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&EPA
EPA/63 5/R-08/006D
www.epa.gov/iris
TOXICOLOGIC AL REVIEW
OF
ETHYLENE GLYCOL
MONOBUTYL ETHER (EGBE)
(CAS No. 111-76-2)
In Support of Summary Information on the
Integrated Risk Information System (IRIS)
November 2009
NOTICE
This document is an Final Agency/Interagency Review draft. This information is distributed
solely for the purpose of pre-dissemination peer review under applicable information quality
guidelines. It has not been formally disseminated by EPA. It does not represent and should not
be construed to represent any Agency determination or policy. It is being circulated for review
of its technical accuracy and science policy implications.
U.S. Environmental Protection Agency
Washington, DC
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1 DISCLAIMER
2
3
4 This document is a preliminary draft for review purposes only. This information is
5 distributed solely for the purpose of pre-dissemination peer review under applicable information
6 quality guidelines. It has not been formally disseminated by EPA. It does not represent and
7 should not be construed to represent any Agency determination or policy. Mention of trade
8 names or commercial products does not constitute endorsement or recommendation for use.
9
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CONTENTS—TOXICOLOGICAL REVIEW OF ETHYLENE GLYCOL MONOBUTYL
ETHER (CAS No. 111-76-2)
LIST OF TABLES vi
LIST OF FIGURES viii
LIST 01 ABBREVIATIONS AND ACRONYMS ix
FOREWORD xi
AUTHORS, CONTRIBUTORS, AND REVIEWERS xii
1. INTRODUCTION 1
2. CHEMICAL AND PHYSICAL INFORMATION 3
3. TOXICOKINETICS 4
3.1. ABSORPTION AND DISTRIBUTION 4
3 .2. METABOLISM AND ELIMINATION 5
4. HAZARD IDENTIFICATION 14
4.1. STUDIES IN HUMANS—EPIDEMIOLOGY, CASE REPORTS, CLINICAL CONTROLS.... 14
4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN ANIMALS-
ORAL AND INHALATION 17
4.2.1. Subchronic Studies 17
4.2.1.1. Oral 17
4.2.1.2. Inhalation 23
4.2.2. Chronic Studies and Cancer Bioassays 28
4.2.2.1. Inhalation 28
4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND INHALATION 33
4.4. OTHER STUDIES 37
4.4.1. Acute and Short-Term Exposure Studies 37
4.4.2. Dermal Exposure Studies 41
4.4.3. Ocular Exposure Studies 42
4.4.4. Genotoxicity 43
4.4.5. Immunotoxicity 46
4.4.6. Other In Vitro Studies 47
4.5. SYNTHESIS AND EVALUATION OF MAJOR NONCANCER EFFECTS AND MODE OF
ACTION: ORAL AND INHALATION 50
4.6. EVALUATION OF CARCINOGENICITY 55
4.6.1. Summary of Overall Weight of Evidence 55
4.6.2. Synthesis of Human, Animal, and Other Supporting Evidence 56
4.6.3. Mode-of-Action Information 57
4.6.3.1. Hypothesized MO A for Liver Tumor Development in Male Mice 57
4.6.3.2. Hypothesized MO A for Fore stomach Tumor Development in Female Mice 65
4.6.3.3. Conclusions About the Hypothesized Modes of Action 70
4.7. SUSCEPTIBLE POPULATIONS 71
4.7.1. Possible Childhood Susceptibility 73
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4.7.2. Possible Gender Differences 74
5. DOSE-RESPONSE ASSESSMENTS 76
5.1. INHALATION REFERENCE CONCENTRATION (RFC) 76
5.1.1. Choice of Principal Study and Critical Effect—with Rationale and Justification 76
5.1.2. Methods of Analysis—Including Models (PBPK, BMD, etc.) 80
5.1.2.1. BMD Approach Applied to Hemosiderin Staining Data 81
5.1.2.2. Selection of the POD 84
5.1.3. RfC Derivation—Including Application of Uncertainty Factors (UFs) 85
5.1.4. RfC Comparison Information 87
5.1.5. Previous Inhalation Assessment 88
5.2. ORAL REFERENCE DOSE (RfD) 88
5.2.1. Choice of Principal Study and Critical Effect—with Rationale and Justification 88
5.2.2. Methods of Analysis—Including Models (PBPK, BMD, etc.) 91
5.2.2.1. BMD Approach Applied to Hemosiderin Endpoint 91
5.2.2.2. Route-to-Route Extrapolation from Inhalation Data 91
5.2.2.3. Selection of the POD 91
5.2.3. RfD Derivation—Including Application of Uncertainty Factors (UFs) 92
5.2.4. RfD Comparison Information 93
5.2.5. Previous Oral Assessment 94
5.3. UNCERTAINTIES IN THE DERIVATION OF THE INHALATION REFERENCE
CONCENTRATION (RfC) AND ORAL REFERENCE DOSE (RfD) 95
5.3.1. Choice of Endpoint 96
5.3.2. Choice of Dose Metric 96
5.3.3. Use of BMC Approach 97
5.3.4. Choice of Model for BMCL Derivations 97
5.3.5. Choice of Animal to Human Extrapolation Method 97
5.3.6. Route -to-Route Extrapolation 98
5.3.7. Statistical Uncertainty at the POD 98
5.3.8. Choice ofBioassay 98
5.3.9. Choice of Species/Gender 98
5.3.10. Human Relevance ofNoncancer Responses Observed in Mice 99
5.3.11. Human Population Variability 99
5.4. CANCER ASSESSMENT 99
5.4.1. Quantification for Oral and Inhalation Cancer Risk 101
5.4.2. Uncertainties in Cancer Risk Assessment 102
5.4.2.1. Choice of Low-Dose Extrapolation Method 102
5.4.2.2. Human Relevance of Cancer Responses Observed in Mice 103
5 .5. POTENTIAL IMPACT OF SELECT UNCERTAINTIES ON THE RFC 103
6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HAZARD AND DOSE
RESPONSE 106
6.1. HUMAN HAZARD POTENTIAL 106
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6.2. DOSE RESPONSE 107
6.2.1. Noncancer—Inhalation 107
6.2.2 Noncancer—Oral 108
6.2.3. Cancer—Oral and Inhalation 109
7. REFERENCES Ill
APPENDIX A. SUMMARY OF EXTERNAL PEER REVIEW AND PUBLIC COMMENTS AND
DISPOSITION A-l
APPENDIX B. CORLEY ET AL. (2005a, 1997, 1994) PBPK MODELS B-l
APPENDIX C. RFD AND RFC DERIVATION OPTIONS C-l
C.l. RfC DERIVATIONS C-l
C. 1.1. RfC Derivations Based on Hematologic Data C-l
C. 1.1.1. NOAEL/LOAEL Method Applied to Hematologic Data C-3
C. 1.1.2. BMC Method Applied to Hematologic Data C-4
C. 1.2. BMC Method Applied to Hemosiderin Data C-10
C. 1.3. BMD Method Applied to Forestomach Lesions in Female Mice C-20
C.2. RfD DERIVATIONS C-27
C.2.1. RfD Derivations Based on Hematologic Data C-27
C.2.1.1. NOAEL/LOAEL Method and PBPK Model Applied to Hematologic Data C-27
C.2.1.2. BMD Method and PBPK Model Applied to Hematologic Data C-28
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LIST OF TABLES
2-1. Physical and chemical properties of EGBE 3
3-1. Summary of species-specific toxicokinetic parameters 7
4-1. Hematology and hemosiderin data from the 13-week drinking water exposure to EGBE
in F344 rats 20
4-2. Incidence and severity of selected histopathological changes from the 13-week drinking
water exposure to EGBE in F344 rats and mice 22
4-3. Hematology and hemosiderin data from a 14-week inhalation study of EGBE in
F344 rats 26
4-4. Hematology and hemosiderin data from a 14-week inhalation study of EGBE in
B6C3Fi mice 27
4-5. Selected female and male rat and mouse nonneoplastic effects from the 2-year chronic
EGBE inhalation study 29
4-6. Comparison of female and male rat and mouse Hct (manual) values from 3- and
12-month inhalation exposures to EGBE 31
4-7. Summary of genotoxicity studies on EGBE, BAL, and BAA 43
4-8. Incidence of liver hemangiosarcomas and hepatocellular carcinomas in studies of NTP
chemicals that caused increased hemosiderin in Kupffer cells in male mice 61
5-1. Results of candidate studies 77
5-2. Female and male rat and mouse liver hemosiderin staining incidence and RBC from
subchronic and chronic EGBE inhalation studies 78
5 -3. Summary of PBPK model s 81
5-4. AUC BAA blood concentrations measured at 12 months in both genders of
B6C3Fi mice and F344 rats 82
5-5. Comparison of BMC/BMCL values for male and female rat liver hemosiderin staining
data from inhalation chronic study using measured blood AUC (12 months) of the EGBE
metabolite BAA as a common dose metric 83
5-6. Comparison of BMC/BMCL values for male and female mouse liver hemosiderin
staining data from inhalation chronic study using measured blood AUC (12 months) of
the EGBE metabolite BAA as a common dose metric 84
5-7. Subchronic 91-day drinking water studies in rats and mice 89
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5-8. Summary of uncertainty in the EGBE noncancer and cancer risk assessments 95
5-9. Illustrative potency estimates for tumors in mice, using a linear analysis approach 102
B-l. Selected parameters used in the PBPK model for EGBE developed by
Corley et al. (2005a, 1997) B-5
C-1. Summary of PBPK model s C-2
C-2. Model estimates of BAA blood levels in female rats following inhalation exposures C-3
C-3. Comparison of BMC/BMCL values for female rat RBC count data from a 14-week
subchronic inhalation study, using modeled blood Cmax (3 months) of the EGBE
metabolite BAA as a common dose metric C-5
C-4. AUC BAA blood concentrations measured at 12 months in both genders of
B6C3Fi mice and F344 rats C-10
C-5. Comparison of BMC/BMCL values for male and female rat liver hemosiderin staining
data from inhalation chronic study using measured blood AUC (12 months) of the
EGBE metabolite BAA as a common dose metric C-12
C-6. Comparison of BMC/BMCL values for male and female mouse liver hemosiderin
staining data from inhalation chronic study using measured blood AUC (12 months)
of the EGBE metabolite BAA as a common dose metric C-13
C-7. PBPK model estimates of BAA Cmax blood levels and incidence of forestomach
epithelial hyperplasia in female mice C-21
C-8. BMDS model estimates of Cmax BMDio and BMDLio values for forestomach
epithelial hyperplasia in female mice C-22
C-9. Female mouse Cmax values for various time points of the NTP (2000) study estimated
by the Lee et al. (1998) model C-22
C-10. Estimated Cmax for BAA in blood for humans continuously exposed to varying
drinking water concentrations of EGBE C-23
C-l 1. Estimated Cmax for BAA in blood for humans continuously exposed to varying
concentrations of EGBE C-23
C-12. Modeled estimates of BAA in human blood exposed to EGBE in water C-28
C-13. Model estimates of BAA blood levels in female rats following oral exposures C-29
C-14. Comparison of female rat RBC count and MCV BMD/BMDL values from an oral
subchronic study using modeled blood Cmax (3 months) of the EGBE metabolite
BAA as a common dose metric C-30
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LIST OF FIGURES
3-1. Proposed metabolic scheme of EGBE in rats and humans 5
4-1. Simulated concentrations of EGBE, BAL, and BAA in liver tissues of female mice
exposed via inhalation for 6 hours to 250 ppm EGBE 64
5-1. PODs for selected endpoints with corresponding applied UFs and derived RfC 87
5-2. PODs for selected endpoints with corresponding applied UFs and derived RfD 94
5-3. Potential impact of select uncertainties on the RfC for EGBE 104
B-l. PBPK model of Corley et al. (1994) B-2
B-2. PBPK model of Corley et al. (2005a) B-4
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LIST OF ABBREVIATIONS AND ACRONYMS
ADH
alcohol dehydrogenase
AIC
Akaike information criterion
ALDH
aldehyde dehydrogenase
AUC
area under the curve
BAA
2-butoxyacetic acid
BAL
butoxyacetaldehyde
BMC
benchmark concentration
BMCL
benchmark concentration, 95% lower bound
BMD
benchmark dose
BMDL
benchmark dose, 95% lower bound
BMDS
benchmark dose software
BMR
benchmark response
CASRN
Chemical Abstracts Service Registry Number
CHO
Chinese hamster ovary
CHR
contact hypersensitivity response
CI
confidence interval
Cls
clearance rate
Cmax
maximum concentration
con-A
concanavalin-A
COP
cardiac output
DNA
deoxyribonucleic acid
EA
ethyl aery late
EG
ethylene glycol
EGBE
ethylene glycol monobutyl ether
EGEE
ethylene glycol ethyl ether
EGME
ethylene glycol methyl ether
GD
gestational day
GFR
glomerular filtration rate
G6PD
glucose-6-phosphate dehydrogenase
GSH
glutathione
Hb
hemoglobin
Hct
hematocrit
HEC
human equivalent concentration
HED
human equivalent dose
HH
hereditary hemochromatosis
Hp
haptoglobin
i.p.
intraperitoneal
i.v.
intravenous
IRIS
Integrated Risk Information System
KLH
keyhole limpet hemocyanin
LOAEL
lowest-observed-adverse-effect level
MAA
2-methoxyacetic acid
MCH
mean corpuscular hemoglobin
MCHC
mean corpuscular hemoglobin concentration
MCV
mean cell volume
ME
2-methoxyethanol
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MN micronuclei
MOA mode of action
NK natural killer
NOAEL no-observed-adverse-effect level
NTP National Toxicology Program
NZW New Zealand white
8-OHdG 8-hydroxydeoxyguanosine
OR osmotic resistance
OXA oxazolone
PBPK physiologically based pharmacokinetic
PFC plaque-forming cell
POD point of departure
RBC red blood cell
RfC reference concentration
RfD reference dose
ROS reactive oxygen species
s.c. subcutaneous
SCE sister chromatid exchange
SD standard deviation
t/2 half-life
TNFa tumor necrosis factor-alpha
TNP-LPS trinitrophenyl-lipopolysaccharide
UF uncertainty factor
U.S. EPA U.S. Environmental Protection Agency
Vd volume of distribution
WBC white blood cell
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FOREWORD
The purpose of this Toxicological Review is to provide scientific support and rationale for
the hazard and dose-response assessment in IRIS pertaining to chronic exposure to ethylene
glycol monobutyl ether (EGBE). It is not intended to be a comprehensive treatise on the
chemical or toxicological nature of EGBE.
The intent of Section 6, Major Conclusions in the Characterization of Hazard and Dose
Response, is to present the major conclusions reached in the derivation of the reference dose,
reference concentration and cancer assessment, where applicable, and to characterize the overall
confidence in the quantitative and qualitative aspects of hazard and dose response by addressing
the quality of data and related uncertainties. The discussion is intended to convey the limitations
of the assessment and to aid and guide the risk assessor in the ensuing steps of the risk
assessment process.
For other general information about this assessment or other questions relating to IRIS, the
reader is referred to EPA's IRIS Hotline at (202) 566-1676 (phone), (202) 566-1749 (fax), or
hotline.iris@epa.gov (email address).
November 2009
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1 AUTHORS, CONTRIBUTORS, AND REVIEWERS
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4 CHEMICAL MANAGERS/AUTHORS
5
6 Paul Reinhart, Ph.D., DABT
7 National Center for Environmental Assessment
U.S. Environmental Protection Agency
Research Triangle Park, NC
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9 Angela Howard, Ph.D.
10 National Center for Environmental Assessment
U.S. Environmental Protection Agency
Research Triangle Park, NC
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12 Jeffrey Gift, Ph.D.
13 National Center for Environmental Assessment
U.S. Environmental Protection Agency
Research Triangle Park, NC
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15 REVIEWERS
16 This document was provided for review to EPA scientists, interagency reviewers from
17 other federal agencies and White House offices, and the public, and peer reviewed by
18 independent scientists external to EPA. A summary and EPA's disposition of the comments
19 received from the independent external peer reviewers and from the public is included in
20 Appendix A.
21
22 INTERNAL EPA REVIEWERS
23
Jane Caldwell, Ph.D.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
IIa Cote, Ph.D., DABT
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Research Triangle Park, NC
Lynn Flowers, Ph.D., DABT
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Gary Foureman, Ph.D.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
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Research Triangle Park, NC
Jennifer Jinot, Ph.D.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Reeder Sams II, Ph.D.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Research Triangle Park, NC
John Vandenberg, Ph.D.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Research Triangle Park, NC
John Whalan
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
November 2009
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EXTERNAL PEER REVIEWERS
David Jollow, Ph.D.
Professor Emeritus
Medical University of South Carolina
Michael Pereira, Ph.D.
College of Medicine and Public Health
Ohio State University
Andrew Salmon, Ph.D.
Senior Toxicologist
Office of Environmental Health Hazard Assessment
California Environmental Protection Agency
Fletcher Hahn, DVM, Ph.D.
Scientist Emeritus
Lovelace Respiratory Research Institute
Rochelle Tyl, Ph.D.
Senior Fellow
Life Sciences and Toxicology
RTI International
D. Alan Warren, M.PH., Ph.D.
Academic Program Director
Environmental Health Sciences
University of South Carolina Beaufort
Gregory Travlos, DVM, DACVP
Veterinary Medical Officer
National Institute of Environmental Health Sciences
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1. INTRODUCTION
This document presents background information and justification for the Integrated Risk
Information System (IRIS) Summary of the hazard and dose-response assessment of ethylene
glycol monobutyl ether (EGBE). IRIS Summaries may include oral reference dose (RfD) and
inhalation reference concentration (RfC) values for chronic and other exposure durations, and a
carcinogenicity assessment.
The RfD and RfC, if derived, provide quantitative information for use in risk assessments
for health effects known or assumed to be produced through a nonlinear (presumed threshold)
mode of action (MOA). The 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. The inhalation RfC (expressed in units of mg/m3) is
analogous to the oral RfD, but provides a continuous inhalation exposure estimate. The
inhalation RfC considers toxic effects for both the respiratory system (portal of entry) and for
effects peripheral to the respiratory system (extrarespiratory or systemic effects). Reference
values are generally derived for chronic exposures (up to a lifetime), but may also be derived for
acute (<24 hours), short-term (>24 hours up to 30 days), and subchronic (>30 days up to 10% of
lifetime) exposure durations, all of which are derived based on an assumption of continuous
exposure throughout the duration specified. Unless specified otherwise, the RfD and RfC are
derived for chronic exposure duration.
The carcinogenicity assessment provides information on the carcinogenic hazard
potential of the substance in question and quantitative estimates of risk from oral and inhalation
exposure may be derived. The information includes a weight-of-evidence judgment of the
likelihood that the agent is a human carcinogen and the conditions under which the carcinogenic
effects may be expressed. Quantitative risk estimates may be derived from the application of a
low-dose extrapolation procedure. If derived, the oral slope factor is a plausible upper bound on
the estimate of risk per mg/kg-day of oral exposure. Similarly, a plausible inhalation unit risk is
an upper bound on the estimate of risk per (j,g/m3 air breathed.
Development of these hazard identification and dose-response assessments for EGBE has
followed the general guidelines for risk assessment as set forth by the National Research Council
(NRC, 1983). U.S. Environmental Protection Agency (U.S. EPA) Guidelines and Risk
Assessment Forum Technical Panel Reports that may have been used in the development of this
assessment include the following: Guidelines for the Health Risk Assessment of Chemical
Mixtures (U.S. EPA, 1986a), Guidelines for Mutagenicity Risk Assessment (U.S. EPA, 1986b),
Recommendations for and Documentation of Biological Values for Use in Risk Assessment (U.S.
EPA, 1988), Guidelines for Developmental Toxicity Risk Assessment {U.S. EPA, 1991), Interim
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Policy for Particle Size and Limit Concentration Issues in Inhalation Toxicity (U. S. EPA,
1994a), Methods for Derivation of Inhalation Reference Concentrations and Application of
Inhalation Dosimetry (U.S. EPA, 1994b), Use of the Benchmark Dose Approach in Health Risk
Assessment {U.S. EPA, 1995), Guidelines for Reproductive Toxicity Risk Assessment (U.S. EPA,
1996), Guidelines for Neurotoxicity Risk Assessment (U.S. EPA, 1998), Science Policy Council
Handbook. Risk Characterization (U.S. EPA, 2000a), Benchmark Dose Technical Guidance
Document (U.S. EPA, 2000b), Supplementary Guidance for Conducting Health Risk Assessment
of Chemical Mixtures (U.S. EPA, 2000c), A Review of the Reference Dose and Reference
Concentration Processes (U.S. EPA, 2002), Guidelines for Carcinogen Risk Assessment (U.S.
