC (2008c); Miyata et al. (2013)
27.6
BMDLio
3
10
1
3
1
100
3 x 10"1
Increased absolute kidney
weight; female rat; 26-wk
JPEC (2008c); Mivata et al. (2013)
13.7
BMDLio
3
10
1
3
1
100
1 x 10"1
Increased absolute kidney
weight; F0 male rat; 18-wk
Gaoua(2004b)
22.6
BMDLio
3
10
1
3
1
100
2 x 10"1
Increased absolute kidney
weight; male rat; 16-wk
Fuiii et al. (2010)
33.4
BMDLio
3
10
1
3
1
100
3 x 10"1
Increased absolute kidney
weight; female rat; 17-wk
Fuiii et al. (2010)
217
BMDLio
3
10
1
3
1
100
2 x 10°
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Increased urothelial hyperplasia; male
rat; 2 year
Suzuki et al. [2012); JPEC (2010a]
Increased absolute kidney weight;
male rat; 2 year
Suzuki et al. (2012); JPEC (2010a)
Increased absolute kidney weight;
female rat; 2 year
Suzuki et al. (2012); JPEC (2010a)
Increased absolute kidney weight;
male rat; 26 week
JPEC (2008c)
Increased absolute kidney weight;
female rat; 26 week
JPEC (2008c)
Increased absolute kidney weight;
male rat; 18 week
Gaoua (2004b)
Increased absolute kidney weight;male
rat; 16 week
Fujii (2010)
Increased absolute kidney weight;
female rat; 17 week
Fujii (2010)
0.1
1 10
mg/kg-day
100
1000
~ Candidate RfD
© PODhed
Composite IIF
Figure 2-1. Oral candidate values with corresponding POD and composite UF.
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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.
Table 2-3. Organ/system-specific RfDs and overall RfD for ETBE
Effect
Basis
RfD (mg/kg-d)
Study exposure
description
Confidence
Kidney
Increased absolute kidney
weight in female rats; Suzuki
etal. (2012); JPEC (2010a)
lx 10°
Chronic
High
Overall RfD
Kidney
1 X 10°
Chronic
High
Kidney Toxicity
For ETBE, candidate values were derived for increases in urothelial hyperplasia or absolute
kidney weight in male or female rats. 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
values based on data from a 2-year oral study are those for increased absolute kidney weight in
female rats and urothelial hyperplasia in male rats (Saito etal.. 2013: Suzuki etal.. 2012: TPEC.
2010a. b). While urothelial hyperplasia in male rats was the most sensitive candidate value, this
endpoint was not observed in female rats or mice of either sex, whereas increased kidney weights
were observed in multiple studies in rats of both sexes and in mice. A lower candidate value based
on increased kidney weights in female rats was derived from a smaller [n = 15), shorter duration
study [26 weeks; TPEC f2008cl]: however, based on comparison of BMD values (204 vs. 224
mg/kg-day) and composite UFs (30 vs. 100), this lower value reflects greater variability and
uncertainty in the data. The composite UF for the candidate value based on increased kidney
weight in female rats treated for 2 years was the lowest uncertainty value, providing greater
certainty in the selection of this candidate value. In addition, the candidate values from female rats
are not potentially confounded by alpha 2u-globulin related processes.
Collectively, these observations indicate that the most appropriate basis for a
kidney-specific RfD would be the increased absolute kidney weight in female rats from the 2-year
oral study (Suzuki etal.. 2012: TPEC. 2010a). The observed increases in kidney weight are likely a
result of increased CPN severity seen with ETBE exposure as CPN is characterized by cell
proliferation and chronic inflammation that results in increased kidney weight (Melnick etal..
2012: Travlos etal.. 2011).
The candidate value for increased absolute kidney weight in female rats
(1x10° mg/kg-day) was selected as the kidney-specific reference dose for ETBE. Confidence in
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this RfD is high. The candidate value is derived from a well-conducted good laboratory practice
(GLP) study, involving a sufficient number of animals per group and assessing a wide range of
kidney endpoints. In addition, the POD is based on benchmark dose modeling, with the POD within
the range of tested doses (e.g., not requiring extrapolation well beyond the experimental range),
and the reference value is associated with less relative uncertainty (as illustrated by the smaller
composite UF). Furthermore, several additional studies demonstrate quantitatively similar PODs
and candidate values for kidney effects (see Table 2-2 and Figure 2-1).