EPA, 2005a), Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to
Carcinogens (U.S. EPA, 2005b), Science Policy Council Handbook: Peer Review (U.S. EPA,
2006a), and A Framework for Assessing Health Risks of Environmental Exposures to Children
(U.S. EPA, 2006b).
The literature search strategy employed for this compound was based on the Chemical
Abstracts Service Registry Number (CASRN) and at least one common name. Any pertinent
scientific information submitted by the public to the IRIS Submission Desk was also considered
in the development of this document. The relevant literature was reviewed through December
2008.
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2. CHEMICAL AND PHYSICAL INFORMATION
EGBE is also known as 2-butoxyethanol. EGBE is widely used as a solvent in various
applications, such as in surface coatings, spray lacquer, quick-dry lacquers, enamels, varnishes,
varnish removers, latex paint, metal cleaners, and in commercially available cleaning products.
EGBE has been estimated to range in concentration from 1 to 30% (volume/volume) in industrial
and commercial products. The average concentration of EGBE in household products in 1977
was 2.6%. EGBE is a high production volume chemical with an estimated 390 million pounds
produced in the United States in 1992 (NTP, 2000). Some relevant physical and chemical
properties of EGBE are shown in Table 2-1.
Table 2-1. Physical and chemical properties of EGBE
CASRN
111-76-2
Empirical formula
c4h9-o-ch2ch2-oh
Molecular weight
118.2
Vapor pressure
0.88 mm Hg at 25°C (about 1,200 ppm)
Water solubility
Miscible
Log Kow
0.81
Henry's law constant
2.08 x 10"7-2.08 x 10-8 atm-m3/mole (25°C)
Flash point
62°C (closed cup); 70°C (open cup)
Conversion factor
1 ppm = 4.83 mg/m3; 1 mg/m3 = 0.207 ppm
EGBE exists as a colorless liquid at ambient temperature and pressure. Its evaporation
rate relative to butyl acetate is 0.08; thus, it is considered a "slow evaporator." It is miscible in
water and partitions about equally between phases of octanol and water. Considering the
magnitude of the octanol: water partition coefficient (~ 7 :1), it is unlikely that EGBE
bioaccumulates. Based on the magnitude of the Henry's law constant, partitioning of EGBE
between water and air greatly favors the water phase.
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3. TOXICOKINETICS
3.1. ABSORPTION AND DISTRIBUTION
EGBE is absorbed and rapidly distributed in humans following inhalation, ingestion, or
dermal exposure. Kumagai et al. (1999) examined 10 polar organic solvents, including EGBE,
during short-term inhalation by humans. Four healthy male research subjects inhaled 25 ppm
EGBE for 10 minutes; the mean uptake was 79.7% in the last 5 minutes of EGBE respiration.
Johanson and Boman (1991) attempted to define the relative importance of the skin to the
total absorption of EGBE vapors by humans by comparing mouth-only and body-only exposures.
Four research subjects were exposed to 50 ppm EGBE mouth-only for 2 hours, followed by
1 hour of no exposure, then 2 hours of 50 ppm body-only exposure (i.e., exposed in a chamber
while breathing fresh air via a respirator). Blood samples were collected periodically for
analysis of EGBE under the assumption that the finger-prick blood samples represented mixed
arterial blood. Since the areas under the curve (AUCs) for the concentration of EGBE in the
subjects' blood samples following body-only exposures were three- to fourfold greater than
following mouth-only exposure, the authors concluded that the skin accounted for approximately
75% of the total uptake of EGBE in a whole-body exposure.
Corley et al. (1994) suggested that Johanson and Boman's (1991) conclusion of greater
absorption of EGBE vapor through the skin than from the respiratory tract was inconsistent with
the physiological differences (relative surface area, blood perfusion, barrier thickness) favoring
absorption of vapors through the lungs. They reanalyzed the kinetic data of Johanson and
Boman, assuming that the finger-prick blood samples represented venous blood draining the skin
prior to mixing systemically. These revised calculations resulted in dermal uptake contributing
no more than 22% of the total uptake of EGBE in a whole-body exposure at average temperature
and humidity (skin permeability coefficient of 3 cm/hour), assuming no clothing to hinder
absorption.
To provide experimental validation of the skin's role in the uptake of EGBE vapors,
Corley et al. (1997) conducted a study in which human research subjects exposed one arm to
50 ppm [13C]-EGBE for 2 hours. Catheters installed in the antecubital vein of the unexposed
arm served as the primary site for blood collection, which was analyzed for both EGBE and
2-butoxyacetic acid (BAA). Finger-prick blood samples were collected from the exposed arm at
the end of the 2-hour exposure. If Johanson and Boman's (1991) assumption that finger-prick
blood samples represented systemic arterial blood was correct, then the concentrations of EGBE
and BAA in the finger-prick blood samples taken from the exposed arm at the end of the 2-hour
exposure should have been comparable to the corresponding catheter sample taken from the
unexposed arm. This was not the case, since the concentration of EGBE averaged nearly
1,500-fold higher in the finger-prick blood samples than in the samples collected from the
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unexposed arm, confirming the potential for portal of entry to have a major effect when using the
finger-prick sampling technique.
3.2. METABOLISM AND ELIMINATION
The metabolism of EGBE has been studied extensively in rodents, particularly in rats,
and the large body of literature on this subject has been thoroughly reviewed (Commonwealth of
Australia, 1996; ECETOC, 1994). Proposed pathways for the metabolism of EGBE in rats and
humans are presented in Figure 3-1. The principal products from metabolic processes in rats or
humans are BAA (all species) and the glutamine or glycine conjugate of BAA (humans). Other
potential metabolic products, such as the glucuronide conjugate of EGBE, ethylene glycol (EG),
butoxyacetaldehyde (BAL), and C02 are minor metabolites or are transitory in nature (e.g.,
BAL) and do not accumulate in blood, tissues, or excreta.
CO.;
carboligase oxidase dehydrogenase
CH,CH2CHjCHjOH
(Butanol)
+ HOCH/ CHyOH
(Ethylene Glycol)
CH3 CH? CH2CH2 OCH2 CH2 O-GIuc
(EGBE - Glucuronide)
de IKyl
CH3 CHaCHz CHi- OCH2 CB2 O-SQs H
(EGBE - Sulfate)
{Rats Only?)
(Rats Only?)
(Rats Only?)
CH3CH2CH2CH2 OCH2CH2OH
(EGBE)
(Rats and Human)
alcohol dehydrogenase
CH3CH2 ch2ch2 och2 CHO
(BAL)
aldehyde dehydrogenase
CHi CH? CH? CHS OCH2 CO? Glii
(BAA - Glutamine)
CHj CH?CH? CH;- OCH.»CO> -Gly
(BAA - Glycine)
(Human C
(Human Only)
CH)CHjCH2CHjOCH2COjH
(BAA)
dealkyl carboligase
Sources: Adapted from Corley et al. (1997) and Medinsky et al. (1990).
Figure 3-1. Proposed metabolic scheme of EGBE in rats and humans
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The two main oxidative pathways of EGBE metabolism observed in rats are via alcohol
dehydrogenase (ADH) and O-dealkylation by a cytochrome P450 dealkylase (CYP 2E1).
Because BAA is excreted in the urine of both rats and humans following EGBE exposure, it has
been suggested that the production of BAA through the formation of BAL by ADH would be
applicable to both rats and humans (Corley et al., 2005a, 1997; Medinsky et al., 1990).
However, the other proposed metabolic pathways of EGBE may only be applicable to rats, since
the metabolites of these pathways (i.e., EQ EGBE glucuronide, and EGBE sulfate) have been
observed in the urine of rats (Bartnik et al., 1987; Ghanayem et al., 1987a), but not in humans
(Corley et al., 1997). In addition, Corley et al. (1997) confirmed an observation of Rettenmeier
et al. (1993) that approximately two-thirds of the BAA formed by humans is conjugated with
glutamine and, to a lesser extent, glycine. The BAA-glutamine and BAA-glycine conjugation
pathways have not been detected in the rat.
Carpenter et al. (1956) first identified BAA as the metabolite responsible for the
hemolytic toxicity of EGBE by incubating BAA with whole blood from a variety of species.
Blood from rats, mice, and rabbits was more rapidly hemolyzed than blood from humans,
monkeys, dogs, or guinea pigs when incubated in vitro at 37.5°C with a saline solution of 0.1%
of the sodium salt of BAA. These results correlated well with osmotic fragility studies using
blood from these same species following in vivo inhalation exposures to EGBE. In contrast, a
much higher concentration (2.5%) of EGBE was required to produce a similar degree of
hemolysis in vitro. Subsequent investigations have shown that hemolytic blood concentrations
of BAA can be produced following oral or dermal administration or inhalation of EGBE.
The uptake and metabolism of EGBE is essentially linear following a 6-hour inhalation
exposure of up to 438 ppm, a concentration that causes mortality in animals (Sabourin et al.,
1992a). BAA is the primary metabolite in rats following drinking water (Medinsky et al., 1990)
and inhalation (Dill et al., 1998) exposures. EGBE is eliminated primarily as BAA in urine.
Lesser amounts of the glucuronide and sulfate conjugates of EGBE have been observed in the
urine of rats (Bartnik et al., 1987; Ghanayem et al., 1987a) but not humans (Corley et al., 1997).
No significant differences in the urinary levels of BAA were found following administration of
equivalent doses of EGBE dermally or in drinking water (Shyr et al., 1993; Sabourin et al.,
1992b; Medinsky et al., 1990). Corley et al. (1997) reported that the elimination kinetics of
EGBE and BAA appear to be independent of the route of exposure. Elimination of EGBE and
BAA following repeated inhalation exposure appears to be dependent on species, gender, age,
time of exposure, and exposure concentration (NTP, 2000; Dill et al., 1998). In rodents,
dose-dependent clearances of EGBE and BAA have been observed (Corley et al., 1994;
Ghanayem et al., 1990). A summary of species-specific toxicokinetic parameters is shown in
Table 3-1 followed by a brief summary of key individual studies.
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Table 3-1. Summary of species-specific toxicokinetic parameters
EGBE toxicokinetics
tin in blood (hr)
Species
Gender
Route
Mean
Reference
Human
Male
Inhalation
0.65
Johanson et al. (1986)
Human
Male
Dermal vapor
0.53-0.6
Johanson and Boman (1991)
Human
Male
Dermal vapor
0.66
Corley et al. (1997)
Human
Male/female
Inhalation
0.93
Jones and Cocker (2003)
F344 rat
Male
i.v.
0.11-0.17
Ghanayem et al. (1990)
F344 rat
Male
Inhalation
0.13-0.69
Dill et al. (1998)
F344 rat
Female
Inhalation
0.12-0.50
Dill et al. (1998)
B6C3Fi mouse
Male
Inhalation
0.05-0.16
Dill et al. (1998)
B6C3Fi mouse
Female
Inhalation
0.06-0.14
Dill et al. (1998)
B6C3FJ mouse
Female
i.p.
0.16
Poet et al. (2003)
B6C3FJ mouse
Female
Gavage
0.35
Poet et al. (2003)
Clearance (mL/min/kg body weight)
Species
Gender
Route
Mean
Reference
Human
Male
Inhalation
16.2
Johanson et al. (1986)
Guinea pig
Female
i.v.
128
Johanson et al. (1986)
F344 rat
Male
i.v.
5.9-13.3
Ghanayem et al. (1990)
Sprague-Dawley rat
Male
Inhalation
2.2-2.3
Johanson (1994)
BAA toxicokinetics
tia in blood (hr)
Species
Gender
Route
Mean
Reference
Human
Male
Inhalation
4.3
Johanson and Johnsson (1991)
Human
Male
Dermal vapor
3.27
Corley et al. (1997)
F344 rat
Male
i.v.
1.5-3.2
Ghanayem et al. (1990)
F344 rat
Male
Inhalation
0.55-1.96
Dill et al. (1998)
F344 rat
Female
Inhalation
0.79-6.6
Dill et al. (1998)
B6C3Fi mouse
Male
Inhalation
0.36-4.0
Dill et al. (1998)
B6C3Fi mouse
Female
Inhalation
0.38-4.5
Dill et al. (1998)
B6C3Fi mouse
Female
i.p.
1.05-1.42
Poet et al. (2003)
B6C3Fi mouse
Female
Gavage
1.55-2.11
Poet et al. (2003)
Clearance (mL/min/kg body weight)
Species
Gender
Route
Mean
Reference
Sprague-Dawley rat
Male
Inhalation
0.49-0.58
Johanson (1994)
i.p. = intraperitoneal; i.v. = intravenous
1
2 Percutaneous absorption of EGBE in rats is rapid and produces measured blood levels of
3 BAA sufficient to produce hemolysis (Bartnik et al., 1987). Metabolism, disposition, and
4 pharmacokinetic studies in male F344 rats conducted by Corley et al. (1994) produced hemolytic
5 blood concentrations of BAA (0.5 mM) following a single oral dose of 126 mg/kg of
6 [14C]-labeled EGBE. Using their physiologically based pharmacokinetic (PBPK) model, they
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predicted that such hemolytic blood concentrations would also be produced in rats following a
single 6-hour EGBE inhalation exposure >200 ppm. A report that evaluated the National
Toxicology Program (NTP ,2000) inhalation bioassay suggests that BAA blood concentrations in
female rats, which achieved higher blood concentrations than males, exceeded 0.5 mM
(approximately 67 jag BAA/g blood), following exposure to 62.5 ppm EGBE for both 1-day and
12-month exposure durations (Dill et al., 1998).
The metabolic basis for the hematotoxicity of EGBE was studied in male F344 rats by
using pyrazole and cyanamide as metabolic inhibitors of ADH and aldehyde dehydrogenase
(ALDH), respectively (Ghanayem et al., 1987b). Male F344 rats, 9-13 weeks old, were
pretreated with pyrazole or cyanamide followed by administration of 500 mg/kg EGBE by
gavage. The use of pyrazole protected rats from EGBE-induced hematotoxicity and resulted in a
10-fold lower ratio of BAA to conjugated EGBE excreted in urine. Cyanamide treatment
significantly reduced the hematotoxic response in a manner similar to that of pyrazole; it also
resulted in a high mortality rate in rats given cyanamide and EGBE. This effect was possibly
due to the increased levels of unmetabolized BAL; the effect was not observed in animals treated
with cyanamide or EGBE alone. Pyrazole completely blocked the increase in spleen weight to
body weight ratios seen in EGBE-treated animals. Gavage administration of either BAL or BAA
at equivalent molar doses to 125 mg/kg EGBE produced identical increased spleen to body
weight ratios and identical increases in free hemoglobin (Hb) levels in plasma. Pretreatment of
rats with cyanamide prior to administration of BAL provided significant protection against BAL-
induced hematotoxicity. These studies confirm the central role of BAA in the hematotoxic
response elicited in rats.
The effects of age, dose, and metabolic inhibitors on the toxicokinetics of EGBE were
studied in male F344 rats (Ghanayem et al., 1990). Rats aged 3-4 and 12-13 months were
administered a single gavage dose of 31.2, 62.5, or 125 mg/kg EGBE. Pretreatments included
pyrazole, cyanamide, or probenecid, an inhibitor of renal anion transport. Toxicokinetic
parameters for EGBE, including AUC, maximum plasma concentration (Cmax), and clearance
rate (Cls), were dose dependent; AUC and Cmax increased and Cls decreased as dose levels
increased. Other measured parameters were unaffected by dose. Age had no effect on half-life
(ty2), volume of distribution (Vd), or Cls of EGBE, but Cmax and AUC increased with age. As
expected from previous studies, inhibition of EGBE metabolism by either pyrazole or cyanamide
resulted in significantly increased ty2 and AUC, as well as decreased Cls. BAA toxicokinetics
were also altered by dose and age, as well as by administration of metabolic inhibitors.
Statistically significant, slight increases in Cmax, AUC, and tu were seen at higher doses; these
results were more pronounced in older rats. Probenecid pretreatment at EGBE dose levels of
31.2 and 62.5 mg/kg produced no changes in the measured toxicokinetic parameters for EGBE.
Pretreatment produced two- to threefold increases in AUC, and two- to sixfold increases in ty2 for
BAA. These results indicate that renal organic acid transport is vital to renal elimination of
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BAA. The increased Cmax, AUC, and ty2 in older versus younger rats may be due to differences in
relative contributions of the two primary metabolic pathways previously discussed, or to
compromised renal clearance.
Dill et al. (1998) described the toxicokinetics of EGBE and BAA in male and female
F344 rats and B6C3Fi mice as part of the 2-year EGBE inhalation toxicity and carcinogenicity
study conducted by the NTP (2000). Blood samples were collected from a satellite group of
animals postexposure (i.e., after the daily 6-hour exposure) after 1 day, 2 weeks, and 3,6, 12, and
18 months of exposure to target EGBE concentrations of 0, 31.2 (rats only), 62.5, 125, or
250 ppm (mice only) by whole-body inhalation; the samples were assayed for EGBE and BAA.
Postexposure time points varied from 10 to 720 minutes following 1 day, 2 weeks, and 3 and
6 months and varied from 10 to 2,880 minutes following 12 months. Postexposure 16-hour urine
samples were collected after 2 weeks and 3, 6, 12, and 18 months of exposure and assayed for
BAA. In addition, a separate set of aged mice were kept in the control chamber and exposed to
EGBE for 3 weeks when they were approximately 19 months old. Postexposure blood samples
were collected after 1 day and 3 weeks of exposure; 16-hour urine samples were collected after
2 weeks of exposure. Overall, mice eliminated both EGBE and BAA from blood faster than rats:
for example after the 1-day exposure, ty2 for rats (males and females, over three concentrations)
averaged 8.6 minutes for EGBE, while the ty2 for mice (males and females, over three
concentrations) averaged 4 minutes for EGBE. In contrast, the rate of BAA elimination from
blood decreased as the exposure concentration increased. As exposure continued, the rates of
elimination for both EGBE and BAA decreased in both species, resulting in longer residence
times in the blood. At 1 day postexposure, ty2 in male rats was 9.4 minutes, and at 18 months
postexposure, ty2 was 15.8 minutes. Female rats were significantly less efficient in clearing BAA
from their blood than males, possibly as a result of reduced renal clearance in female rats. The
aged mice were observed to eliminate BAA from blood >10 times slower than young mice after
1 day of exposure, but this difference was less obvious after 3 weeks of exposure. These
findings provide evidence that the elimination kinetics of EGBE and BAA, following repeated
inhalation exposure to EGBE, appear to be dependent on various factors, including species,
gender, age, time of exposure, and exposure concentration.
Green et al. (2002) explored reasons that female mice develop marked hyperkeratosis in
the forestomach when given oral doses (1/day for 10 days) of either EGBE or BAA. Irritation
from the carboxylic acid BAA is hypothesized to cause cell damage followed by cell
proliferation and eventually the observed hyperkeratosis. Their studies examined the activity and
localization of ADH and ALDH (the principal enzymes involved in the metabolism of EGBE in
the stomach tissues of mice and rats) the localization of these enzymes in a human stomach
sample, and whole body autoradiography of mice exposed to radiolabeled EGBE (see
Section 4.4.1). Tissue homogenates were prepared from female B6C3Fi mice (n = 30) and rats
(n = 10; gender and species not specified) and centrifuged at 41,000 x g3 with the supernatants
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used to examine the metabolism of EGBE by ADH and ALDH. The stomachs were separated
into fore and glandular sections and used to measure the metabolism of EGBE to BAL and BAL
to BAA by ADH and ALDH, respectively. A marked species difference in ALDH activity was
observed between rats and mice:
Rats: Km = 0.29 mM
Vmax = 1.627 nmol/minute per mg protein, forestomach, Km = 0.73 mM
Vmax = 2.170 nmol/minute per mg protein, glandular stomach
Mice: Km= 46.59 mM
Vmax = 17.094 nmol/minute per mg protein, forestomach, Km = 87.01 mM
Vmax = 13.986 nmol/minute per mg protein, glandular stomach
Km values were over two orders of magnitude greater in mice compared to rats. Based
upon the Km and Vmax values reported, while the mouse ALDH enzyme has a lower affinity than
the rat enzyme for EGBE, the mouse enzyme has a much greater capacity to metabolize EGBE
to the intermediate without becoming saturated. The fact that EGBE is held in the forestomach,
along with the information that rates for the ADH enzyme were of the same order of magnitude
for rats and mice, suggests that mice are capable of generating more BAA in the forestomach
than rats for the same dose and exposure duration.
Green et al. (2002) also examined the distribution of ALDH and ADH in rat, mouse, and
human stomach tissue sample from a single individual, using histochemical staining. The
stratified squamous epithelium of the forestomach of both rats and mice contained the highest
staining intensity for ALDH and ADH. These enzymes were found throughout the mucosa in the
human stomach tissue sample; the highest concentration was found in the mucus-producing cells
at the surface. Data indicates that the distribution of these enzymes in humans is more closely
comparable to that found in the rodent glandular stomach than in the rodent forestomach. This
finding, combined with the difference in ALDH and ADH activity between mouse and rat
forestomach, suggests that humans are at much lower risk for the tissue irritation seen in the
mouse forestomach.
Deisinger and Boatman (2004) determined the extent of the in vivo formation of BAL
and BAA from EGBE and their elimination kinetics from blood, liver, and forestomach of mice.
Male and female B6C3Fi mice (4/gender/time point) were administered oral doses of 600 mg/kg
EGBE dissolved in distilled water. At 5, 15, 45, and 90 minutes following the dose, blood, liver,
and forestomach tissues, along with forestomach contents, were collected and processed to
determine EGBE, BAL, and BAA concentrations in the samples. High EGBE concentrations
were measured at all time points; maximum concentrations occurred 5 minutes after dosing, with
a mean of 123 mM in females and 129 mM in males. EGBE levels in blood and liver were also
at maximum concentrations at 5 minutes postdosing, but at levels that were roughly 50-fold
lower than in the forestomach. BAA concentrations in all organs were substantial in the
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5-minute samples, and concentrations continued to increase until leveling off in the 45- and
90-minute samples. Concentrations of BAA measured in the forestomach were lower than
concentrations in blood and liver tissues. Furthermore, BAA was found to be associated with
forestomach tissues, rather than forestomach contents. BAL levels were highest in the initial
samples, 5 minutes postdose, and then declined. Levels of BAL measured in the forestomach
were 10-fold to 100-fold lower than the parent compound or carboxylic acid metabolite. No
differences between male and female mice were apparent in the parent compound or BAA organ
concentrations at comparable time points following dosing, but the BAL concentrations were up
to twofold greater at some time points in the liver and forestomach of female mice compared to
male mice.
Using rate constants derived from mouse stomach fractions (Green et al., 2002) and
making several assumptions about the use of these enzyme activity data, Corley et al. (2005a)
estimated that 250 ppm EGBE would result in peak Cmax values of 7 |iM EGBE, 0.5 [jM BAL,
and 3,250 |iM BAA in liver tissue of male mice at the end of a 6-hour inhalation exposure. The
model includes the metabolism of EGBE to BAL via ADH, and the subsequent metabolism of
BAL to BAA via ALDH in both the liver and forestomach. The model predicts that the
concentrations of BAL in gastrointestinal tract tissues of male and female mice at 5 minutes
postdosing, the time of maximal concentration, would be 18 and 33 [jM, respectively, following
gavage exposure to 600 mg/kg EGBE. This compares well with the levels of BAL actually
observed in forestomach tissue of male and female mice at 5 minutes postdosing: 19 and 33 [jM,
respectively, following gavage exposure to EGBE at 600 mg/kg (Deisinger and Boatman, 2004).
For humans, the elimination kinetics of EGBE and BAA appear to be independent of the
route of exposure. For whole-body exposures under exercise conditions, the elimination ty2 for
EGBE and BAA were 0.66 and 4 hours, respectively (Johanson and Johnsson, 1991; Johanson,
1986). For dermal exposure to neat liquids, the ty2 for elimination of EGBE and BAA were
1.3 and 3.1 hours, respectively (Johanson et al., 1988). For dermal exposure to vapors, the
elimination ty2 for EGBE was 0.53-0.6 hours.
Haufroid et al. (1997) conducted a study on 31 male workers exposed to low levels of
EGBE in a beverage package production plant. The average airborne EGBE exposure
concentration was 2.91 ± 1.30 mg/m3 (0.59 ± 0.27 ppm). Postshift urine samples showed an
average BAA concentration of 10.4 mg/g creatinine. One exposed individual who exhibited a
very low urinary BAA excretion was found to possess a genetic polymorphism for CYP 2E1 that
produced increased oxidative activity. However, the researchers did not measure BAA
conjugated to glutamine, an alternative pathway for BAA excretion in humans. Further
investigations on the influence of genetic polymorphisms for CYP 2E1 on urinary BAA
excretion rate are needed before conclusions can be drawn.
Johanson and Johnsson (1991) analyzed venous blood samples collected at 0, 2, 4, and
6 hours postexposure from five healthy, male research subjects exposed to 20 ppm EGBE via
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inhalation for 2 hours during light physical exercise on a bicycle ergometer. Blood samples were
analyzed for BAA concentrations. An average peak blood concentration of 45 [xM BAA was
reached 2-4 hours after exposure. The range of concentration was from 36 to 60 |iM, The
average ty2 for elimination of BAA from blood was 4.3 hours, with a range from 1.7 to 9.6 hours,
suggesting that blood levels of BAA would probably not increase following prolonged
occupational exposures to concentrations of EGBE vapor at or below existing occupational
exposure limits of 20-25 ppm. Thus, blood levels would not reach those shown to cause adverse
effects in vitro. The average renal clearance of BAA was 23-39 mL/minute, which was only
about one-third of the glomerular filtration rate (GFR). The authors suggested that the low
clearance of BAA relative to the GFR could have been related to the binding of BAA to proteins
in blood or to a low efficiency in renal tubular secretion. The low pKa of 3.5 estimated by the
researchers indicates that tubular reabsorption is unlikely, because more than 99% of the BAA in
normal human urine (pH ~6) is ionized. The Vd averaged 15 L (range 6.5-25 L) based on whole
blood measurements, and was approximately equal to the volume of extracellular water (13-
16 L), a further indication of binding of BAA to blood proteins.
Laitinen (1998) reported BAA levels in a study of eight silkscreen printers (gender not
specified) exposed to a mixture of EGBE and 2-butoxyethylacetate. Daily mean 8-hour air
concentrations ranged from 0.1 to 0.6 ppm during a 5-day period. Urine samples from these
workers contained 75 mg BAA/g creatinine immediately after the work shift, and 58 mg/g
creatinine the following morning, 14-16 hours postshift. Laitinen et al. (1998) reported similar
postshift urinary levels of 60 mg BAA/g creatinine in another group of 37 male and 15 female
silkscreen workers exposed to 5 ppm EGBE and its acetate for one 8-hour workday.
Several PBPK models have been developed for EGBE, each sequentially building upon
the advances from the previous model. The first model was developed by Johanson (1986) to
describe the kinetics of EGBE in the blood of human volunteers exposed for 2 hours to 20 ppm
EGBE in air while exercising. Shyr et al. (1993) published a model to describe the
pharmacokinetics of EGBE in male F344 rats based upon the drinking water exposure data of
Medinsky et al. (1990) and the inhalation and dermal data of Sabourin et al. (1992a, b). Corley
et al. (1994) then extended the Johanson (1986) model to describe the kinetics of EGBE, as well
as the major metabolite, BAA, in rats and humans and later validated the human dermal exposure
model in Corley et al. (1997). Lee et al. (1998) followed with a model that included young and
old, male and female rats and mice to describe the kinetics of EGBE and BAA in the NTP 2-year
inhalation bioassay (data in Dill et al., 1998). Based upon the data of Dill et al. (1998), there
were species, gender, age, and exposure concentration-dependent differences in the kinetics of
BAA. Lee et al. (1998) made several assumptions about the plasma protein binding of BAA, the
metabolism of EGBE to BAA and other metabolites, and the renal clearance of BAA (all initially
described by Corley et al. [1994] and Shyr et al. [1993] in male rats) to describe the kinetic data
in female rats and mice as a function of age and chronic exposure to EGBE. Corley et al.
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1 (2005a) replaced the assumptions used by Lee et al. (1998) with experimental data. This model,
2 along with the Lee et al. (1998) rat and mouse model and Corley et al. (1997) human model, is
3 used in this current review to calculate the internal dose of EGBE (Cmax of BAA in blood) used
4 in the development of the RfC and RfD. This model is described in more detail in Appendix B.
5
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4. HAZARD IDENTIFICATION
4.1. STUDIES IN HUMANS—EPIDEMIOLOGY, CASE REPORTS, CLINICAL
CONTROLS
Carpenter et al. (1956) conducted three controlled inhalation studies. In the first study, a
group of two men and six rats were exposed simultaneously for 4 hours to an EGBE
concentration of 113 ppm in a 1,250 cubic foot room. Effects observed in humans included nasal
and ocular irritation, a metallic taste, and belching. Erythrocyte osmotic fragility did not change
for the men, yet rose appreciably for the rats. In a second study, a group of two men, one
woman, and three rats were exposed to 195 ppm EGBE for two 4-hour periods, separated by a
30-minute recess, in a 6.5 cubic foot room. There was no change in the subjects' blood pressure,
erythrocyte fragility, or pulse rate. They experienced nose and throat irritation, followed by
ocular irritation and disturbed taste; one subject reported a headache. In the rats, an increase in
erythrocyte fragility values was noted. In the third study, two men and two women were exposed
for 8 hours to a 100 ppm EGBE concentration. No changes in blood pressure, erythrocyte
fragility, or pulse rate were observed. Again, nasal and throat irritation followed by ocular
irritation and a disturbing metallic taste were experienced. Two subjects reported headaches.
There are a number of case reports of acute ingestion of EGBE, consisting primarily of
accidental or intentional ingestion. Bauer et al. (1992) reported the effects of acute ingestion of
500 mL of window cleaner containing 9.1% EGBE and 2.5% ethanol by a 53-year-old alcoholic
male. He was comatose with metabolic acidosis, shock and noncardiogenic pulmonary edema
when brought to a hospital, approximately 10 hours after ingestion. He had increased heart rate,
decreased blood pressure, and transient polyuria and hypoxemia. Hypochromic anemia was
evident with an Hb concentration of 9.1 g/100 mL, a hematocrit (Hct) of 25%, and thrombo-
cytopenia. The patient recovered and was discharged after 15 days.
Gijsenbergh et al. (1989) reported that a 23-year-old woman weighing 64 kg ingested
approximately 25-30 g of EGBE (-400-500 mg/kg) and ethanol (~4:1 ratio) as a window
cleaner in an apparent suicide attempt. She was comatose when admitted to the hospital,
exhibiting dilated pupils, obstructive respiration, and metabolic acidosis, including depression of
blood Hb concentration and hematuria. The presence of EGBE in the blood and dialysis fluid
was confirmed. Treatment consisted of supportive therapy, forced diuresis, bicarbonate
administration, and hemodialysis. Her Hb concentration fell from 11.9 g Hb/100 mL upon
admission to 8.9 g Hb/100 mL. She was discharged after 8 days.
Gualtieri et al. (2003, 1995) reported a case of a suicide attempt with an industrial-
strength window cleaner. The 18-year-old male weighed 71 kg; he consumed between 360 and
480 mL of a concentrated glass cleaner that contained 22% EGBE, a dose equivalent to 1,131—
1,509 mg/kg. He was admitted to the hospital with no abnormalities other than epigastric
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discomfort within 3 hours postingestion. Approximately 10 hours postadmission, the patient was
noticeably lethargic, weak, and hyperventilating, symptoms consistent with the onset of
metabolic acidosis. BAA was measured; the highest serum concentration found was
4.86 mmol/L, collected approximately 16 hours postingestion. The patient was transferred to a
tertiary care hospital where hemodialysis was initiated at approximately 24 hours postingestion.
Ethanol therapy was started 30 minutes later. Treatment also consisted of intravenous (i.v.) doses
of 100 mg thiamine and 50 mg folic acid every 12 hours and 50 mg pyridoxine every 6 hours.
Following 4 hours of dialysis, the patient was alert and remained hemodynamically stable. Ten
days after discharge, the patient was readmitted following a second ingestion of 480 mL of the
same cleaner, an EGBE dose equivalent to 1,509 mg/kg. Treatment included ethanol therapy and
hemodialysis, and was initiated within a few hours of ingestion to control the metabolic acidosis.
Due to this early treatment, ethanol therapy had an impact on the disposition of EGBE and BAA.
As with the first episode, metabolic acidosis was manifest. This high-dose oral ingestion was
nearly 1.1-1.5 g EGBE/kg body weight. The highest serum BAA concentration was
2.07 mmol/L, collected 22 hours postingestion. No evidence of hemolysis or renal abnormalities
was detected.
A 5 0-year-old woman ingested approximately 250-500 mL of a window cleaner
containing 12% EGBE, representing -30-60 mL, in an apparent suicide attempt (Rambourg-
Schepens et al., 1988). She was diagnosed with metabolic acidosis, hypokalemia, a rise in serum
creatinine level, and a marked increase in urinary excretion of oxalate crystals. Moderate
hemoglobinuria appeared on the third day postexposure, and a progressive erythropenia was
noted. In the absence of more complete hematologic details from this and other similar case
studies, it is not possible to determine whether these effects were due to hemolysis or other
factors related to the profound blood chemistry changes observed. The clinical status improved
gradually and the patient was discharged on the 10th day.
Burkhart and Donovan (1998) summarized the case of a 19-year-old male who ingested
20-30 ounces (or -590-885 mL) of a product that contained 25-35% EGBE (an exposure
equivalent to -177-265 mL, estimated at >3,000 mg/kg) along with 15-25%) propylene glycol,
5-10%o monoethanolamine, and l-3%> potassium hydroxide. On his arrival at the hospital
3.5 hours after ingestion, the patient was deeply comatose with severe hypotension. Hematuria
developed on the second day, with no evidence of renal or hepatic toxicity; however, pulmonary
toxicity consisting of severe aspiration pneumonia was present. The patient had a significant
recovery, despite severe neurologic deficits that were slow to resolve.
Osterhoudt (2002) reported on a 16-month-old girl who ingested an unknown amount of
cleaning solution containing EGBE (10—30%>), monoethanolamine (5-10%>), alkoxylated linear
alcohols (l-5%>), ethylenediaminetetraacetic acid (l-5%>), and potassium hydroxide (l-5%>).
Metabolic acidosis was manifest, and a single dose (15 mg/kg) of the ALDH inhibitor
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fomepizole was administered. Within 2 hours, the metabolic acidosis was completely resolved,
and there was no evidence of alkaline mucosal injury, hepatic or renal dysfunction, or hemolysis.
Dean and Krenzelok (1991) reported that 24 children, aged 7 months to 9 years, were
observed subsequent to oral ingestion of at least 5 mL of glass window cleaner containing EGBE
in the 0.5-9.9% range. Two children drank more than 15 mL, and were treated by gastric lavage.
No symptoms of EGBE poisoning, such as metabolic acidosis, and no hemolysis were observed
in any of the children.
Raymond et al. (1998) reported on seven clerical workers who were evaluated 8 months
after they entered a file room where the supervisor believed that EGBE had been applied
overnight to strip the floor. Exact details of the product used were unknown, but based on
containers found and exposure symptoms of noted intense eye and respiratory irritation, marked
dyspnea, nausea, and faintness, the authors suggested that they were exposed to EGBE
concentrations of 200-300 ppm. Of major concern were skin spots—cherry angiomas—that
appeared between 4 and 22 weeks after exposure in six of the seven workers. All workers
continued to experience recurrent eye and tracheobronchial irritation; four had a dry cough.
Workplace air sampling conducted by a certified industrial hygienist 1 week after the floor
stripping found no detectable EGBE, although traces (0.1-0.2 ppm) of formaldehyde were
identified. Five years after the exposure, four of the workers who could be contacted reported
that they continued to have outbreaks of new cherry angiomas. It should be noted that no other
studies linking EGBE exposure to outbreaks of cherry angiomas are available in the literature.
The authors included the observation that, since this report, they had seen three patients who they
believe were also exposed to EGBE vapor in an unrelated incident, and who did not develop any
skin spots. Cherry angiomas are the most common cutaneous vascular lesion; they are benign
and formed by a proliferation of dilated venules. The spots occur more frequently with
increasing age but can appear in younger individuals. There are reports in the literature of cherry
angiomas appearing following individual exposure to other chemicals, such as bromides (Cohen
et al., 2001), glutaraldehyde (Raymond et al., 1998), and sulfur mustard gas (Firooz et al., 1999).
A cross section of 31 male workers, aged 22-45, employed for 1-6 years, who were
exposed to low levels of EGBE in a beverage packing production plant were monitored by
Haufroid et al. (1997). The effect of external EGBE exposure and internal BAA levels on
erythrocyte lineage were investigated by studying red blood cell (RBC) count, Hb, Hct, mean
cell volume (MCV), mean corpuscular hemoglobin (MCH), mean corpuscular hemoglobin
concentration (MCHC), haptoglobin (Hp), reticulocyte count, and osmotic resistance (OR), a
measure of osmotic fragility. Also studied were serum glutamic-oxaloacetic and glutamic-
pyruvic transaminases and renal creatinine and urinary retinol binding protein parameters. The
average airborne concentration of EGBE was 2.91 mg/m3, or 0.6 ppm (standard deviation [SD]
of ±1.30 mg/m3 or 0.27 ppm). In addition, there was coexposure to methyl ethyl ketone. Single
determinations of BAA in postshift urine samples were used to assess exposure to low levels of
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EGBE. No differences were observed for RBC counts, Hb, MCV, MCH, Hp, reticulocyte count,
or between exposed and control workers. The only statistically significant change observed in
exposed workers when compared with a matched control group (n = 21) was a 3.3% decrease in
Hct (p = 0.03) and a 2.1% increase in MCHC (p = 0.02). The implications of these small
erythroid effects are unclear. Both values are within their corresponding normal clinical ranges
and, given that no statistically significant changes were observed in other erythroid parameters,
they do not appear to be related to the more severe effects observed in laboratory animals.
Furthermore, no correlation was found between any of the nine erythroid parameters measured
and the parameters of internal exposure. No significant differences were observed in hepatic and
renal biomarkers.
Several human studies investigated the dermal absorption of EGBE. Jakasa et al. (2004)
dermally exposed six male research subjects, ages 22-55, to 50%, 90%, or neat EGBE for
4 hours on the forearm over an area of 40 cm2. The dermal absorption of EGBE from aqueous
solutions was markedly higher than from neat EGBE. In Jones et al. (2003), four research
subjects were exposed via inhalation of 50 ppm EGBE for 2 hours on nine separate occasions,
with each occasion separated by 3 weeks, at varying temperatures and humidity levels. Results
show that "baseline" dermal contribution to total body absorption of EGBE vapor in
appropriately dressed workers was, on average, 11%. Higher temperature (30°C, mean 14%,
p = 0.03) and greater humidity (65% relative humidity, mean 13%,p = 0.1) both increased
dermal absorption. The wearing of whole-body overalls did not attenuate absorption (mean
10%). By combining several factors together in the industrial scenario, dermal absorption of
vapors was reported to be as high as 39% of the total absorbed dose.
4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN
ANIMALS—ORAL AND INHALATION
4.2.1. Subchronic Studies
4.2.1.1. Oral
A number of subchronic studies by the gavage route of exposure have been conducted.
Krasavage (1986) conducted a toxicity study using groups of 10 COBS CD (Sprague-Dawley)
BR adult male rats treated by gavage with 222, 443, or 885 mg/kg-day undiluted EGBE
5 days/week for 6 weeks. Endpoints evaluated throughout the study included body weight, food
consumption, clinical signs, and survival. Hematology and serum clinical chemistry parameters
were determined after the last treatment. Dose-related changes were observed in the RBC counts
of all treatment groups, including statistically significant decreases in RBC count and Hb
concentration and a statistically significant increase in MCH. Statistically significant
hematological changes occurring at 443 and 885 mg/kg-day were increased MCV and decreased
MCHC. The increased MCV at higher doses is likely due to both an increase in MCV and in the
number of larger reticulocytes in the circulation following the erythropoietic response (NTP,
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2000). Based on decreased RBC count and trends in Hb and other hematological endpoints, the
lowest-observed-adverse-effect level (LOAEL) was determined to be 222 mg/kg-day, the lowest
dose tested. A no-observed-adverse-effect level (NOAEL) was not identified.