2.1.5. Selection of the Overall Reference Dose
For ETBE, kidney effects were identified as the strongest hazard and carried forward for
dose-response analysis; thus, only one organ/system-specific reference dose was derived.
Therefore, the kidney-specific RfD of 1 x 10° mg/kg-day is the overall RfD for ETBE. This value is
based on increased absolute kidney weight in female 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 (Suzuki 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 the existing evidence base does not raise concerns that
additional studies would be likely to lead to the identification of a more sensitive endpoint.
Furthermore, multiple studies demonstrate quantitatively similar PODs and candidate values for
kidney effects (see Table 2-2 and Figure 2-1), providing additional support 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 reference concentration (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 of ETBE 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 there are quantitative measures of exposure and reported effects that are determined to be
associated with that exposure. Data on the effects of inhaled ETBE in humans is limited to a small
number of 2-hour inhalation studies at concentrations up to 208.9 mg/m3 fNihlen et al.. 1998b:
Vetrano. 19931. These studies were not considered for dose-response assessment because they are
of acute duration and primarily focused on toxicokinetics.
The database for ETBE includes inhalation studies and data sets from animal studies 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
There is evidence supporting kidney effects following ETBE exposure in rats, including
organ-weight changes, histopathology (urothelial hyperplasia and exacerbation of CPN), and
altered clinical chemistry (creatinine, BUN, cholesterol). The most consistent,
concentration-related findings across multiple studies, in both sexes, were for kidney-weight
changes and increased CPN severity. In the case of kidney-weight changes, one chronic and
numerous subchronic studies investigated this endpoint following inhalation exposure fSuzuki et
al.. 2012: Hagiwar a et al.. 2011: Fuiii etal.. 2010: TPEC. 2010b. 2008b. c; Gaoua. 2004b: Medinskv et
al.. 1999). A 2-year inhalation study (Saito etal.. 2013: TPEC. 2010b) reported increased urothelial
hyperplasia in male rats only, and increased kidney weight and CPN severity in both sexes.
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Increased kidney weights from the 13-week studies were also considered for dose-response
analysis fSaito etal.. 2013: IPEC. 2010b).
In the Saito etal. T20131 2-year inhalation study, male and female F344 rats
(50/sex/concentration group) were exposed to concentrations of 0, 2,090, 6,270, or 20,900 mg/m3
fTPEC. 2010bl. Increased incidence of urothelial hyperplasia was only observed in males and
significantly increased at 6,270 and 20,900 mg/m3. Increased severity of CPN was significantly
increased in males and females at 20,900 mg/m3. Absolute kidney weight was significantly
increased in males at concentrations >2,090 mg/m3 and in females at >6,270 mg/m3.
In the IPEC 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. 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. (19991 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 concentrations in both males 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 (U.S. EPA. 2012).
the benchmark concentration (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 (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 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.
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Table 2-4. Summary of derivation of PODs following inhalation exposure
Endpoint and reference
Species/
sex
Model3
BMR
BMC
(mg/m3)
BMCL
(mg/m3)
PODADJb
(mg/m3)
PODHEcc
(mg/m3)
Kidney
Increased urothelial
hyperplasia; 2-yr
Saito etal. (2013); JPEC
(2010b)
F344 rat/M
Gamma
10%
2,734
1,498
268
265
Increased CPN severity; 2-yr
Saito etal. (2013); JPEC
(2010b)
F344 rat/M and
F
NOAEL:d 6,270 mg/m3
1,120
1,110
Increased absolute kidney
weight; 2-yr
Saito etal. (2013); JPEC
(2010b)
F344 rat/F
NOAEL:de 6,270 mg/m3
6% 1" in kidney weight
1,120
1,110
Increased absolute kidney
weight; 13-wk
JPEC (2008b)
S-D rat/M
NOAEL:d 627 mg/m3
10% 1" in kidney weight
112
111
Increased absolute kidney
weight; 13-wk
JPEC (2008b)
S-D rat/F
Linear
10% RD
28,591
16,628
2,969
2,946
Increased absolute kidney
weight; 13-wk
Medinskv et al. (1999)
F344 rat/M
Hill
10% RD
6,968
2,521
450
447
Increased absolute kidney
weight; 13-wk
Medinskv et al. (1999)
F344 rat/F
Exponential
(M4)
10% RD
5,610
3,411
609
604
F = female; M = male.