Nagano et al. (1979) performed a toxicity study in male JCL/ICR mice (five/group) using
gavage doses of 0, 357, 714, or 1,430 mg/kg-day EGBE 5 days/week for 5 weeks. Parameters
evaluated at the end of the study were hematology (RBC and white blood cell [WBC] counts,
MCV, and Hb), absolute and relative weights of testes, and testicular histology. Mean RBC
counts were significantly lower than the control values in the 357 and 714 mg/kg-day groups, but
WBC counts were not affected. All animals in the 1,430 mg/kg-day group died before
examinations were performed; mortality was not observed in the lower dose groups, and no
differences in testes weights or histology were found. The LOAEL for this study, based on the
reduced RBC count, was 357 mg/kg-day. A NOAEL was not determined.
Siesky et al. (2002) investigated whether subchronic exposure to EGBE in F344 male
rats, 8-10 weeks old, and B6C3Fi male mice, 8-10 weeks old, produced an increase in oxidative
damage and deoxyribonucleic acid (DNA) synthesis in endothelial cells and hepatocytes in the
mouse liver, the putative cancer target cells. Mice (60/group) were treated via gavage with doses
of 0, 225, 450, and 900 mg/kg-day and rats (20/group) with 0, 225, and 450 mg/kg-day for
90 days. A dose-related increase in hemolysis was observed in both rats and mice. An increase
in the percentage of iron-stained Kupffer cells was observed following 450 and 900 mg/kg in
mice and 225 and 450 mg/kg in rats. An increase in oxidative damage, as measured by
8-hydroxydeoxyguanosine (8-OHdG) levels, was seen in mouse livers at 450 and 900 mg/kg-day
after 7 or 90 days, while no increase was seen in rat livers at any dose or time point examined.
Vitamin E levels were reduced by all doses of EGBE in the mouse and rat liver (statistically
significant at 7 and 90 days in both mice and rats); however, the basal level of vitamin E was
~2.5-fold higher in rat than in mouse liver. The LOAEL for this study was 450 mg/kg-day in
mice and 225 mg/kg-day in rats, based on the percentage of iron-stained Kupffer cells. The
NOAEL was 225 mg/kg-day in mice, and a NOAEL was not determined in rats.
NTP (1993) performed a 13-week toxicity study in F344 rats and B6C3Fi mice where
groups of 10 animals/gender/species received EGBE in drinking water at doses of 0, 750, 1,500,
3,000, 4,500, and 6,000 ppm in rats and 0, 750, 1,500, 3,000, 4,500, and 6,000 ppm in mice. The
corresponding doses in mg/kg-day, based on measured drinking water consumption were: 0, 69,
129, 281, 367, or 452 mg/kg-day in male rats; 0, 82, 151, 304, 363, or 470 mg/kg-day in female
rats; 0, 118, 223, 553, 676, or 694 mg/kg-day in male mice; and 0, 185, 370, 676, 861, or
1,306 mg/kg-day in female mice. Due to a dose-related decrease in water consumption in the
2-week studies, the test chemical was administered at a constant concentration (ppm) in the
13-week studies rather than on a mg/kg body weight basis. Hematology was performed on rats
but not on mice. Complete histological exams were performed on all control animals and all
animals in the highest dose group. Vaginal cytology and sperm indices were evaluated in rats
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and mice from the control and three highest dose groups. Hematologic changes in both genders
persisting until or developing by 13 weeks included dose-related indications of mild to moderate
anemia. Portions of the hematologic results from the NTP 13-week rat drinking-water study are
presented in Table 4-1. The various results shown in this table are indicative of the various
degrees of hemolysis caused by exposure to increasing concentrations of EGBE. Overall, the
dose-related increase in MCV is indicative of erythrocyte swelling that would be expected to
precede cell lysis and an increase in the number of reticulocytes. Deficits in RBCs as a result of
lysis manifest through dose-related decreases in the measures of both RBC count and Hb
concentration. Hct would be expected to decrease but did not. The increases noted both in
reticulocytes (young RBCs) and, at higher doses, in nucleated erythrocytes (immature and
prematurely released blood cells) are homeostatic responses that would be anticipated to occur as
the lysed blood cells are being replaced. More specifically, male rats evaluated at 13 weeks
showed significantly reduced RBC counts at >281 mg/kg-day (3,000 ppm) and reduced Hb
concentration, reduced platelets, and increased bone marrow cellularity at >367 mg/kg-day
(4,500 ppm). These data also suggest that female rats are more sensitive to the effects from
EGBE, since several statistically significant effects occurred at the 750 ppm concentration, the
lowest level tested in females; males did not show statistically significant effects until two dose
levels higher (3,000 ppm). In addition, the degree to which these various measures were affected
was somewhat greater in females than males (indicated as percent control in the tables),
particularly at the three highest exposure concentrations.
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Table 4-1. Hematology and hemosiderin data from the 13-week drinking water exposure to EGBE in F344 rats
Endpoint"
Control
750 ppm
1,500 ppm
3,000 ppm
4,500 ppm
6,000 ppm
N
Males
Females
8
10
10
10
10
10
10
10
10
10
9
9
Hct (%)
Males
Females
44.8 ±0.8
44.8 ±0.6
45.0 ±0.6 (100)
43.2 ±0.8 (96)
44.7 ± 0.4 (99)
42.8 ±0.7 (95)
44.1 ±0.7 (98)
43.6 ±0.7 (97)
42.3 ± 0.6b (94)
44.4 ± 0.7 (99)
43.4 ±0.4 (97)
46.1 ±0.7 (103)
Hb (g/dL)
Males
Females
15.0 ±0.2
14.9 ±0.2
15.2 ±0.1 (101)
14.4 ± 0.2b (97)
14.9 ±0.1 (99)
13.9 ±0.2 (93)
14.6 ±0.1 (97)
14.2 ±0.2 (95)
14.0 ± 0.1° (93)
14.0 ± 0.2° (94)
13.7 ± 0.2° (91)
13.4 ± 0.2° (90)
Erythrocytes (10'/|iL)
Males
Females
8.64 ±0.15
8.15 ±0.09
8.74 ±0.10 (101)
7.59 ± 0.15° (93)
8.54 ± 0.09 (99)
7.09 ± 0.14° (87)
8.11 ± 0.12b (94)
7.00 ± 0.12° (86)
7.48 ± 0.12° (86)
6.80 ±0.11° (83)
7.18 ± 0.12° (83)
6.58 ± 0.14° (81)
Reticulocytes (106/jxL)
Males
Females
0.14 ±0.03
0.12 ±0.02
0.24 ± 0.06
0.17 ±0.03
0.15 ±0.02
0.19 ±0.03
0.18 ±0.02
0.28 ± 0.03°
0.22 ±0.05
0.28 ± 0.05°
0.46 ± 0.07°
0.27 ± 0.05°
Nucleated erythrocytes (10 7|iL)
Males
Females
0.00 ± 0.00
0.01 ±0.01
0.00 ± 0.00
0.03 ±0.02
0.01 ±0.01
0.02 ±0.01
0.01 ±0.01
0.05 ±0.02
0.00 ± 0.00
0.10 ± 0.03b
0.04 ± 0.02°
0.16 ± 0.04°
MCV (fL)
Males
Females
52.0 ±0.4
54.8 ±0.3
51.5 ±0.3 (99)
57.0 ±0.4C (104)
52.3 ±0.4 (100)
60.5 ±0.4C (110)
54.4 ±0.3C (105)
62.4 ±0.6C (114)
56.7 ±0.5C (109)
65.3 ± 0.6°(119)
60.6 ± 1.1° (116)
70.1 ± 0.9° (128)
MCH (pg)
Males
Females
17.4 ±0.2
18.3 ±0.2
17.4 ±0.1
18.9 ±0.2
17.5 ±0.2
19.7 ±0.2
18.0 ± 0.2b
20.2 ± 0.3°
18.7 ± 0.3°
20.6 ± 0.2°
19.1 ± 0.3°
20.4 ± 0.1°
Hemosiderin (incidence)
Males
Females
0/10
0/10
0/10
0/10
0/10
2/10
0/10
10/10
0/10
10/10
7/10
10/10
"Values listed are mean ± standard error at various EGBE concentrations in ppm for the 13-week time point. Percent of control values in parentheses.
Statistically significant difference, p < 0.05.
Statistically significant difference,/) <0.01.
Source: NTP(1993).
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Statistically significant hematologic effects in female rats at week 13 included reduced
RBC counts and Hb concentrations at >82 mg/kg-day and increased reticulocytes, decreased
platelets, and increased bone marrow cellularity at approximately 304 mg/kg-day, all being
indicative of hemolysis. There were no histopathological changes in the testes or epididymis at
>129 mg/kg-day.
Table 4-2 shows that liver effects, including cytoplasmic alterations, hepatocellular
degeneration, and pigmentation were observed in the mid- and high-dose groups (129, 281, 367,
and 452 mg/kg-day for males and 151, 304, 363, and 470 mg/kg-day for females; statistics not
reported). As with the hematologic effects, these effects appeared to be more severe in females
than in males. Cytoplasmic alterations of liver hepatocytes, consisting of hepatocytes staining
more eosinophilic and lacking the basophilic granularity of the cytoplasm present in hepatocytes
from control animals, were observed in the lowest-dose groups tested (69 mg/kg-day for males
and 82 mg/kg-day for females). The lack of cytoplasmic granularity or "ground-glass"
appearance of the hepatocytes suggests that this response was not due to enzyme induction
(Greaves, 2000). The cytoplasmic alterations were judged to increase in severity in both
genders, but especially in females, with the severity in the two highest dose groups being judged
as "moderate." Liver pigmentation, colored brown to green and staining strongly positive for
iron (indicative of hemosiderin accumulation), was noted in the cytoplasm of Kupffer cells in
both genders of rats. In females, liver pigmentation was noted in 0/10 controls and 0/10 at
82 mg/kg-day, 2/10 with a severity grade of 1 (minimal) at 151 mg/kg-day, and 10/10 in the three
highest dose levels; the severities increased from a numerical grade of 1.2 in the 304 mg/kg-day
group to 1.9 the upper two dose groups. In males, the hemosiderin pigmentation was noted in
animals exposed to the highest dose only (452 mg/kg-day) at an incidence of 7/10 and a severity
rating of 1 (minimal). No hepatic pigmentation was reported in the mice exposed for 13 weeks.
The hematological (decreased RBC count and Hb) and hepatic changes were dose-related; 69-
82 mg/kg-day was considered a LOAEL. ANOAEL was not identified.
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Table 4-2. Incidence3 and severity of selected histopathological changes from the 13-week drinking water exposure to
EGBE in F344 rats and mice
Control
750 ppm
1,500 ppm
3,000 ppm
4,500 ppm
6,000 ppm
Rat
N
10
10
10
10
10
10
Liver cytoplasmic alterations (NR)
Males
0
4(1.0)
8(1.0)
7(1.1)
10 (2.0)
10(1.8)
Females
0
5 (1.4)
9 (2.0)
10 (2.2)
10 (3.0)
10 (3.0)
Hepatocellular degeneration (NR)
Males
0
0
0
8(1.0)
8(1.0)
8(1.0)
Females
0
0
0
10(1.3)
10(1.3)
10(1.1)
Kupffer cell pigmentation (NR)
Males
0
0
0
0
0
7(1.0)
Females
0
0
2(1.0)
10(1.2)
10(1.9)
10(1.9)
Mouse
N
10
10
10
10
10
10
Necropsy body weight (g)b
Females (only)
31.1 ± 0.7
31.8 ± 0.8
30.9 ± 1.5
28.0 ± 0.7°
28.4 ± 0.5°
27.8 ± 0.9d
Relative kidney weight (right) (mg organ wt/g body wt)b
Females (only)
6.33 ±0.10
7.69 ± 0.14d
8.06 ± 0.29d
7.47 ± 0.19d
7.55 ± 0.18d
8.21 ±0.26d
incidences represent the number of animals with lesions. Average severity (in parentheses) is based on the number of animals with lesions; 1 = minimal, 2 = mild,
3 = moderate, 4 = marked.
bMean ± standard error.
Statistically significant difference, p < 0.05.
Statistically significant difference, p < 0.01.
NR = Statistics not reported
Source: NTP(1993).
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38
Female mice showed statistically significant reductions in body weight gain starting at
3,000 ppm and statistically significant increases in relative kidney weight at all doses. Changes
at the higher doses followed the reductions in body weight at those dose levels, while the
increases at lower doses (750 and 1,500 ppm) were due to increased absolute kidney weights at
those doses. Body weight reduction followed the decreased water consumption data. No
histopathologic changes were noted at any dose level, even though relative kidney weights
showed a statistically significant increase at 750 and 1,500 ppm in the absence of reduction in
body weight gain.
Keith et al. (1996) administered EGBE at 120 mg/kg-day for 120 days by gavage to
transgenic FVB/N mice (25 mice/gender/group) carrying the v-Ha-ras oncogene and observed
the animals for an additional 120 days. EGBE did not induce an increase in tumors at any site.
4.2.1.2. Inhalation
Wistar rats (23 animals/group, gender not specified) were exposed to 0, 135, or 320 ppm
EGBE 7 hours/day, 5 days/week for 5 weeks (Werner et al., 1943a). Hematologic endpoints—
RBC, WBC, differential, and reticulocyte counts and Hb concentration—were evaluated.
Exposure to 320 ppm EGBE resulted in an increased percentage of circulating immature
granulocytes, a decreased Hb concentration and RBC count, and an increased reticulocyte count.
These hematologic changes were not severe; they were reversed 3 weeks after discontinuing
exposure. No effect on WBC count was observed. In another study, Werner et al. (1943b)
exposed groups of two dogs of unspecified strain to subchronic inhalation doses of 0 or 415 ppm
EGBE 7 hours/day, 5 days/week for 12 weeks. Necropsies were performed 5 weeks
postexposure; hematologic parameters were examined before, during, and after the exposure. No
statistical analysis was presented. The authors concluded that exposure of dogs to EGBE vapors
resulted in decreased Hb concentration and RBC count with increased hypochromia,
polychromatophilia, and microcytosis. These hematologic effects were not severe and they were
reversed 5 weeks after the end of exposure.
Carpenter et al. (1956) studied the hemolytic effects of EGBE vapor inhalation in rats,
mice, dogs, and monkeys, in addition to humans. An unspecified strain of rats (15/gender/group)
was exposed via inhalation to 54, 107, 203, 314, or 432 ppm EGBE 7 hours/day, 5 days/week for
6 weeks. Erythrocyte osmotic fragility was observed in rats immediately after a single 7-hour
exposure to >107 ppm. Osmotic fragility in females exceeded that for males. In almost all
cases, these high fragility values returned to normal after the rats rested overnight. In the same
study, the authors exposed groups of 10 male C3H mice to 100, 200, or 400 ppm EGBE
7 hours/day for 30, 60, or 90 days. An increase in erythrocyte osmotic fragility occurred at all
concentrations and was consistent throughout the exposures. In all instances, erythrocyte
osmotic fragility was normal after a 17-hour rest period. The LOAELs for these rat and mouse
studies were 54 and 100 ppm, respectively. NOAELs were not reported. The authors reported
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34
35
36
37
38
slight increases in erythrocyte osmotic fragility for a male and a female dog (basenji hybrids)
exposed to 200 ppm EGBE for 31 days (7 hours/day). RBC counts and Hb concentrations were
slightly decreased in the female. Erythrocyte permeability, as determined by radio-iodine
uptake, was increased in both genders, but was not statistically different when compared with
control values. A female dog succumbed after 8 days of inhalation exposure to 385 ppm of
EGBE (7 hours/day). Symptoms included loss of weight, transitory increases in erythrocyte
osmotic fragility, nasal and ocular infection, weakness, apathy, anorexia, and increased WBC
count. Necropsy of this animal revealed severe congestion and hemorrhage of the lungs and
congestion of the liver and both kidneys. In addition, a severe subcapsular hemorrhage was
found in one adrenal gland. A male dog survived after 28 days of inhalation exposure to
385 ppm of EGBE for 7 hours/day. Effects in the male were similar to the female, but developed
more slowly. At autopsy, congestion of the kidneys was not observed. In studies on male and
female monkeys, occasional rises in erythrocyte osmotic fragility were reported that were more
frequent in the female than in the male following 90-day inhalation exposure to 100 ppm of
EGBE.
A90-day subchronic inhalation study was performed using F344 rats (16/gender/group)
exposed to EGBE for 6 hours/day, 5 days/week at concentrations of 0, 5, 25, and 77 ppm (Dodd
et al., 1983). After 31 completed exposures (6 weeks), the 77 ppm female rats had slight but
statistically significant decreases in RBC counts (13% below control value) and Hb
concentrations, accompanied by an 11% increase above the control value in MCH. The 77 ppm
males exhibited slight (5%) but statistically significant decreases in RBC counts and Hb
concentration that were accompanied by increases in MCH. At the end of the 90-day study
(66 exposures), the hematologic effects seen in the 77 ppm exposed animals had either lessened
or returned to the ranges of control values and were no longer statistically significant. The
NOAEL was determined to be 25 ppm, and the LOAEL was 77 ppm.
In the subchronic portion of the inhalation NTP (2000) study, F344 rats and B6C3Fi mice
(10/gender) were exposed to EGBE concentrations of 0, 31, 62.5, 125, 250, and 500 ppm (0, 150,
302, 604, 1,208, and 2,416 mg/m3) 6 hours/day, 5 days/week for 14 weeks. Hematologic and
hemosiderin staining results are presented in Tables 4-3 and 4-4. These results are indicative of
the various degrees of hemolysis caused by exposure to increasing concentrations of EGBE.
Both rat genders exhibited clinical signs at the three highest doses, consistent with the hemolytic
effects of EGBE, including: (1) deficits in RBCs as a result of lysis manifestation through the
dose-related decrease in Hct—a finding consistent with decreases noted for both RBC count and
Hb concentrations; and (2) increases in both reticulocytes and nucleated erythrocytes at higher
doses—homeostatic responses that would be anticipated to occur as the lysed blood cells are
being replaced. Female rats may be somewhat more sensitive; several statistically significant
effects occur at the 31 ppm level in females, as opposed to a single parameter for males. In
addition, the degree to which these various measures are affected is somewhat greater in females
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than males (indicated as percent control) particularly at the three highest concentrations.
Hematologic evaluation showed mild-to-moderate regenerative anemia at all concentrations in
females and at the three highest concentrations in males. Exposure-related trends were noted for
reticulocyte count, RBC count, MCV, Hb concentration, and Hct. Liver-to-body-weight ratios
increased significantly in males at the two highest concentrations and in females at the highest
concentration. Histopathologic effects at concentrations >62.5 ppm for male rats and >31 ppm
for females consisted of excessive splenic congestion in the form of extramedullary
hematopoiesis, hemosiderin accumulation in Kupffer cells, liver necrosis, centrilobular
hepatocellular degeneration, renal tubular degeneration, intracytoplasmic Hb and hemosiderin
deposition, and bone marrow hyperplasia. In addition, five moribund female rats were sacrificed
from the highest concentrations and one from the 250 ppm group. The LOAEL for
hematological alterations was 31 ppm for female rats and 62.5 ppm for male rats. The 31 ppm
exposure level was considered a NOAEL for male rats.