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 hours) 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).
dNOAELwas used because of the lack of a suitable model fit (see Appendix C).
eAbsolute kidney weight was increased 5, 6, and 18% at 2,090, 6,270, and 20,900 mg/m3. A NOAEL was selected
based on the concentration closest to a 10% change in order to more closely approximate a minimally
biologically significant change.
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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 the inhalation
studies (TPEC. 2010b. 2008b: Medinskv etal.. 19991 were adjusted to reflect a continuous exposure
by 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) (2-3)
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 extrarespiratory 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 (U.S. EPA. 1994). ETBE is a Category 3 gas
because extrarespiratory effects were observed. Therefore, the BMCLadj is multiplied by the ratio of
animal/human blood:air partition coefficients (La/Lh)- As detailed in Appendix Section 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 fNihlenetal.. 19951. This
allowed a BMCLhec to be derived as follows:
BMCLhec = BMCLadj (mg/m3) x (LA 4 LH) (interspecies conversion) (2-4)
= 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.
2.2.3. Derivation of Candidate Values
In EPA's A Review of the Reference Dose and Reference Concentration Processes (U.S. EPA.
2002: Section 4.4.51 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
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variability of response in the human population following inhalation exposure to ETBE (U.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 (U.S. EPA.
1994).
The 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, suggesting
that toxicity would be expected to increase with exposure durations greater than 13 weeks.
Therefore, a UFs of 10 was applied for studies of 13 weeks, and 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 in absolute kidney weight and a
10% extra risk of histological lesions were selected assuming that they represent minimal
biologically significant response levels. In some cases, the data were not amenable to modeling,
thus a NOAEL was selected to represent a minimal biologically significant response.
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 flPEC. 2008b: Medinskv et al.. 19991. Generally, if the RfD
or RfC is based on animal data, a factor of 3 is often applied if either a prenatal toxicity study or a
two-generation reproduction study is missing, or a factor of 10 if both are missing (U.S. EPA. 2002).
There are no developmental or multigeneration reproductive studies by the inhalation route;
however, the oral studies of prenatal developmental toxicity in rats and rabbits (Aso etal.. 2014:
Asano etal.. 20111. and single- and multigeneration reproductive toxicity and developmental
toxicity in rats fFuiii etal.. 2010: Gaoua. 2004a. b) are available to inform the inhalation database.
In addition, systemic effects 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. Similarly, the oral ETBE data set does
not indicate immunotoxicity and differences in outcome would not be expected for inhalation
exposures (Bantonetal.. 2011: Li etal.. 2011). Although most of the studies are in rats, the
available mouse study showed effects that were less severe than those in rats, suggesting that mice
are less sensitive than rats. The ETBE inhalation database, supported by the information from the
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Toxicological Review of Ethyl Tertiary Butyl Ether
oral database, adequately covers all major systemic effects, including reproductive, developmental,
immunological, and neurological effects, and the available evidence does not raise concern that
additional studies would likely lead to the 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.
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-yr
Saito et al. (2013); JPEC (2010b)
265
BMCLio
3
10
1
1
1
30
9 x 10°
Increased CPN severity; male and
female rat; 2-yr
Saito et al. (2013); JPEC (2010b)
1,110
NOAEL
3
10
1
1
1
30
4x 101
Increased absolute kidney weight;
female rat; 2-yr
Saito et al. (2013); JPEC (2010b)
1,110
NOAEL
3
10
1
1
1
30
4x 101
Increased absolute kidney weight;
male rat; 13-wk
JPEC (2008b)
111
NOAEL
3
10
1
10
1
300
4 x 10"1
Increased absolute kidney weight;
female rat; 13-wk
JPEC (2008b)
2,946
BMCLio
3
10
1
10
1
300
lx 101
Increased absolute kidney weight;
male rat; 13-wk
Medinsky et al. (1999)
447
BMCLio
3
10
1
10
1
300
2 x 10°
Increased absolute kidney weight;
female rat; 13-wk
Medinskv et al. (1999)
604
BMCLio
3
10
1
10
1
300
2 x 10°
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Increased urothelial hyperplasia;
male rat; 2 year
Saito et al. (2013); JPEC [2010b]
Increased CPN severity;
male and female rats; 2 year
Saito et al. (2013); JPEC 2010b
Increased absolute kidney weight;
female rats; 2-year Saito et al.