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Table 4-3. Hematology and hemosiderin data from a 14-week inhalation study of EGBE in F344 rats
Endpoint"
Control
31 ppm
(150 mg/m3)
62.5 ppm
(302 mg/m3)
125 ppm
(604 mg/m3)
250 ppm
(1,208 mg/m3)
500 ppm
(2,416 mg/m3)
Hct (%)
Males
Females
46.8 ±0.5
48.5 ±0.5
45.8 ±0.6 (98)
46.0 ± 0.5° (95)
47.0 ±0.4 (100)
45.2 ± 0.5° (93)
44.5 ± 0.5° (95)
42.9 ± 0.4° (88)
41.1 ± 0.3° (88)
40.0 ± 0.3° (82)
37.3 ± 0.4° (80)
36.2 ± 0.6° (75)
Hb (g/dL)
Males
Females
15.5 ±0.1
15.6 ±0.1
14.8 ±0.3 (95)
15.0 ± 0.1° (96)
15.4 ±0.1 (99)
14.6 ± 0.1° (94)
14.5 ± 0.2° (94)
13.6 ± 0.1° (87)
13.1 ± 0.1° (85)
12.5 ± 0.1° (80)
11.7 ± 0.1° (75)
10.5 ± 0.3° (67)
Erythrocytes (106/|iL)
Males
Females
9.05 ±0.08
8.48 ±0.05
8.71 ± 0.14b (96)
8.08 ± 0.07° (95)
8.91 ±0.06 (98)
7.70 ± 0.08° (91)
8.01 ± 0.08° 89)
6.91 ± 0.05° (81)
7.10 ± 0.07° (78)
6.07 ± 0.04° (72)
5.97 ± 0.05°(66)
4.77 ± 0.15° (56)
Reticulocytes (106/jxL)
Males
Females
0.16 ±0.02
0.13 ±0.02
0.17 ±0.03
0.10 ±0.01
0.15 ±0.02
0.16 ±0.02
0.30 ± 0.04°
0.26 ± 0.04b
0.48 ± 0.06°
0.34 ± 0.04°
0.68 ± 0.07°
0.40 ± 0.11°
Nucleated erythro (107|iL)
Males
Females
0.04 ± 0.02
0.04 ± 0.02
0.05 ±0.01
0.05 ±0.02
0.04 ±0.03
0.12 ± 0.03b
0.11 ±0.03
0.18 ±0.07
0.17 ± 0.04°
0.61 ± 0.24°
0.20 ± 0.06°
0.73 ± 0.27°
MCV (fL)
Males
Females
50.4 ±0.3
55.1 ±0.3
50.2 ±0.2 (100)
55.3 ±0.2 (100)
50.7 ±0.2 (100)
56.4 ±0.2 (102)
53.1 ±0.2C (105)
58.7 ±0.2C (107)
53.8 ±0.3C (107)
61.6 ±0.2C (112)
58.5 ± 0.3°(117)
66.8 ±0.9C (121)
MCH (pg)
Males
Females
17.1 ±0.1
18.4 ±0.1
17.0 ±0.1
18.6 ±0.2
17.3 ±0.1
19.0 ± 0.0°
18.1 ± 0.1°
19.6 ± 0.1°
18.4 ± 0.1°
20.6 ± 0.1°
19.5 ± 0.1°
22.0 ± 0.1°
Hemosiderin (incidence)
Males
Females
0/10
0/10
0/10
0/10
0/10
10/10
7/10
10/10
10/10
9/9
10/10
5/5
"Values listed are mean ± standard error (percent of control).
Statistically significant difference, p < 0.05.
Statistically significant difference,/) <0.01.
Source: NTP (2000).
1
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Table 4-4. Hematology and hemosiderin data from a 14-week inhalation study of EGBE in B6C3Fi mice
Endpoint"
Control
31 ppm
(150 mg/m3)
62.5 ppm
(302 mg/m3)
125 ppm
(604 mg/m3)
250 ppm
(1,208 mg/m3)
500 ppm
(2,416 mg/m3)
Hct (%)
Males
Females
47.3 ± 1.0
46.2 ±0.3
48.3 ±0.4 (102)
45.9 ±0.3 (99)
47.6 ±0.5 (101)
45.8 ±0.3 (99)
46.6 ± 0.4 (99)
45.1 ± 0.2° (98)
44.2 ± 0.4° (93)
42.3 ± 0.4° (92)
36.3 ± 1.4° (77)
37.8 ± 1.0° (82)
Hb (g/dL)
Males
Females
15.7 ±0.4
15.7 ±0.1
16.0 ±0.1 (102)
15.4 ± 0.1b (98)
15.9 ±0.1 (101)
15.4 ± 0.1b (98)
15.4 ± 0.1° (98)
14.8 ± 0.1° (94)
14.4 ± 0.1° (92)
13.7 ± 0.1° (87)
11.4 ± 0.4° (73)
11.6 ± 0.1° (74)
Erythrocytes (106/|iL)
Males
Females
9.71 ±0.22
9.72 ±0.05
10.04 ±0.08 (103)
9.55 ± 0.06b (98)
9.77 ±0.1 (101)
9.51 ± 0.06b (98)
9.47 ± 0.06b (98)
9.18 ± 0.05° (94)
8.90 ± 0.07° (92)
8.57 ± 0.06° (88)
7.21 ±0.23c(74)
7.35 ± 0.07° (76)
Reticulocytes (10'7|.iL)
Males
Females
0.21 ±0.03
0.18 ±0.02
0.22 ±0.03
0.21 ±0.03
0.21 ±0.02
0.19 ±0.02
0.32 ± 0.03b
0.29 ± 0.02°
0.45 ± 0.04°
0.47 ± 0.04°
0.79 ± 0.20°
1.17 ± 0.28°
MCV (fL)
Males
Females
49.1 ±0.4
48.3 ±0.3
48.5 ±0.3 (99)
48.8 ±0.2 (101)
49.0 ±0.4 (100)
48.8 ±0.2 (101)
49.7 ±0.4 (101)
49.5 ±0.5 (102)
49.8 ±0.4 (101)
49.0 ±0.3 (101)
48.3 ±0.9 (98)
48.8 ± 1.0(101)
MCH (pg)
Males
Females
16.2 ±0.1
16.1 ±0.1
16.0 ±0.1 (99)
16.0 ±0.1 (99)
16.2 ±0.1 (100)
16.2 ±0.1 (101)
16.2 ±0.0 (100)
16.1 ±0.1 (100)
16.2 ±0.1 (100)
16.0 ± 0.0 (99)
15.8 ±0.2 (98)
15.8 ±0.1 (98)
Hemosiderin (incidence)
Males
Females
0/10
0/10
0/10
0/10
0/10
0/10
0/10
0/10
0/10
10/10
6/6
6/6
aValues listed are mean ± standard error (percent of control).
Statistically significant difference, p < 0.05.
Statistically significant difference,/) <0.01.
Source: NTP (2000).
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25
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27
28
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30
31
32
33
34
35
The mice exposed via the inhalation route exhibited clinical signs consistent with the
hemolytic effects of EGBE at the two highest concentrations for both genders (NTP, 2000).
Hematologic evaluation indicated a moderate regenerative anemia (marked by decreased RBC
counts, increased reticulocyte counts, and increased MCV) with an increase in platelets at the
three highest concentrations in both genders. Histopathological effects consisted of excessive
extramedullar splenic hematopoiesis, renal tubular degeneration, hemosiderin deposition in the
spleen and kidney and accumulation in Kupffer cells, and testicular degeneration. Forestomach
necrosis, ulceration, inflammation, and epithelial hyperplasia were observed at concentrations
>31 ppm for females and >62.5 ppm for males. In addition, four females and four males either
died or were sacrificed moribund at the highest concentration. The NOAEL for male and female
mice was 31 ppm and the LOAEL was 62.5 ppm, based on histopathological changes in the
forestomach.
4.2.2. Chronic Studies and Cancer Bioassays
4.2.2.1. Inhalation
NTP (2000) also completed a two-species, 2-year inhalation study on EGBE in both
genders of rats and mice. In this chronic study, animals were exposed to EGBE 6 hours/day,
5 days/week at concentrations of 0, 31, 62.5, and 125 ppm (0, 150, 302, and 604 mg/m3) for
groups of 50 F344/N rats and 0, 62.5, 125, and 250 ppm (0, 302, 604, and 1,208 mg/m3) for
groups of 50 B6C3Fi mice. The researchers stated that the highest exposure was selected to
produce a 10-15% depression in hematologic indices. They reported that no effect on survival
was observed in rats, but survival was statistically significantly decreased in male mice exposed
to 125 or 250 ppm, compared with chamber controls (54, 52, and 78% respectively [NTP, 2000]).
Although statistics were not reported for mean body weights, the rats exposed to 31 and
62.5 ppm had similar mean body weights to the control rats. Mean body weights of the exposed
mice were generally less than those of controls, with females experiencing greater and earlier
reductions. From week 17 to the end of the study, the mean body weights of 125 ppm female
rats were generally less than those of controls. Nonneoplastic effects in rats included hyaline
degeneration of the olfactory epithelium in males (13/48, 21/49, 23/49, 40/50) and females
(13/50, 18/48, 28/50, 40/49) and Kupffer cell pigmentation in the livers of males (23/50, 30/50,
34/50, 42/50) and females (15/50, 19/50, 36/50, 47/50) (Table 4-5). The severity of the olfactory
lesion was not affected by exposure. The Kupffer cell pigmentation is a result of hemosiderin
accumulation and is a recognized secondary effect of the hemolytic activity of EGBE (NTP,
2000).
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Table 4-5. Selected female and male rat and mouse nonneoplastic effects
from the 2-year chronic EGBE inhalation study
Control
31 ppm
62.5 ppm
125 ppm
250 ppm
Rat
Kupffer cell pigmentation, hemosiderin in the
liver
Male
Female
23/50
15/50
30/50
19/50
34/503
36/503
42/503
47/503
NT
NT
Hyaline degeneration of the olfactory
epithelium
Male
Female
13/48
13/50
21/493
18/48
23/49a
28/503
40/503
40/493
NT
NT
Mouse
Kupffer cell pigmentation, hemosiderin in the
liver
Male
Female
0/50
0/50
NT
NT
0/50
5/50a
8/4 9b
25/49b
30/49b
44/50b
Hematopoietic cell proliferation in the spleen
Male
Female
12/50
24/50
NT
NT
11/50
29/50
26/48b
32/49
42/50b
35/503
Hemosiderin in the spleen
Male
Female
0/50
39/50
NT
NT
6/50a
44/50
45/48b
46/49b
44/49b
48/50b
Forestomach ulcers
Male
Female
1/50
1/50
NT
NT
2/50
7/50a
9/4 9b
13/49b
3/48
22/50b
Forestomach epithelial hyperplasia
Male
Female
1/50
6/50
NT
NT
7/50a
27/50b
16/49b
42/49b
21/48b
44/50b
Hyaline degeneration of the olfactory
epithelium
Females (only)
6/50
NT
14/50
11/49
12/50
Bone marrow hyperplasia
Males (only)
0/50
NT
1/50
9/49a
5/50a
"Statistically significant difference, p < 0.05.
Statistically significant difference, p £ 0.01,
NT = not tested
Source: NTP (2000).
1
2 Nonneoplastic, statistically significant effects in mice included forestomach ulcers and
3 epithelial hyperplasia, hematopoietic cell proliferation and hemosiderin pigmentation in the
4 spleen, Kupffer cell pigmentation in the livers, and bone marrow hyperplasia (males only).
5 Hyaline degeneration of the olfactory epithelium (females only) was increased relative to
6 chamber controls but was not statistically significant. As in the rats, the Kupffer cell
7 pigmentation was considered a secondary effect of the hemolytic activity of EGBE. Bone
8 marrow hyperplasia, hematopoietic cell proliferation, and hemosiderin pigmentation in the
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spleen were also attributed to the primary hemolytic effect; it was followed by regenerative
hyperplasia of the hematopoietic tissue. The forestomach lesions did not appear to be related to
the hemolytic effect of EGBE. Incidences of ulcer were significantly increased in all exposed
female groups, as well as males exposed to 125 ppm. Incidences of epithelial hyperplasia,
usually focal, were significantly increased in all exposed groups of males and females. The
hyperplasia was often associated with ulceration, particularly in the females, and consisted of
thickness of the stratified squamous epithelium and sometimes the keratinized layer of the
forestomach. Ulceration consisted of a defect in the forestomach wall that penetrated the full
thickness of the epithelium and frequently contained accumulations of inflammatory cells and
debris.
Using the same exposure levels described above, additional groups of rats (27/gender/
exposure group) and mice (30/gender/exposure group) in the 2-year study were examined at 3, 6,
and 12 months (8-10 animals/time point) for hematologic effects (NTP, 2000). Nine male and
nine female rats were exposed to 31 ppm EGBE, specifically to evaluate hematology at 3 months
and to receive a total evaluation at 6 months. Animals were continuously exposed, as described
above, until their sacrifice at 3, 6, or 12 months. As in the 14-week study, inhalation of EGBE by
both species resulted in the development of exposure-related hemolytic effects, inducing a
responsive anemia. In rats, the anemia was persistent and did not progress or ameliorate in
severity from 3 months to the final blood collection at 12 months. Statistically significant
(p < 0.05) decreases in automated and manual Hct values, Hb concentrations, and RBC counts
occurred at 3, 6, and 12 months in the 125 and 250 ppm female mice and the 250 ppm male
mice. Statistically significant decreases in these same endpoints were also observed in 62.5 ppm
females at 6 months and in 125 ppm males at 6 and 12 months (decreases in Hct were observed
only at 3 and 6 months). MCV was increased in female mice at the highest duration (12 months)
and exposure (250 ppm) levels. Reticulocyte counts were increased significantly in the 125 ppm
females at 3 and 6 months and in the 125 ppm males at 6 months of exposure. Table 4-6 shows
the responses available for a representative measure of the hematologic effects from EGBE
exposure. Hct levels for male and female rats and mice measured after 3 months or 12 months
are presented.
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Table 4-6. Comparison of female and male rat and mouse Hct (manual)
values from 3- and 12-month inhalation exposures to EGBE
Control
31.2 ppm
62.5 ppm
125 ppm
250 ppm
Female rats3
3 mos
12 mos
46.5 ±0.5
45.4 ±0.2
46.1 ±0.5 (95)
43.3 ± 0.5° (93)
45.3 ±0.3 (100)
42.2 ± 0.5° (91)
42.3 ± 0.4° (93)
-
Male rats3
3 mos
12 mos
44.9 ±0.2
47.8 ±0.4
46.9 ±0.5 (104)
44.8 ±0.4 (100)
45.9 ± 0.8b (96)
42.9 ± 0.5b (95)
42.9 ± 1.2° (90)
-
Female mice3
3 mos
12 mos
49.3 ±0.5
46.9 ±0.4
-
48.9 ±0.4 (99)
46.3 ± 0.4 (99)
46.2 ± 0.5° (94)
43.8 ± 0.4° (93)
43.7 ± 0.5° (89)
41.8 ± 0.3° (89)
Male mice3
3 mos
12 mos
47.5 ±0.3
47.9 ±0.4
-
47.3 ±0.5 (100)
48.7 ± 1.9 (102)
46.0 ± 0.4b (97)
46.4 ± 1.0 (97)
43.7 ± 0.2° (92)
42.1 ± 0.4° (88)
aThese results are from a serial sacrifice conducted as a part of the 2-year chronic inhalation study. Values listed
are mean ± standard error (percent of control).
Statistically significant difference, p < 0.05.
Statistically significant difference,/) <0.01.
- = data were not available
Source: NTP (2000).
In vitro studies by Ghanayem (1989) have shown that the hemolysis caused by the EGBE
metabolite BAA is preceded by erythrocyte swelling. If the observed increase in MCV is in
response to cell swelling, it could be a preliminary indicator of the hemolytic effect. Other
researchers, however, have attributed the increased MCV at all exposures and the increased
MCH at higher exposures to the erythropoietic response subsequent to hemolysis and the
corresponding increase in the number of larger reticulocytes in circulation (cited in NTP, 2000).
Reticulocyte counts were significantly increased in female rats at 62.5 ppm (6 and 12 months)
and in male rats at 125 ppm (3 and 6 months). Since a statistically significant increase in
reticulocyte count was not observed at any time point in males or females exposed to 31 ppm or
in males exposed to 62.5 ppm, it appears that reticulocyte count alone cannot account for the
increase in MCV. The observed increases in MCV may be a combined result of both erythrocyte
swelling prior to, and an increased number of reticulocytes subsequent to, hemolysis; the former
would be more influential at lower exposure levels, and the latter would have more relative
impact at higher levels.
Similar effects indicating anemia were also observed in mice, where females were more
sensitive. As in rats, the anemia response was observed at slightly higher doses, but was
persistent and did not progress or ameliorate in severity from 3 months to the final blood
collection at 12 months. Table 4-6 shows the manual Hct values in male and female rats and
mice at 3 months and 12 months. Statistically significant (p < 0.05) decreases in automated and
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manual Hct values and Hb and RBC counts occurred at 3, 6, and 12 months in the 125 ppm
females and the 250 ppm males and females. Statistically significant decreases in these
endpoints were also observed in 62.5 ppm females exposed for 6 months and in 125 ppm males
exposed for 6 and 12 months (decreases in Hct were observed only at 3 and 6 months). No
changes were observed in the MCV of mice, except for an increase in female mice at the highest
duration (12 months) and exposure (250 ppm) levels. Reticulocyte counts were increased
significantly in 125 ppm females at 3 and 6 months and in 125 ppm males at 6 months of
exposure.
At the end of the 2-year chronic bioassay (NTP, 2000), no significant neoplastic effects
were observed in male or female rats. In female rats, the combined incidence of benign and
malignant pheochromocytoma of the adrenal medulla was 3/50, 4/50, 1/49, and 8/49. The
incidence in the high-dose group (16%) does not represent a statistically significant increase over
the chamber control group (6%), but did exceed the historical control range (6.4 ± 3.5%; range
2-13%) for this effect.
Male mice exposed to 125 and 250 ppm EGBE had a low survival rate. A high rate of
hepatocellular carcinomas was found in these exposure groups (10/50 [control], 11/50, 16/49,
21/49); the increase at the high-exposure level was statistically significant (p < 0.01). When
hepatocellular adenomas and carcinomas are combined, no significant increase was observed in
any exposure group. However, the incidence of hemangiosarcomas in males exposed to
250 ppm (8%>) was significantly increased (p = 0.046) relative to chamber controls (0/50, 1/50,
2/49, 4/49) and exceeded the range of historical controls (14/968; 1.5 ± 1.5%; range 0-4%). No
organisms consistent with Helicobacter hepaticus were found in the 14 mice evaluated (NTP,
2000). The researchers concluded from this that H. hepaticus was not a factor in the
development of liver neoplasms. No significant increases in benign or malignant hepatocellular
tumors or hemangiosarcomas were noted in the female mice, and the incidence of hepatocellular
adenomas actually decreased significantly (p < 0.05) in relation to the control chamber group
(16/50, 8/50, 7/49, 8/49).
Forestomach squamous cell papillomas and carcinomas, combined, were significantly
increased (trend test = 0.003) in female mice relative to the chamber controls (0/50, 1/50, 2/50,
6/50). The incidence of these tumor types (12%) at the highest exposure level was statistically
significant and exceeded the range for the occurrence of these tumors in historical controls
(0.9 ± 1.1%; range 0-3%). The first incidence of these tumors appeared in the group exposed to
250 ppm at 582 days, compared with 731 days at 62.5 and 125 ppm. This indicates a decreased
latency period in the highest exposure group. While the incidence of these types of forestomach
tumors was not significantly increased over controls in male mice (1/50, 1/50, 2/50, 2/50), the
incidence of squamous cell papillomas (4%) in the two highest exposure groups exceeded the
range for historical controls (0.5 ± 0.9%; range 0-2%). The increased incidences of forestomach
neoplasms in males, as in females, occurred in groups with ulceration and hyperplasia.
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Section 4.6 has a discussion of the cancer data from this study. With respect to the
noncancer findings, a NOAEL could not be determined, and a LOAEL of 62.5 ppm was
determined in mice for hemosiderin deposition. In rats, a NOAEL could not be determined, and
a LOAEL of 31 ppm was determined for hemosiderin deposition.
4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND INHALATION
Due to the known reproductive toxicity, such as effects to male testes and sperm, of two
other glycol ethers, ethylene glycol methyl ether (EGME; 2-methoxyethanol) and ethylene glycol
ethyl ether (EGEE; 2-ethoxyethanol), the reproductive toxicity of EGBE was studied in a variety
of oral studies (NTP, 1993; Exon et al., 1991; Heindel et al., 1990; Foster et al., 1987; Grant
et al., 1985; Nagano et al., 1984, 1979) and inhalation studies (NTP, 2000; Dodd et al., 1983)
using rats, mice, and rabbits. Several developmental studies have addressed EGBE toxicity from
conception to sexual maturity, including toxicity to the embryo and fetus, following oral (Sleet
et al., 1989; Wier et al., 1987), inhalation (Nelson et al., 1984; Tyl et al., 1984), and dermal
(Hardin et al., 1984) exposures in rats, mice, and rabbits. In many instances, LOAELs and
NOAELs were reported for both parental and developmental effects; therefore, the
developmental studies can also be used to assess systemic toxicity as well as developmental
toxicity.
EGBE did not cause biologically significant effects in the reproductive organs, including
testes, in any study. In a two-generation reproductive toxicity study, fertility was reduced in
mice only at very high, maternally toxic doses (>1,000 mg/kg). Maternal toxicity, related to the
hematologic effects of EGBE, and relatively minor developmental effects have been reported in
developmental studies and are discussed below. No teratogenic effects were noted in any of the
studies. It can be concluded from these studies that EGBE is not significantly toxic to the
reproductive organs of adult males or females, or to the developing fetuses of laboratory animals.
As discussed in Section 4.2, Nagano et al. (1979) performed a toxicity study in male
JCL/ICR mice (five/group), using gavage doses of 0, 357, 714, or 1,430 mg/kg-day EGBE
5 days/week for 5 weeks. A LOAEL of 357 mg/kg-day based on reduced RBC count was
identified, but no changes in testes weight or histology were observed. In another study, Nagano
et al. (1984) used the same dosing regimen up to 2,000 mg/kg-day to test EGBE and other glycol
ethers. Testicular atrophy was observed for EGEE and EGME, but not for EGBE.
Grant et al. (1985) exposed male F344 rats (six/group) to gavage doses of 0, 500, or
1,000 mg/kg-day EGBE and EGME for 4 days. Severe testicular atrophy was observed in rats
fed 500 mg/kg-day EGME, but no significant effect was noted in rats fed up to 1,000 mg/kg-day
EGBE.
Krasavage (1986) conducted a toxicity study using groups of 10 COBS CD(Sprague-
Dawley)BR adult male rats treated by gavage with 222, 443, or 885 mg/kg-day undiluted EGBE
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5 days/week for 6 weeks. They found no effects on testicular weight and no histopathological
lesions in the testes, seminal vesicles, epididymides, or prostate gland at any exposure level.
Foster et al. (1987) fed Alpk/AP (Wistar-derived) male rats single gavage doses of 0, 174,
434, or 868 mg/kg BAA. Occasional significant decreases in the weight of the prostate gland
and seminal vesicles were observed, but the decreases were not time- nor dose-related.
No treatment-related lesions were noted following histologic examination of the testes,
epididymides, or prostate. BAA did not produce any changes in testicular cell populations when
introduced in vitro at 5 mM. Simultaneous testing of the acids of EGME and EGEE resulted in
significant spermatocyte cell loss and damage in vivo and in vitro.
Subchronic reproductive studies were conducted using male and female Swiss CD-I
mice. Heindel et al. (1990) exposed them to EGBE in drinking water at doses of 0, 700, 1,300,
and 2,000 mg/kg-day for 7 days premating; subsequently, they exposed the mice for 98 days
while cohoused as breeding pairs. The higher two dose levels resulted in mortality: 13/20 died
during the study in the 2,000 mg/kg-day group and 6/20 died in the 1,300 mg/kg-day dose group,
as compared with 1 each in the 700 mg/kg-day group and the control group. Statistically
significant toxic effects seen in the 1,300 and 2,000 mg/kg-day dose groups with adult mice
included decreased body weight gain, increased kidney and liver weights, and dose-related
decreases in water consumption. Statistically significant developmental effects observed in the
1,300 and 2,000 mg/kg-day dose groups included decreased pup weight and fewer and smaller
litters produced per pair. A significant reduction (5%) in live pup weight was also observed in
the 700 mg/kg-day dose group. No statistically significant effect on fertility was observed in the
700 mg/kg-day dose group.
At the completion of the 98-day continuous breeding phase, F0 breeding pairs were
separated and housed individually, while exposure to EGBE continued. When the last litter was
weaned, a 1-week crossover mating trial was performed to determine effects by gender.
F0 males and females from the 1,300 mg/kg-day dose group were mated with male and female
control animals. The exposed mice had significantly lower body weights and increased relative
kidney weights, but reproductive organ weights, sperm motility and morphology, and estrous
cycle length and frequency did not differ from control mice. In the only histopathological
examination carried out on treated females, no kidney lesions were observed. The proportion of
successful copulation was equivalent in all groups, and no developmental effects were observed
in any offspring. However, the number of fertile females was significantly reduced in the group
where treated females were mated with control males, suggesting that fertility effects were
primarily due to effects on the female mice.
A final phase of this study assessed the fertility and reproductive effects of EGBE in first-
generation (Fl) pups. There were insufficient numbers of offspring to assess the two highest
dose groups, and no statistically significant effect on fertility was noted when offspring of the
low, 700 mg/kg-day dose group were mated. Thus, the researchers concluded that the 700 and
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1,300 mg/kg-day dose levels are considered to be NOAEL and LOAEL values, respectively, for
both maternal and reproductive effects. A minimal LOAEL for developmental effects was
700 mg/kg-day, where a very slight decrease in pup weight was observed.
A study by Exon et al. (1991), also discussed in Section 4.4.5, looked at reproductive
parameters in male rats. Groups of six Sprague-Dawley rats were exposed to EGBE in drinking
water at doses of 0, 180, or 506 mg/kg-day (males) for 21 days. While testicular atrophy and
necrosis and a reduced number of spermatogenic cells were observed in males exposed to
EGME, no statistically significant effects on fertility parameters were seen in males at any of the
tested doses of EGBE.
NTP (1993) evaluated the effects of EGBE on the reproductive systems of male and
female B6C3Fi mice (five/gender/group) following 2-week drinking water exposure to doses of
93, 148, 210, 370, or 627 mg/kg-day for males and 150, 237, 406, 673, or 1,364 mg/kg-day for
females. No deaths were reported, and there were no effects on body weight. Thymus weights
were decreased in the highest male dose group. There were no treatment-related gross lesions in
any of the reproductive organs and histopathological examinations were not performed. NTP
(1993) also exposed male and female F344 rats (five/gender/group) to EGBE for 2 weeks in
drinking water. Male rats received doses of 73, 108, 174, 242, or 346 mg/kg-day, and females
received 77, 102, 152, 203, or 265 mg/kg-day. No treatment-related deaths occurred during the
study, and no changes in body weight were observed in male rats that could be related to
treatment. However, female rats had lower weight gain in the highest dose group. Water
consumption was lowest in the highest dose group in both genders, and no treatment-related
gross lesions of reproductive organs were reported.
Dodd et al. (1983) (also discussed in Section 4.2) performed a 90-day subchronic
inhalation study on F344 rats. Male and female rats (16/gender/group) were exposed to EGBE
for 6 hours/day, 5 days/week at concentrations of 0, 5, 25, and 77 ppm. They reported no
changes in testicular weight or in the pathology of the epididymides or testes of male rats at any
exposure level; reproductive organs of the female rats were not examined histologically.
NTP (2000) performed chronic and subchronic inhalation studies of EGBE in F344 rats
and B6C3Fi mice in which reproductive organs were examined. In the subchronic portion of the
NTP (2000) studies, rats and mice (10/gender/group) were exposed to concentrations of 0, 31,
62.5, 125, 250, and 500 ppm of EGBE 6 hours/day, 5 days/week for 14 weeks. Testicular
degeneration was reported in 2/4 mice from the 500 ppm group that died or were killed
moribund. No other effects were noted in the reproductive organs of rats or mice. Exposure
concentrations were 0, 31, 62.5, and 125 ppm for groups of 50 F344/N rats and 0, 62.5, 125, and
250 ppm for groups of 50 B6C3Fi mice. No effects were noted in the reproductive organs of
either species, but the researchers reported that survival was significantly decreased in male mice
at 125 and 250 ppm (54.0 and 53.1%, respectively).
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Prenatal and postnatal developmental toxicity tests were conducted in CD-I mice by Wier
et al. (1987). Animals received 0, 350, 650, 1,000, 1,500, or 2,000 mg/kg-day via gavage on
gestational days (GDs) 8-14. Maternal toxicity included mortality of 3/6 animals in the
1,500 mg/kg-day group and 6/6 in the 2,000 mg/kg-day group. Treatment-related clinical
observations were lethargy, abnormal breathing, and green or red vaginal discharge (the latter at
>1,500 mg/kg-day). Based on clinical signs in the prenatal study, the LOAEL for maternal
effects was 650 mg/kg-day. The LOAEL for developmental toxicity was determined to be
1,000 mg/kg-day based on a statistically significant increase in the number of resorptions and a
reduced number of viable fetuses. The corresponding NOAEL for prenatal effects was
350 mg/kg-day. In the postnatal study, reproductive effects were evaluated in offspring of
CD-I mice administered EGBE via gavage at 0, 650, or 1,000 mg/kg-day on GDs 8-14.
Maternal body weight was lowered at 1,000 mg/kg-day. Survival and body weight gain of
offspring were unaffected by treatment. No statistically significant developmental effects were
observed. In a simultaneous study with EGEE, developmental toxicity was noted at doses below
maternal toxicity levels.
Developmental toxicity was investigated following the administration of EGBE in
distilled water by gavage to groups of 28-35 pregnant F344 rats at doses of 0, 30, 100, or
200 mg/kg-day on GDs 9-11, or doses of 0, 30, 100, or 300 mg/kg-day on GDs 11-13 (Sleet
et al., 1989). GDs 9-13 were the most critical periods of fetal cardiovascular development.
Food and water measurements, body and organ weights, clinical signs, hematologic analyses of
dams, amount of corpora lutea, uterine contents, and number of dead and live fetuses were
monitored. Maternal effects of EGBE given in either dosing sequence included marked, dose-
related reductions in body weight and/or weight gain, increases in kidney and spleen weights,
severe hematotoxicity as evidenced by a decrease in HCT, Hb, and RBC counts, and an increase
in reticulocytes at doses >100 mg/kg-day. No indications of developmental toxicity were
observed at the two lower doses. Viability of embryos was reduced by EGBE treatment at the
200 mg/kg-day dose (GDs 9-11) but not at 300 mg/kg-day (GDs 9-13). A decreased platelet
count was noted in the fetuses at 300 mg/kg-day (GDs 9-13). Cardiovascular or other types of
malformations were not found at any dose. The LOAEL for maternal toxicity was 100 mg/kg-
day based on signs of hematotoxicity, with a NOAEL established at 30 mg/kg-day. The LOAEL
for developmental toxicity was 200 mg/kg-day based on decreased viability of embryos, with a
NOAEL for this endpoint at 100 mg/kg-day.
Sprague-Dawley rats (15/group) were exposed to 0, 150, or 200 ppm EGBE via
inhalation for 7 hours/day on GDs 7-15 (Nelson et al., 1984). Rats exposed to 200 ppm showed
some evidence of hematuria on the first day of exposure; no biologically significant effects were
noted thereafter, or at any time in offspring. The LOAEL was 200 ppm for slight maternal
toxicity; a NOAEL was identified at 150 ppm. The NOAEL for developmental toxicity was
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200 ppm. Simultaneous testing revealed that 50 ppm exposures to EGME were toxic at all levels
of embryonic and fetal development.
Pregnant F344 rats (36/group) and New Zealand white (NZW) rabbits (24/group) were
exposed to 0, 25, 50, 100, or 200 ppm EGBE via inhalation for 6 hours/day on GDs 6-15 for rats
or GDs 6-18 for rabbits (Tyl et al., 1984). Fetuses were weighed and evaluated for viability,
body weight, and morphological development, including external, visceral, and skeletal
malformations. In rats, fetotoxicity was observed at 100 and 200 ppm in the form of retarded
skeletal ossification of vertebral arches or centra, sternebrae, or phalanges. Maternal toxicity
was also evident at 100 and 200 ppm by increased incidence of hematuria, reduced RBC count,
decreased weight gain, and reduced food consumption. For both maternal and developmental
toxicity in the rat, the NOAEL and LOAEL values were 50 and 100 ppm, respectively. In
rabbits, fetal skeletal ossification of sternebrae and rudimentary ribs was delayed at 200 ppm.
Maternal toxicity was also evident at 200 ppm as an increased incidence of clinical signs,
reduced gravid uterine weight, and decreased weight gain. For both maternal and developmental
effects in the rabbit, the NOAEL and LOAEL values were 100 and 200 ppm, respectively.
Reproductive toxicity tests were performed on female Sprague-Dawley rats (number not
specified) via dermal administration during GDs 6-15, 4 times/day at 1,800 and 5,400 mg/kg-
day (Hardin et al., 1984). In the highest dose group, 10/11 rats died between days 3 and 7 of
treatment. Signs associated with treatment included red-stained urine, ataxia, inactivity, rough
coats, and necrosis of the tail tip. At the lower dose, body weight was slightly reduced, yet there
was no evidence of embryo- or fetotoxicity or gross malformations or variations.
4.4. OTHER STUDIES
4.4.1. Acute and Short-Term Exposure Studies
Ghanayem et al. (1987c) conducted acute toxicity studies designed to assess the effect of
age on toxicity by comparing effects in treated young (4-5 weeks old) and adult rats (9-
13 weeks, 5-6 months, and/or 16 months old). The researchers exposed male F344 rats
(six/group) using single gavage doses of EGBE in water (99% purity) to concentrations of 0, 32,
63, 125, 250, 250, or 500 mg/kg-day. Evaluations included total RBC and WBC counts, urine
Hb concentration, organ weights, and histology of the liver, spleen, bladder, kidney, and testes.
Focal necrosis of the liver was observed in adult rats exposed at a dose of 250 or 500 mg/kg.
Hematologic effects were found to be dose- and age-dependent, with older rats being more
sensitive than younger rats. Significant decreases in RBC counts, Hct, and Hb and increases in
free plasma Hb occurred at 125 mg/kg-day in both young and adult rats, with the younger rats
exhibiting significantly less pronounced responses. The incidence of hemoglobinuria was also
dose- and age-dependent. Concentrations of free Hb in urine were also significantly higher in
older rats than in younger rats at all doses. These researchers suggested that the metabolic basis
of the age-dependent toxicity of EGBE may be due to a reduced ability in older rats to
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metabolize the toxic metabolite BAA to CO2 and a diminished ability to excrete BAA in the
urine. Based on increased Hb concentrations in the urine and associated hemolytic effects at
higher doses, the LOAEL for this study was determined to be 32 mg/kg-day for both young and
adult rats. ANOAEL was not identified.
Ghanayem and Sullivan (1993) performed acute oral toxicity studies in male F344 rats
(N not specified), using single gavage doses of 250 mg/kg-day EGBE in tap water. MCV and
Hct values increased immediately after treatment and decreased with time following exposure.
Hemolysis and decreases in Hb concentrations and RBC counts were reported.
Grant et al. (1985) gavaged groups of 24 male F344 rats with EGBE (purity 99.9%) in
water at doses of 0, 500, or 1,000 mg/kg-day for 4 days. Six rats per dose were examined at 1, 4,
8, and 22 days after the last dose. The animals were evaluated for changes in body weight,
hematology, organ weight, and histology. Hematology evaluations showed marked dose-related
effects on circulating RBCs and WBCs. Changes at 500 and 1,000 mg/kg-day on postdosing
day 1 included significant dose-related decreases in Hb concentration and total WBC and
lymphocyte counts and increases in MCV, reticulocyte counts, and MCH. Hct was also reduced
at 1,000 mg/kg-day. Most of the RBC changes subsequently returned to normal, although MCV
and MCH remained increased at day 22. Body weight gain was sufficiently reduced throughout
the posttreatment period at 1,000 mg/kg-day. Changes in relative organ weights were evident on
posttreatment day 1, including increased liver and spleen weights at 500 and 1,000 mg/kg-day
and increased kidney and reduced thymus weights at 1,000 mg/kg-day. These changes returned
to normal by posttreatment day 22, except for liver and spleen weights at 1,000 mg/kg-day,
which increased somewhat (~5 and -20%, respectively). The authors determined that EGBE
appears to be relatively inactive as a bone marrow toxicant due to the observed proliferative
response and the lack of hemorrhage at any time in the bone marrow of EGBE treated animals.
Based on hemolytic anemia with associated reticulocytosis and increased hematopoiesis, a
LOAEL was established at 500 mg/kg-day, the lowest dose tested. ANOAEL was not identified.
Ghanayem et al. (1992) administered EGBE to male F344 rats (six/group) via gavage for
12 days at dose levels of 0 and 125 mg/kg-day. These investigators identified effects of EGBE
exposure similar to those identified above. Significant hemolysis occurred, becoming more
pronounced up to the third day of dosing. Gradual recovery was observed up to day 12. MCV,
ATP concentration, reticulocyte counts, and relative spleen-to-body weight ratios increased up to
the sixth day of dosing and declined thereafter. Liver-to-body-weight ratios were slightly
lowered on days 3 and 6 and slightly increased on day 12.
Several studies investigated EGBE-induced effects on specific organs and cells. Four
male and four female F344 rats were exposed to two, three, or four daily doses of EGBE at
250 mg/kg-day. Ezov et al. (2002) investigated hemolytic anemia and disseminated thrombosis
in rats by investigating the organs for hemolysis and histopathologic evidence of disseminated
thrombosis. Significant morphological changes in erythrocytes were noted in both genders of
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rats, while disseminated thrombosis and infarction were seen mainly in females and consisted of
tissue necrosis in the brain, liver, bones, eyes, lungs, and heart. Renal tubular necrosis associated
with Hb casts was seen in both genders. Koshkaryev et al. (2003) measured changes in
adherence, aggregability, and deformability of RBCs. EGBE exposure did not affect RBC
aggregability, and its effect on deformability was inconclusive; however, the exposure enhanced
RBC adherence to endothelial cells, with adherence highest at day 2 (the first day examined),
after which it decreased sharply with time. Shabat et al. (2004) studied bone marrow injury and
reported extensive vascular thrombosis resulting in necrosis of bone marrow cells, bone-lining
cells, and cortical and trabecular osteocytes. The authors concluded that, in EGBE-treated rats,
interactions of several factors may generate a thrombotic crisis, such as the release of
procoagulant factors from destroyed erythrocytes; they further concluded that disturbed blood
flow may result from alterations in the rheology of erythrocytes, including self-aggregation,
deformation, and adherence to the endothelium of the blood vessel wall. Redlich et al. (2004)
investigated the dental effects from EGBE-induced hemolysis and thrombosis. Odontoblastic
necrosis in the dental pulp of incisors and molars and muscle-cell damage in the tongue were
observed; the most severe changes occurred in females. These effects were probably the result
of ischemic events in the blood vessels supplying these tissues, rather than a direct cytotoxic
effect of EGBE.
Corley et al. (1999) conducted a series of studies in B6C3Fi mice investigating aspects of
EGBE toxicity, including the occurrence of forestomach lesions in both oral and inhalation
exposure routes, the dose-response of forestomach irritation, and the occurrence of forestomach
lesions as a consequence of systemic-only exposure. To determine the extent that activities
during inhalation exposures (e.g., grooming) could contribute to observed forestomach lesions,
groups of female mice were exposed for 6 hours to target concentrations of 250 ppm EGBE via
whole-body (n = 20) and nose-only (n = 20) exposures and concentrations on the fur that were
available for oral consumption via grooming measured. For whole-body exposures,
9.2 ± 2.9 mg/kg was available compared with 7.5 ± 2.3 mg/kg for the nose-only exposures.
Little difference was detected in the internal dose of EGBE from whole-body versus nose-only
exposures, as measured by the analysis of EGBE and in BAA detected in blood and urine
postexposure. To inform the dose response of toxicity in the forestomach tissues in mice, neat
EGBE was administered to male and female mice (five/gender/dose) via gavage (no vehicle) for
1 week at doses of 100, 400, or 800 mg/kg-day. The dose in the 100 mg/kg-day group was
increased to 1,200 mg/kg-day after 2 days. Severe hemolysis and mortality were seen, and the
2-week study was terminated after only four doses. Forestomach lesions consisting of focal
areas of irritation and epithelial hyperplasia were seen at all exposure levels in this study. Next,
the researchers administered saline solutions of EGBE to groups of three mice by either
intraperitoneal (i.p.) or subcutaneous (s.c.) injection at 400 or 600 mg/kg-day for 3 consecutive
days or 0 and 400 mg/kg-day for 5 consecutive days. Focal irritation in the forestomach, similar
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to that seen in gavage and inhalation studies, was seen in the three mice administered EGBE by
i.p. injection at 600 mg/kg for 3 days, while 1/3 mice at 400 mg/kg i.p. and at 400 and 600 mg/kg
s.c. for 3 days also had forestomach lesions, minimal effects. At 400 mg/kg (5-day study),
1/6 mice (i.p.) and 2/6 mice (s.c.) also had minimal lesions. From these results, it can be
concluded that: (1) the contribution to forestomach exposure from grooming activities, etc.,
during whole-body inhalation exposures is incidental; (2) the exposure of forestomach lesions to
EGBE is similar from gavage and inhalation exposures; and (3) forestomach tissues show a
similar irritative response whether EGBE exposure is systemic or portal of entry.