(2013); JPEC (2010b)
Increased absolute kidney weight;
male rat; 13 week
JPEC (2008b)
Increased absolute kidney weight;
female rat; 13 week
JPEC (2008b)
Increased absolute kidney weight;
male rat; 13 week
Medinsky et al. (1999)
Increased absolute kidney weight;
female rat; 13 week
Medinsky et al. (1999)
0.1 1 10 100 1000 10000
mg/m3
~
Candidate RfC
•
P0Dhec
¦
Composite UF
Figure 2-2. Inhalation 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
As stated previously, reference values based on lifetime exposure are preferred over
subchronic exposures to minimize uncertainty and extrapolation to a chronic exposure duration.
The only candidate reference values based on data from a 2-year inhalation study are those for
urothelial hyperplasia in male rats and increased absolute kidney weight in female rats fSaito etal..
2013: TPEC. 2010bl For ETBE, candidate values were derived for increased kidney weight in both
sexes of rats and urothelial hyperplasia in males. While the increase in urothelial hyperplasia in
male rats resulted in the lowest candidate value, this endpoint was observed only in male rats and
thus the biological relevance of this lesion is more uncertain than the kidney endpoints observed in
both sexes. Increased kidney weight was observed in both sexes of rats, however, the kidney
weight changes in female rats was selected to avoid potential confounding from alpha 2u-globulin
processes. The candidate RfC for increased absolute kidney weight in female rats (4 x 101 mg/m3)
was ultimately selected as the kidney-specific RfC for ETBE and is 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 general
indicator of kidney toxicity, which has been observed in both sexes of rats (and to a lesser extent in
mice). This RfC is identical to the candidate RfC for the increased severity of CPN in male and
female rats. That the observed increases in kidney weight seen in male and female rats is a result of
the increases in severity of CPN seen with ETBE exposure is likely, as a characteristic feature of CPN
is increasing kidney size and weight fHard etal.. 20131. Confidence in this kidney-specific RfC is
medium. The candidate value is derived from a well-conducted study, involving a sufficient number
of animals per group, including both sexes, and assessing a wide range of kidney endpoints, but the
inability to model the selected endpoint resulted in some reduction in confidence.
Table 2-6. Organ/system-specific RfCs and overall RfC for ETBE
Effect
Basis
RfC (mg/m3)
Study exposure
description
Confidence
Kidney
Increased absolute kidney weight in female
rats; Saito et al. (2013); JPEC (2010b)
4x 101
Chronic
Medium
Overall RfC
Kidney
4x 101
Chronic
Medium
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 4 x 101 mg/m3 is
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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
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 increased level of concern, even if average levels over the full exposure duration were less
than or equal to the RfC. In the case of ETBE, no specific potential for early lifestage susceptibility
to ETBE exposure was identified, as discussed in Section 1.3.3.
2.2.6. 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 (U.S. EPA.
19941. The overall confidence in this RfC is medium. Confidence in the principal study, Saito et al.
(20131: TPEC (2010bl. is medium. 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; however, the inability to model the POD resulted in some reduction
in confidence. Confidence in the database is high; the available studies evaluated a comprehensive
array of endpoints, and there is no indication that additional studies would lead to identification of
a more sensitive endpoint Reflecting medium confidence in the principal study and high
confidence in the database, overall confidence in the RfC for ETBE is medium.
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 of ETBE 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 in rats and mice.
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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 limited or no
information available regarding effects in humans or 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
database lacks an adequate model that would inform potential interspecies differences. Generally,
male rats appear more susceptible than females to ETBE toxicity. The underlying mechanistic basis
of this apparent difference, however, is not understood (except in the case of kidney effects in males
which are likely exacerbated by alpha 2u-globulin nephropathy). Most importantly, it is unknown
which animal species and sexes are more comparable to humans.
Overall, the ETBE data are insufficient to conclude that the alpha 2u-globulin process is
operative; however, key noncancer effects related to alpha 2u-globulin (including exacerbation of
CPN in male rats) were observed and considered not appropriate for hazard identification and,
therefore, not suitable for dose-response consideration. Instead, candidate values in female rats
were prioritized to avoid confounding by alpha 2u-globulin processes. Only one subchronic study
is available in mice, which are not affected by alpha 2u-globulin. This study seemed to indicate
male mice were more sensitive to kidney-weight changes than female mice, although these changes
did not reach statistical significance (Medinskv etal.. 1999: Bond etal.. 1996). Therefore, there is
uncertainty regarding whether other factors (unrelated to alpha 2u-globulin) may increase the
susceptibility of male rats to ETBE-related kidney effects.