In another series of studies, male and female B6C3Fi mice (16/gender/group) were
exposed by gavage to 0, 400, 800, or 1,200 mg/kg-day of neat EGBE for 2 days (Poet et al.,
2003). A high level of mortality was seen; the dose was reduced by half and the dosing
discontinued when survival did not improve after two additional doses. Lesions, including
epithelial hyperplasia and inflammation of the forestomach, were seen at the higher dose levels
in both males and females, and minimal-to-mild forestomach epithelial hyperplasias were seen in
both genders of the lower-dose groups. In a study similar to that performed by Corley et al.
(1999), female mice were exposed, either by whole-body or nose-only inhalation, to a single
6-hour exposure of 250 ppm EGBE; the concentrations on the fur available for oral consumption
via grooming were measured. An average of 205 ± 69 jag of EGBE was detected on the fur of
the mice exposed whole-body, while an average of 170 ± 52 jag was detected on the fur of the
mice exposed nose-only (Poet et al., 2003).
Green et al. (2002) conducted a series of experiments to examine the effects and
distribution of EGBE in vivo. First, female B6C3Fi mice (five/group) were given a single, daily
gavage dose of either EGBE or BAA (0, 50, 150, or 500 mg/kg) for 10 days. Eighteen hours
after the last dose, animals were sacrificed to look at cell proliferation in the forestomach and
glandular stomach. The only dose-dependent finding was a thickening of the keratinized layers
or hyperkeratosis of the forestomach (statistics not reported). ANOAEL of 150 mg/kg EGBE
and 50 mg/kg BAA was observed. No effects were seen in the glandular stomachs at any dose
levels. In the next set of studies, female B6C3Fi mice (n = 12) were exposed whole-body to
250 ppm 2- butoxy[l-14C]ethanol (specific activity 0.365 mCi/mmol) for 6 hours. Following the
exposure, animals were given free access to food and water; they were terminated (four per time
point) at 5 minutes and 24 and 48 hours after exposure, and whole body autoradiography was
performed. Female B6C3Fi mice (n = 12) were given a single i.v. injection of 10 mg/kg
2-butoxy[l-14C]ethanol (850 (j,Ci/kg). The animals were then given free access to food and water
and terminated (four per time point) at 4, 24, and 48 hours after dosing, and whole body
autoradiograms were prepared. These studies showed that whether EGBE was delivered by
inhalation or by i.v., radiolabeled EGBE was found in the buccal cavity, esophagus, and stomach.
This suggests that EGBE somehow enters the stomach via the buccal cavity and esophagus
following inhalation exposure.
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Administration of a 2,000 mg/kg oral dose of EGBE to guinea pigs caused complete
mortality of females and 60% mortality of males (Shepard, 1994a, b), but a dose of 1,000 mg/kg
caused only a 20% mortality of either gender. Clinical signs and gross necropsy indicated
toxicity was due to irritation of the stomach. There was no evidence of hemolytic toxicity.
Gingell et al. (1998) performed acute oral and inhalation toxicity studies in the guinea
pig. A dose of 2,000 mg/kg EGBE was administered by gavage to five males and five females.
After excessive mortality (3/5 males, 5/5 females) was observed at this dose, reduced doses of
500 and 1,000 mg/kg were administered. No animals died at 500 mg/kg, and 1/5 males and
1/5 females died at 1,000 mg/kg. The acute oral median lethal dose for both genders was
1,414 mg/kg (95%) confidence interval [CI] = 1,020-1,961 mg/kg). Clinical signs in the guinea
pigs included slight-to-severe weakness, salivation and staining of face or abdomen hair, and
respiratory difficulties in a few males. No evidence of RBC toxicity or hemolysis was observed.
In the inhalation study, male and female guinea pigs were exposed for 1 hour to 633 ppm and
691 ppm, respectively. No mortality, clinical signs of toxicity, or exposure-related pathological
signs were noted. Thus, the median lethal concentration for a 1-hour exposure for guinea pigs
was >633 ppm in males and >691 ppm in females.
4.4.2. Dermal Exposure Studies
EGBE appears to be readily absorbed after contact with animal skin. Rats and rabbits
exhibit varying degrees of hematotoxicity following dermal application of EGBE (Allen, 1993 a,
b, c, d; Bartnik et al., 1987; Tyler, 1984). Bartnik et al. (1987) performed acute dermal toxicity
tests using Wistar rats (six/gender/group). A single application of 200, 260, 320, 375, or
500 mg/kg EGBE was placed on the dorsal shaved skin of rats and covered with a glass capsule.
Hemolytic and/or hemoglobinuria effects were observed at 500 mg/kg EGBE within 6 hours of
application. No effects were observed at 200 mg/kg.
Repeated occluded application of EGBE either neat or as a dilute aqueous solution to
NZW rabbits (five/gender/group) at 18, 90, 180, or 360 mg/kg (6 hours/day, nine consecutive
applications) produced hemoglobinuria in males at 360 mg/kg and in females at 180 or
360 mg/kg (Tyler, 1984). Only female rabbits showed decreased RBC counts, decreased Hb and
MCHC, and increased MCH concentrations at the highest treatment level. Recovery was noted
after 14 days. In a separate 13-week study, occluded dermal administration of EGBE to NZW
rabbits (10/gender/group) at exposure levels of 10, 50, or 150 mg/kg for 6 hours/day, 5 days/
week produced no observable hematological effects (Tyler, 1984).
Occlusion or semi-occlusion of the site of EGBE administration was also a determining
factor. For example, some studies have shown no clinical signs of hematotoxicity in Sprague-
Dawley rats (five/gender/group) administered EGBE dermally at 2,000 mg/kg (24-hour
exposure) either semi-occluded or occluded (Allen, 1993a, b). However, clinical signs of
systemic toxicity were noted following the occluded exposure. In similar studies in NZW rabbits
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(five/gender/group), red-stained urine was reported at semi-occluded doses of 2,000 mg/kg
EGBE, along with other clinical signs of systemic toxicity (Allen, 1993c, d). Similar effects
occurred at occluded doses of 500, 707, and 1,000 mg/kg in this species; deaths occurred at the
500 and 1,000 mg/kg exposures. Thus, hematotoxicity varied from nonexistent to severe. In
guinea pigs, dermal administration of EGBE at 2,000 mg/kg produced no deaths, clinical signs of
toxicity, or treatment-related signs of organ toxicity (Gingell et al., 1998; Shepard, 1994b).
In an assessment of immune parameters, female BALB/c mice (five/group) were
topically exposed to EGBE at 100, 500, 1,000, and 1,500 mg/kg-day for 4 consecutive days
(Singh et al., 2001). A statistically significant increase in spleen-to-body-weight ratio, and a 29%
increase in splenic cellularity was observed at 1,500 mg/kg. Splenic proliferative responses to
the T-cell mitogen, concanavalin-A(con-A), were significantly decreased by 32% at 500 mg/kg-
day and 35% at 1,000 mg/kg-day. Allogeneic antigen-driven lymphoproliferative responses in
the mixed lymphocyte response were significantly reduced by 55% at 500 mg/kg-day and 56% at
1,000 mg/kg-day. However, natural killer (NK) cell activity, cytotoxic T-lymphocyte activity,
and the T-dependent plaque-forming cell (PFC) response were not significantly affected by
EGBE exposure. A dose of 100 mg/kg-day was a NOAEL.
Singh et al. (2002) exposed female BALB/c mice (five/group) via gavage to 50, 150, or
400 mg/kg EGBE, or topically on the ear to 0.25, 1.0, 4.0, or 16.0 mg EGBE. The researchers
measured the oxazolone (OXA)-induced contact hypersensitivity response (CHR). Mice that
received the gavage doses of EGBE for 10 consecutive days did not exhibit a significantly
altered OXA-induced CHR as measured by ear swelling 24 hours postchallenge. In contrast,
topical exposure to EGBE significantly suppressed the OXA-induced CHR at a dose of 4.0 mg
EGBE/ear, but not at any other dose.
The studies indicate that while the dermal route of exposure can be expected to contribute
to overall exposure, the concentrations at which effects occur in animals are higher than those
found following oral and inhalation exposure. In humans, toxicokinetic studies have shown that
dermal absorption of EGBE vapors do contribute to the total body burden, showing the
importance of the dermal route of exposure.
4.4.3. Ocular Exposure Studies
EGBE has been found to be an irritant when instilled in rabbit (Jacobs and Marten, 1989;
Kennah et al., 1989). Kennah et al. (1989) performed the Draize eye irritation test in rabbits.
The percent EGBE concentration and corresponding scores by the Texaco single-digit toxicity
classification system were 100%—66, 70%—49, 30%—39, 20%—2, and 10%—1. In an
assessment that measured corneal thickness, the highest concentration was classified as severely
irritating, the 70% concentration was moderately irritating, and the others were mildly irritating.
Jacobs and Marten (1989) conducted ocular tests on NZW rabbits (n = 6) to determine the effects
of EGBE (100 [xL, 99% pure) on eye irritation. The undiluted chemical was dropped onto the
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lower lid of one eye; the other eye served as a control. The eyes were examined and graded for
ocular reactions at 4, 24, 48, 72, 96, and 168 hours postinstillation. The authors determined that
EGBE should be classified as an irritant based on the mean erythema scores and percent corneal
thickening.
4.4.4. Genotoxicity
Although weakly genotoxic responses have been obtained in two laboratories (Elias et al.,
1996; Hoflack et al., 1995), EGBE is not expected to be mutagenic or clastogenic based on the
available data (summarized in Table 4-7). The NTP reported negative responses for
mutagenicity when EGBE was tested in Salmonella typhimurium strains TA97, TA98, TA100,
TA1535, and TA1537 at up to 10 mg/plate with and without metabolic activation (Zeiger et al.,
1992). However, Hoflack et al. (1995) reported that at 38 [j,mol/plate (4.5 mg/plate), EGBE
induced a weak mutagenic response in salmonella tester strain TA97a in the absence of S9 mix
(Hoflack et al., 1995). The work of Hoflack and colleagues was repeated by Gollapudi et al.
(1996), and EGBE was found to be negative in these tester strains when evaluated at 0.5, 1.0,
2.5, 5.0, 8.5, and 10 mg/plate in the presence and absence of Aroclor-induced rat liver S9 mix.
Thus, the weak positive result reported in salmonella TA97a by Hoflack et al. (1995) is
unconfirmed. A plausible explanation put forth by Gollapudi et al. (1996) is that, given the
sensitivity of the Ames test, perhaps the weak positive result reported by Hoflack et al. (1995) is
attributed to an impurity in their test material.
Table 4-7. Summary of genotoxicity studies on EGBE, BAL, and BAA
Type of test, test species
Dose3
Result
Reference
In vitro tests: EGBE
Reverse mutation, S. typhimurium,
TA97, TA98, TA100, TA1535,
TA1537
10 mg/plate
Negative (w/ and w/o
metabolic activation)
Zeiger etal. (1992)
(work done for NTP)
Reverse mutation, S. typhimurium his-
TA 98, TA 100, TA 102
115 (imol/plate
(14.0 mg/plate)
Negative (w/ and w/o
metabolic activation)
Hoflack et al. (1995)
Reverse mutation, S. typhimurium his-
TA97a
38 (imol/plate
(4.5 mg/plate)
Weakly positive (w/o
metabolic activation)
Hoflack et al. (1995)
Reverse mutation, S. typhimurium his-
TA97a, TA 100; Escherichia coli
WP2uvrA
10 mg/plate
Negative (w/ and w/o
metabolic activation)
Gollapudi et al.
(1996)
Sister chromatid exchanges (SCEs),
micronuclei (MN) and aneuploidy (AP)
in V79 cells
10-100 mM (SCE)
8.46 mM (MN)
16.8 mM (AP)
Weakly positive (w/o
metabolic activation)
Elias etal. (1996)b
Potentiation of clastogenicity induced
by methyl methanesulfonate
8.5 mM
Positive (w/o metabolic
activation)
Elias etal. (1996)b
Chromosomal aberrations, V79 cells
and human lymphocytes
Not available
Negative (w/o metabolic
activation)
Elias etal. (1996)b
Gene mutation, Chinese hamster ovary
cells
38.1 mM°
(4.5 mg/mL)
Negative (w/o metabolic
activation)
Chiewchanwit and Au
(1995)
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Table 4-7. Summary of genotoxicity studies on EGBE, BAL, and BAA
Type of test, test species
Dose3
Result
Reference
DNA damage, SVEC4-10 mouse
endothelial cells
10 mM
Negative
Klaunig and
Kamendulis (2005)
In vitro tests: BAL
Reverse mutation, S. typhimurium his-
TA 97a, TA 98, TA 100 and TA 102
43 (imol/plate
(5.0 mg/plate)
Negative (w/ and w/o
metabolic activation)
Hoflack et al. (1995)
Chromosomal aberrations, V79 cells
and human lymphocytes
0.1-1 mM;
cytotoxicity not
reported
Positive (w/o metabolic
activation)
Elias et al. (1996)b
DNA damage, SVEC4-10 mouse
endothelial cells
1 mM
Negative
Klaunig and
Kamendulis (2005)
In vitro tests: BAA
Reverse mutation, S. typhimurium his-
TA 97a, TA 98, TA 100 and TA 102
8 (imol/plate
(0.9 mg/plate)
Negative (w/ and w/o
metabolic activation); dose
limited by toxicity
Hoflack et al. (1995)
SCEs and CAs, V79 cells
0.8 mM
Negative (w/o metabolic
activation)
Elias et al. (1996)b
Aneuploidy, V79 cells
0.38 mM
Weakly positive (w/o
metabolic activation)
Elias et al. (1996)b
MN assay, V79 cells
10 mM
Positive (w/o metabolic
activation
Elias et al. (1996)b
DNA damage, SVEC4-10 mouse
endothelial cells
10 mM
Negative
Klaunig and
Kamendulis (2005)
In vivo tests: EGBE
MNs, bone marrow erythrocytes of
male mice or rats
550 mg/kg-d, mice
450 mg/kg-d, rats
Negative
Negative
NTP (1996)
DNA adducts FVB/N mice
Sprague-Dawley rats
120 mg/kg-d, mice
and rats
No changes in DNA
methylation
Keith et al. (1996)
"Doses are either the lowest effective dose or the highest ineffective dose.
bAll in vitro assays were performed without the addition of an exogenous metabolic activation system.
The authors found that this dose was cytotoxic.
Elias et al. (1996) reported that EGBE did not induce chromosomal aberrations in
Chinese hamster V79 fibroblast cells but that EGBE, at treatment concentrations of >8.5 mM,
weakly induced sister chromatid exchanges (SCEs) and micronuclei and potentiated the
clastogenicity induced by methyl methanesulfonate. Elias et al. (1996) also reported that EGBE
weakly induced aneuploidy (numerical chromosomal anomalies) in V79 cells; however, this
response was found at very high concentrations (16.8 mM EGBE).
When tested at doses nearing toxicity, EGBE and its metabolite BAL were not mutagenic
in an in vitro gene mutation assay using Chinese hamster ovary (CHO) cells (CHO-AS52)
(Chiewchanwit and Au, 1995). In contrast, Elias et al. (1996) reported that both EGBE and BAL
weakly induced gene mutations in Chinese hamster V79 cells only at high treatment
concentrations (>7.5 mg/mL). It should be noted that Chiewchanwit and Au (1995) reported
high cytotoxicity at 38.1 mM EGBE (4.5 mg/mL). The gene mutation data presented by Elias et
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al. (1996) is in graphic form only with mean values and no SDs presented. The presence or
absence of cytotoxicity was not reported. BAL was also tested for induction of DNA damage in
the mouse endothelial cell line, SVEC4-10, using the comet assay. BAL failed to produce a
statistically significant increase in DNA strand breaks at any of the concentrations or time points
examined (Klaunig and Kamendulis, 2005, 2004; Reed et al., 2003). Other lines of evidence
indicate that direct interaction of BAL with the DNA molecules does not play a significant role
in the carcinogenic activity of EGBE. First, BAL causes cytotoxicity at levels associated with
chromosome effects, and cytotoxicity itself can have effects that result in chromosome damage,
such as reduction in the repair of SCEs. Second, acetaldehyde is recognized as "weakly
mutagenic" and structural comparisons of the aldehyde metabolites of glycol ethers shows that
longer-chain aldehydes such as BAL are less mutagenic (Chiewchanwit and Au, 1995). Third, if
BAL were a stable mutagenic metabolite in any of the in vitro assays exposed to EGBE, one
would expect them to give positive results; however, the results were generally negative. Elias et
al. (1996) suggested that the V79 cells possess neither ALDH nor ADH. The relevance of these
studies, or of any systems that lack these enzymes, is of limited value in elucidating the MOA of
toxicity in biological systems that possess these enzymes. BAA has been found negative for
reverse mutations in S. typhimurium his" with and without metabolic activation (Hoflack et al.,
1995). Concentrations of up to 8 [j,mol/plate were tested, and dose was limited by toxicity. BAA
(up to 10 mM) was also found negative for induction of DNA damage in SVEC4-10 mouse
endothelial cells (Klaunig and Kamendulis, 2005) and in an SCE assay in V79 cells (Elias et al.,
1996). BAA was weakly positive for aneuploidy in V79 cells at 0.38 mM and positive for
micronuclei induction in the same cell line at 10 mM, as reported by Elias et al. (1996). As noted
above, the data means are presented in graphic form without SDs and cannot be critically
evaluated; no cytotoxicity data are reported.
EGBE did not increase the incidence of micronuclei in the bone marrow cells of male
mice or rats (NTP, 1996). Animals were given three i.p. injections of EGBE 24 hours apart and
sacrificed 24 hours after the last injection; rats were dosed at 0, 7, 14, 28, 56, 112.5, 225, or
450 mg/kg and mice were dosed at 0, 17, 34, 69, 137.5, 275, or 550 mg/kg (NTP, 1996). There
was high mortality (2/5 mice survived) in mice injected with 1,000 mg/kg doses of EGBE. Keith
et al. (1996) treated Sprague-Dawley rats and transgenic FVB/N mice carrying the v-Ha-ras
oncogene with a single oral dose of 120 mg/kg EGBE; there was no increase in DNA adducts in
the brain, liver, kidney, testes, or spleen of the rats, and no changes in DNA methylation patterns
in either species.
In conclusion, EGBE has been tested in conventional genotoxicity tests for its potential to
induce gene mutations in systems and cytogenetic damage both in vitro and in vivo. Available
data do not support a mutagenic or clastogenic mechanism for EGBE. Two laboratories (Elias
et al., 1996; Hoflack et al., 1995) reported weak genotoxicity responses in vitro at toxic doses.
These results, however, are questionable given limited published information. Elliott and Ashby
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(1997) reviewed the results of the available genotoxicity studies on EGBE and concluded that the
data indicate that EGBE has no significant genotoxic activity.
4.4.5. Immunotoxicity
Based on the results of the Exon et al. (1991) study, it appears that the immune system is
not a sensitive target of EGBE. Groups of six Sprague-Dawley rats were exposed to EGBE in
drinking water at doses of 0, 180, or 506 mg/kg-day (males) or 0, 204, or 444 mg/kg-day
(females) for 21 days. All rats were injected s.c. with heat-aggregated aqueous keyhole limpet
hemocyanin (KLH), a T-cell dependent antigen, on days 7 and 13 following the start of dosing.