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.
As noted in Section 1.3.2, EPA concluded that there is inadequate information to assess
carcinogenic potential for oral exposure to ETBE because the two available chronic oral bioassays
for ETBE were negative in rats f Suzuki etal.. 2012: TPEC. 2010a: Maltoni et al.. 19991 and because
no chronic oral bioassays are available in mice. Furthermore, PBPK analysis indicated the absence
of a consistent dose-response relationship for liver tumors when compared across oral and
inhalation exposures; therefore, a route-to-route extrapolation was not performed, and no oral
slope factor is derived.
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
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Toxicological Review of Ethyl Tertiary Butyl Ether
derived by applying a low-dose extrapolation procedure. If derived, the inhalation unit risk is a
plausible upper bound on the estimate of risk per |J.g/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 inhalation
exposure to 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. In each case, the rationale for the quantitative analysis is
explained, considering the uncertainty in the data and the suggestive nature of the
weight of evidence. These analyses generally would not be considered Agency
consensus estimates.
In the case of ETBE, an inhalation unit risk is derived. A description of the carcinogenicity
data is presented in the discussions of biological considerations for cancer dose-response analysis
(see Section 1.3.2). Briefly, a well-conducted study demonstrated a significant, positive
exposure-response for hepatocellular adenomas and carcinomas in male rats. While the majority of
liver tumors occurred at the high dose, statistical tests conducted by the study authors found a
significant dose-response trend using both the Peto test (incidental tumor test) and the
Cochran-Armitage test The available data do not demonstrate that the liver tumors are the result
of an excessively high dose (e.g., overt toxicity or altered toxicokinetics) in male rats rather than the
carcinogenicity of ETBE (see discussion in Section 1.2.2). Although decreased body weight gain and
survival was noted in the high dose males and females, the study authors did not detect changes to
the animals' general condition (e.g. abnormal behavior or clinical signs) associated with ETBE. In
addition, the study provided no indication that altered to toxicokinetics was responsible for the
significantly increased incidence of liver tumors in male rats. Considering these data, along with the
uncertainty associated with the suggestive weight of evidence, quantitative analysis of the tumor
data may be useful for providing a sense of the magnitude of potential carcinogenic risk
(e.g., including workers and consumers). Therefore, the hepatocellular adenomas and carcinomas
in male rats were considered for dose-response modeling and calculation of a quantitative risk
estimate. Because of the suggestive nature of the tumorigenic response (U.S. EPA. 2005a). there is
increased uncertainty in this risk estimate, and this is noted below where relevant
2.4.2. Dose-Response Analysis—Adjustments and Extrapolation Methods
The EPA Guidelines for Carcinogen Risk Assessment fU.S. EPA. 2005al recommends that the
method used to characterize and quantify a chemical's cancer risk be determined by what is known
about its mode of action (MOA) and the shape of its dose-response curve. EPA uses a two-step
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Toxicological Review of Ethyl Tertiary Butyl Ether
approach that distinguishes analysis of the observed dose-response data from inferences about
lower doses fU.S. EPA. 2005al. Within the observed range, the preferred approach is to use
biologically based modeling to incorporate a wide range of data into the analysis, 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 (Gehlhaus etal.. 2011) 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 fU.S.
EPA. 2005al. Otherwise, linear low-dose extrapolation is recommended if the MOA of
carcinogenicity is mutagenic or has not been established fU.S. EPA. 2005al. 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. A POD 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 fU.S. EPA. 19941
provides mechanisms for (1) adjusting experimental exposure concentrations to a value reflecting
continuous exposure duration and (2) determining a HEC from the animal exposure data. The
former uses an inverse concentration-time relationship to derive a health-protective duration
adjustment to time weight the intermittent exposures used in the study. The animal BMCL (see
Table 2-7) estimated from the inhalation study fSaito etal.. 2013: TPEC. 2010bl 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/m^) x (6 -h 24) x (5 4 7) (2-5)
= 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 for
rats (La) of 11.6 fKaneko etal.. 20001 and a blood:air partition coefficient for humans (Lh) of 11.7
(Nihlen et al.. 1995). This allowed a BMCLhec to be derived as follows:
BMCLhec = BMCLadj (mg/m3) x (LA 4 Lh) (interspecies conversion) (2-6)
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Toxicological Review of Ethyl Tertiary Butyl Ether
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
Table 2-7. Summary of the inhalation unit risk derivation
Tumor
Species/
sex
Selected
model
BMR
BMCadj
(mg/m3)
POD = BMCLadj
(mg/m3)
BMCLhec
(mg/m3)
Slope
factorab
(mg/m3)"1
Hepatocellular
adenomas or
carcinomas;
Saito etal. (2013); JPEC
(2010b)
F344
rat/M
Multistage
(1 degree)
10%
1,944
1,271
1,261
8 x 10"5
M = male.
aHuman equivalent slope factor = 0.1/BMCLlohec; see Appendix C of the Supplemental Information for details of
the modeling results.
bThis value is uncertain because it is based on a determination of suggestive evidence of carcinogenic potential.
The uncertainties in the data leading to this suggestive weight of evidence determination for carcinogenicity are
detailed in Sections 1.3.2, 2.4.1, and 2.4.4.
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 T able 2-7. 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 2-7.
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"5 per mg/m3, based on the liver tumor response in male
F344 rats (Saito etal.. 2013: TPEC. 2010b). This value is uncertain because it is based on a
determination of suggestive evidence of carcinogenic potential; however, the IUR may be useful for
some decision purposes such as providing a sense of the magnitude of potential risks or ranking
potential hazards (U.S. EPA. 2005a). This unit risk should not be used with continuous exposures
exceeding 1,271 mg/m3 (the POD) because above this level the dose-response relationship is
nonlinear. If risk estimates are needed for exposure corresponding to expected overall cancer risks
greater than 10%, the full dose-response model should be consulted. The slope of the linear
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Toxicological Review of Ethyl Tertiary Butyl Ether
extrapolation from the central estimate1 BMCio is (0.1 4- 0.992 x [1,944 mg/kg-day])
= 5 x 10"5 per mg/m3.
2.4.4. Uncertainties in the Derivation of the Inhalation Unit Risk
There is uncertainty 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 inhalation unit risk. Although the chronic studies did not report an increase in
liver tumorigenesis following oral exposure in rats, no other inhalation studies are available to
replicate these findings and none examined other animal models (e.g. mice). In addition, no data in
humans are available to confirm a general cancer response or the specific tumors observed in the
rat bioassay fSaito etal.. 2013: TPEC. 2010bl. An MOA for liver carcinogenicity could not be
established; therefore, in the absence of data to indicate otherwise, rat liver tumors were
considered relevant to humans (U.S. EPA. 2005a). Although changing the methods used to derive
the inhalation unit risk could change the results, standard practices were used due to the lack of a
human PBPK model, and no other data supported alternative derivation approaches.
Table 2-8. Summary of uncertainties in the derivation of the inhalation unit
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; in absence of data to indicate otherwise,
rat liver tumors were considered relevant to humans
(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) and JPEC (2010b) was a
well-conducted inhalation study which reported a
positive exposure-response trend for liver tumors in
male rats. Additional bioassays might add support to
the findings or provide results for different (possibly
lower) concentrations.
Selection of dose metric:
Alternative could T*
inhalation unit risk.
Administered
concentration was used.
Modeling based on the PBPK-derived internal dose
metric of ETBE metabolism decreased the BMCL by
35%. However, PBPK modeling was not used for the
dose-response because there was no human model.
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 unknown, this
^he Cancer Guidelines (U.S. EPA, 2005a) recommend that the Agency calculate and present the central estimate
as well as the corresponding 95% statistical bounds on the POD. The central estimate may be useful in activities
that involve formal uncertainty analysis that are required by OMB, as well as in ranking agents as to their
carcinogenic hazard.
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Consideration and impact
on cancer risk value
Decision
Justification and discussion
Alternatives could 4^ or T*
Inhalation unit risk.
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:
T* 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, as 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.
2.4.5. Previous IRIS Assessment
No previous cancer assessment for ETBE is available in IRIS.
2.5. APPLICATION OF AGE-DEPENDENT ADJUSTMENT FACTORS
As discussed in the Supplemental Guidance for Assessing Susceptibility from Early-Life
Exposure to Carcinogens fU.S. EPA. 2005bl either default or chemical-specific age-dependent
adjustment factors (ADAFs) are recommended to account for early-life exposure to carcinogens
that act through a mutagenic MOA. Because chemical-specific lifestage susceptibility data for
cancer are not available, and because the MOA for ETBE carcinogenicity is not known (see
Section 1.3.2), application of ADAFs is not recommended.
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