Endpoints evaluated on day 21 included body weight, absolute and relative organ weights
(spleen, thymus, liver, kidney, testis), and histology of thymus, liver, kidney, and testis. Splenic
histology was not assessed because this tissue was used as a source of cells for immune function
assays. Immune function assays included NK cell cytotoxicity, antibody response to a T-cell
dependent antigen as determined by measurement of KLH-specific serum immunoglobulin G
levels, delayed-type hypersensitivity reaction, interleukin-2 and interferon production, and spleen
cell counts. Terminal body weights were somewhat lower than controls in all exposed groups,
and the decreases were statistically significant in all groups except the 180 mg/kg-day males. No
dose-related changes in organ weights or histology were observed. NK cell cytotoxic responses
were significantly enhanced in males at 180 mg/kg-day and females at 204 mg/kg-day. At the
highest dose tested in males and females, the NK cell cytotoxic response was increased above
control, but this finding was not statistically significant. A decreased NK cell cytotoxic response
is an indication of compromised nonspecific immune system integrity. Given that this study
showed an increased response and no dose-response relationship, these findings are not
considered to be an indication of an adverse response. No significant alterations in other
immune parameters were noted.
Smialowicz et al. (1992a) reported on results of experiments that examined the primary
antibody response to a T-cell independent antigen (trinitrophenyl-lipopolysaccharide [TNP-LPS])
measured with a PFC assay to determine the immunotoxic potential of a variety of glycol ethers.
This author had previously reported that the shorter-chain glycol ether 2-methoxyethanol (ME)
and its principal metabolite 2-methoxyacetic acid (MAA) suppressed the antibody response to
TNP-LPS as measured by the PFC assay in F344 rats but not CD-I mice (Smialowicz et al.,
1992b). Having established the sensitivity of F344 rats to suppression of the antibody response
to the T-cell independent antigen TNP-LPS by ME and MAA, the authors examined other glycol
ethers, including 2-butoxyethanol, in the same dose range, for their ability to suppress the
antibody response to TNP-LPS by using the same PFC assay. Male F344 rats were immunized
with a single i.v. injection of 0.5 mL of 40 [j,g/mL TNP-LPS, then dosed (six/dose group) by
gavage with 50-400 mg/kg-day of various glycol ethers, including EGBE (0, 50, 100, 200,
400 mg/kg-day) for 2 days. All rats exposed to 400 mg/kg-day EGBE died, and the 200 mg/kg-
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day EGBE dose resulted in one dead and one moribund rat. This finding was not unexpected, as
the hematotoxicity of EGBE in older rats has been reported in the literature (Ghanayem et al.,
1987c; Tyler, 1984; Carpenter et al., 1956). EGBE did not suppress the primary antibody
response to TNP-LPS in the PFC assay.
In an assessment of immune parameters, female BALB/c mice (five/group) were
topically exposed to EGBE at 100, 500, 1,000, and 1,500 mg/kg-day for 4 consecutive days
(Singh et al., 2001). A statistically significant increase in spleen-to-body-weight ratio was
observed at 1,500 mg/kg, and splenic cellularity was increased by 29% at this dose. Splenic
proliferative responses to the T-cell mitogen, con-A, were significantly decreased by 32% at
500 mg/kg-day and 35% at 1,000 mg/kg-day. Allogeneic antigen-driven lymphoproliferative
responses in the mixed lymphocyte response were significantly reduced by 55% at 500 mg/kg-
day and 56%) at 1,000 mg/kg-day. However, NK cell activity, cytotoxic T-lymphocyte activity,
and the T-dependent PFC response were not significantly affected by EGBE exposure. For those
immune parameters measured, 100 mg/kg-day was aNOAEL.
Singh et al. (2002) exposed female BALB/c mice (five/group) via gavage to 50, 150, or
400 mg/kg EGBE or topically to 0.25, 1.0, 4.0, or 16.0 mg EGBE on the ear and measured the
OXA-induced CHR. Mice that received the gavage doses of EGBE for 10 consecutive days did
not exhibit a significantly altered OXA-induced CHR as measured by ear swelling 24 hours
postchallenge. In contrast, topical exposure to EGBE significantly suppressed the OXA-induced
CHR at a dose of 4.0 mg EGBE/ear, but not at any other dose.
4.4.6. Other In Vitro Studies
Ghanayem (1989) compared the metabolic and cellular basis of EGBE-induced
hemolysis of rat and human erythrocytes in vitro. EGBE is not metabolized when incubated with
blood from male F344 rats and causes no hemolysis or metabolic alterations at concentrations of
up to 10 mM. A concentration of 20 mM EGBE was required to produce significant hemolysis
of rat blood. This may be due to a nonspecific effect occurring at a concentration that is not
physiologically relevant. In contrast, incubation of rat blood with BAL or BAA at concentrations
of 0.5, 1.0, or 2.0 mM caused a time- and concentration-dependent increase in cell swelling (i.e.,
increased Hct) followed by hemolysis. This response was more pronounced for BAA, with
nearly complete hemolysis observed after a 4-hour incubation at 2.0 mM. BAL produced only
slight hemolysis under the same conditions. The addition of ADH (with nicotine adenine
dinucleotide cofactor) to rat blood followed by BAL produced a potentiation of the hemolytic
effects. Addition of cyanamide, an ADH inhibitor, significantly decreased the effects with or
without added ADH. Both BAL and BAA caused a time- and concentration-dependent decrease
in blood ATP concentrations, although this effect may be secondary to the swelling and lysis
observed. Addition of exogenous ATP failed to reverse the hemolytic effects. Neither EGBE nor
its metabolites, BAL and BAA, caused any detectable changes in the concentrations of
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glutathione (GSH) or glucose-6-phosphate dehydrogenase (G6PD) in rat erythrocytes. Blood
from male and female human volunteers was unaffected by 4-hour incubations with BAA at
concentrations of up to 4.0 mM. At 8 mM, slight but significant hemolysis of human blood was
observed: blood from female volunteers showed a slightly greater sensitivity. These studies
show that the erythrocyte membrane is the likely target for BAA, and that humans of both
genders are relatively insensitive to the hemolytic effects of BAA, as compared with rats.
Udden (2000) compared effects on RBC morphology in male F344 rats (five/group)
dosed with 125 or 250 mg/kg EGBE via gavage with the effects of incubation of rat erythrocytes
in vitro with BAA (1 or 2 mM). In vivo exposure resulted in stomatoacytosis and spherocytosis
in blood smears and cup-shaped cells and spherocytosis in fixed samples. In vitro incubation
resulted in erythrocytes with cup shapes and spherocytosis in the fixed samples. Since in vivo
and in vitro exposure caused similar changes in RBC morphology in rats, this study provides
additional evidence for the role of BAA in causing the hemolytic effects from EGBE exposure
in rats.
Udden (2002) performed in vitro comparisons of sub-hemolytic and hemolytic effects in
rat and human RBCs in response to exposure of high BAA, using samples from hospitalized
adults, hospitalized children, and well adults. Erythrocyte parameters measured included
deformability, size distribution, density, MCV, count, osmotic fragility, and hemolysis. Exposure
for 4 hours resulted in loss of deformability, although at concentrations 150-fold different; the
first significant alterations noted in cells exposed was at 0.05 mM BAA in rats and at 7.5 mM
BAA in humans. Significant increases in MCV after a 4-hour exposure to BAA were noted at
0.05 mM in rats and at 10.0 mM in humans, a 200-fold difference. A similar differential in
concentration range was noted in osmotic fragility. Testing of erythrocytes for changes in MCV
and percent hemolysis from in vitro exposure to either 0 or 10 mM BAA showed no differences
among the well adults or the hospitalized children. The percent hemolysis seen in the average
response of the hospitalized adults to 10 mM BAA was significantly increased, from 0.54 to
0.69%. Significant increases in MCV were noted for all three of these groups, including the well
adults. The mechanisms underlying erythrocyte damage and the resistance of human cells to
these effects remain unknown. It should be noted, however, that although the sub-hemolytic
responses were similar for human and rat erythrocytes, other aspects of the erythrocytes,
including morphology and cell density, were altered in the rats but not in human cells. These
differences between species suggest that the mechanisms underlying the observed prehemolytic
effects may be different between rats and humans.
The possibility exists that certain human subpopulations, including the aged and those
predisposed to hemolytic disorders, might be at increased risk from EGBE exposure. Udden
(1995b, 1994) investigated this possibility using blood from the elderly (mean age 71.9; range
64-79 years; five men and four women), from seven patients with sickle cell disease, and from
four subjects with hereditary spherocytosis, three of whom were studied postsplenectomy and
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one studied pre splenectomy. Using a sensitive assay for erythrocyte deformability (Udden,
1994; Udden and Patton, 1994), it was shown that blood from all of these potentially sensitive
groups was unaffected by incubations of up to 4 hours with 2 mM BAA.
Udden and Patton (2005) examined the role of osmolarity and cation composition of the
cell suspension buffers in the mechanism of BAA-induced hemolysis of rat RBCs. Adding
sucrose to the cell suspension media or replacing external sodium with potassium protected rat
RBCs from BAA-induced hemolysis. The authors also observed that adding as little as 0.05 mM
CaCh to the buffer delayed the time course of the hemolytic response, while adding MgCh had
no effect. Use of the calcium-activated potassium channel inhibitor charybdotoxin blocked the
protective effect of calcium. From this, the authors suggest that BAA causes sodium and calcium
to enter the cell. While calcium initially has a protective effect via a loss of potassium through
the calcium-activated potassium channel, compensating for the osmotic effect of increased cell
sodium, calcium may subsequently have other deleterious effects through activation of proteases
and other calcium-activated processes.
Corthals et al. (2006) tested EGBE, BAL, and BAA in the comet assay to determine their
ability to induce DNA damage in SVEC4-10 mouse endothelial cells. EGBE (up to 10 mM),
BAL (up to 1 mM), and BAA (up to 10 mM) did not produce significant increases in DNA
damage relative to controls at any of the time points examined (2, 4, and 24 hours). The
researchers next tested the effect of hemolyzed mouse RBC lysate and ferrous sulfate in the same
system and found that the hemolyzed RBCs produced a statistically significant increase in DNA
damage at the highest concentration tested (10 x 106 hemolyzed RBCs) for 4 hours. No other
time points were significant. Ferrous sulfate produced statistically significant increases in DNA
damage at the highest time point and the lowest concentration tested (24 hours, 0.1 |iM) and at
all time points (2, 4, and 24 hours) in the mid- and high doses tested (0.5 [iM and 1.0 [xM). The
next experiment examined the ability of EGBE, BAA, ferrous sulfate, and hemolyzed RBCs to
stimulate tumor necrosis factor-alpha (TNFa)D release from cultured mouse macrophages
(RAW 264.7 cells). Hemolyzed RBCs (10 x 106 cells) resulted in a statistically significant
increase (p < 0.05) in TNFa release following a 4-hour treatment. Treatment with EGBE, BAA,
or ferrous sulfate did not result in increased TNFa release. Finally, the authors report that
macrophages activated with hemolyzed RBCs (10 x 106) for 4 hours were able to increase DNA
synthesis in mouse endothelial cells through co-culturing for 24 hours. These macrophages were
not, however, able to increase endothelial cell DNA damage (after 4- or 24-hour treatment) as
measured by the comet assay. The authors did find that LPS activated macrophages, after a
4-hour treatment, produced statistically significant increases (p < 0.05) in endothelial cell DNA
damage, as measured by the comet assay.
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4.5. SYNTHESIS AND EVALUATION OF MAJOR NONCANCER EFFECTS AND
MODE OF ACTION: ORAL AND INHALATION
Hemolytic anemia is the primary response elicited in sensitive species following
inhalation, oral, or dermal administration of EGBE. The carboxylic acid metabolite of EGBE,
BAA, has been shown to be the causative agent in this hemolysis. The mechanisms underlying
the hemolytic events are unknown. In vitro tests have shown that BAA produces a
concentration- and time-dependent swelling of rat erythrocytes, and changes the normal
erythrocyte morphology from the typical discocyte form to a spherocyte form prior to lysis. One
potential mechanism that could explain the RBC effects of BAA is the activation of membrane
phospholipid scrambleases. The outer leaflet of the RBC membrane normally has a net positive
charge due to the abundance of phosphatidylcholine while the cytoplasmic side contains
phosphatidylserine. Activation of phospholipid scrambleases results in the translocation of
phosphatidylserine to the outer leaf while shuffling phosphatidylcholine to the cytoplasmic side
of the membrane. The externalization of the phosphatidylserine changes the outer membrane net
charge to negative. The perturbations in membrane homeostasis may result in a loss of
membrane integrity leading to a loss of deformability of the erythrocyte. The fate of the
damaged red cells, whether by direct lysis in the circulation, or splenic sequestration, in
unknown. It is likely that under low EGBE exposure conditions, splenic sequestration will
predominate but at higher exposure conditions, splenic spillage and / /or frank intravascular lysis
will occur. Heme transport mechanisms would then be overloaded and iron containing fragments
will accumulate in the phagocytic cells of the liver (Kupffer cells) as hemosiderin. It is
interesting to note that phosphatidylserine externalization has been associated with impending
apoptosis in nucleated cells. Older erythrocytes are more sensitive to the hemolytic effects of
BAA than are younger cells or newly formed reticulocytes. Hemolysis can be induced in vivo
following administration of EGBE or in vitro following addition of BAA to either whole blood
or washed erythrocytes.
Additionally, it has been reported that EGBE exposure to rats caused disseminated
thrombosis and infarction (Ramot et al., 2007; Yoshizawa et al., 2005; Ezov et al., 2002; Nyska
et al., 1999a, b). The mechanism by which this occurs is presently unknown but several
possibilities exist. The changes in erythrocyte morphology and decrease in deformability could
result in intravascular occlusion. Secondly, the lysis of the erythrocyte would result in release of
procoagulants. Finally, the thrombotic response could be the result of the appearance of anionic
phospholipids, particularly phosphatidylserine, on the cell surface activating the prothrombinase
complex resulting in the formation of thrombin (Connor et al., 1989; Bevers et al., 1982).
Nevertheless, thrombosis only occurs at higher EGBE exposure levels and is not the most
sensitive endpoint.
The primary response of hemolysis is indicated via dose-related clinical observations of
decreases in Hct, Hb concentration, and RBC count in the blood of laboratory animals exposed
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to EGBE. The hemolysis-related events of macrocytosis and increased MCV were also observed
in the rat, considered to be a sensitive species, and are attributed at least partly to the increased
number of larger reticulocytes in the circulation following the erythropoietic compensatory
response (NTP, 2000). These alterations were persistent throughout the chronic animal
exposures but do not appear to progress with extended exposure (from 3 months to 1 year for
rats). These changes persist despite functioning, compensatory, homeostatic mechanisms.
Liver effects were noted in the NTP (2000) reports of subchronic and chronic inhalation
studies in rats and mice and in the subchronic drinking water study in rats. These included
statistically significant increases in the iron-staining hepatic pigmentation attributed to
hemosiderin, the storage protein for insoluble iron, believed to be derived from the Hb released
during hemolysis. Nyska et al. (2004) examined the possible association between chemically
induced hemosiderin deposition and hemangiosarcomas in the liver of mice from 130 NTP
bioassays, and found a highly significant association between liver hemangiosarcoma and
apparent Kupffer cell pigmentation (see Roberts et al. [2007] for an overview of the role of
Kupffer cells in hepatotoxicity and carcinogenicity). The cause for the hemosiderin deposition in
all cases was the erythrocyte hemolytic effect of the compounds. However, in an NTP (1993)
subchronic drinking water study, hepatocellular cytoplasmic changes were observed in male rats
at an exposure level (750 ppm) below the level at which hematological changes were recorded
(1,500 ppm). This finding raises the possibility of a direct, primary hepatic toxicity due to either
EGBE or an EGBE metabolite. Similar liver effects observed in female rats at the 750 ppm
exposure level were accompanied by hematological effects.
In the same NTP (1993) report, no liver lesions were reported in mice exposed to
drinking water containing up to 6,000 ppm EGBE. The lesions reported in rats consisted of
cytoplasmic alterations, hepatocellular degeneration, and pigmentation. The cytoplasmic
alterations, the only lesion observed at the 750 ppm exposure level (corresponding to a
consumption of roughly 55 mg/kg-day EGBE for adult male rats), were described as hepatocytes
staining more eosinophilic and lacking the amphophilic-to-basophilic granularity of the
cytoplasm present in hepatocytes from control animals. Greaves (2000) suggested that the lack
of cytoplasmic granularity or ground-glass appearance of the hepatocytes is an indication that the
response does not involve enzyme induction. The hepatocellular degeneration and pigmentation
observed at the higher exposure levels in both genders was centrilobular, which is consistent with
the Kupffer cell pigmentation and hemosiderin deposition reported in the NTP (2000) inhalation
studies. This information, along with the observation that all other rat and mouse oral and
inhalation studies of EGBE report hemolysis at or below exposure levels that result in liver
effects, suggests that at least these cytoplasmic hepatocellular changes in male rats reported in
the NTP (1993) drinking water study may reflect adaptation to a subclinical level of hemolysis.
However, focal necrosis of the liver observed in male rats following gavage administration of
250 and 500 mg/kg EGBE (Ghanayem et al., 1987b) was judged to be inconsistent with typical
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anoxic centrilobular necrosis associated with anemia (Edmonson and Peters, 1985). The effects
observed in the Ghanayem et al. (1987b) study may be associated with the high bolus exposures
employed.
The liver alterations documented throughout these studies suggest a defined progression
of pathological events with increasing doses of EGBE with increasing levels of hemolysis. In
particular, hepatic hemosiderin deposition in the liver is a dose-related sequela of the hemolytic
activity caused by EGBE exposure. This deposition was noted to follow a dose-response
relationship as well as to increase in severity in the chronic rat and mouse NTP studies; it shows
a statistically significant increase relative to the chamber controls. Although some hemosiderin
deposition in the spleen and liver may be expected with increasing age, the extent of hemosiderin
deposits noted by NTP in the livers of EGBE exposed animals is not normal. For these reasons,
hemosiderin deposition in the liver has been considered a pathological finding (Muller et al.,
2006). The following issues relate to the relevance of these hemolytic and hepatic effects to
humans and to the MO A of EGBE.
The weight of evidence from a variety of studies in animals and humans suggests that
certain species are more susceptible to the hemolytic effects of EGBE. It appears that humans
are less sensitive to the hemolytic effects of EGBE than rats and mice. On one end of the
sensitivity range is the guinea pig, which displays no hemolytic effects from EGBE at exposure
levels as high as an oral dose of 1,000 mg/kg or a dermal dose of 2,000 mg/kg. The rat,
conversely, displays increased osmotic fragility of erythrocytes at single inhalation exposures
below 100 ppm and single oral exposures below 100 mg/kg EGBE. No hemolysis was observed
in controlled laboratory acute inhalation exposures of human volunteers at up to 195 ppm EGBE.
Some reversible hemolytic effects have been described in addition to more debilitating effects in
humans who consumed single oral doses of 400-1,500 mg/kg EGBE in cleaning formulations
(see Section 4.1). Effects in humans from chronic exposure to EGBE have not been studied.
With respect to gender sensitivity, it has been consistently noted (Ezov et al., 2002; NTP,
2000, 1993; Dodd et al., 1983; Carpenter et al., 1956) that female rats are more sensitive to
EGBE-induced hemolysis than males. This gender difference is consistent with toxicokinetic
data for male and female rats reported by the NTP (2000) 2-year study. Female rats eliminated
BAA, the toxic metabolite of EGBE, more slowly from the blood, resulting in a larger AUC for
the blood concentration of BAA versus time. This may be a result of the reduced renal excretion
observed in female versus male rats. NTP (2000) also reported that, like female rats, female
mice tended to have greater blood concentrations of BAA at any given time than males. This
may explain the slight increase in incidence and severity of the anemic response found in female
mice, as compared to males. However, unlike female rats, female mice excrete slightly more
BAA than male mice; no significant difference between female and male mice has been noted in
the overall rate of elimination or the ty2 of BAA.
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Some studies (Ghanayem et al., 1990, 1987c) were designed to assess the effect of age on
the toxicokinetics and hemolytic effects in young and adult rats treated with single EGBE gavage
exposures. Both blood retention (Ghanayem et al., 1990) and hematologic effects (Ghanayem
et al., 1987c) were found to be dose- and age-dependent; older rats retained more of the EGBE
metabolite BAA in their blood and were more sensitive than younger rats. The increased blood
retention of BAA (as measured by increased Cmax, AUC, and tu) in older rats versus younger rats
may be due to metabolic differences or compromised renal clearance. The researchers suggested
that the pharmacokinetic basis of the age-dependent toxicity of EGBE may be due to a reduced
ability of older rats to metabolize the toxic metabolite BAA to CO2, and a diminished ability to
excrete BAA in the urine. No comparable studies exist for an analysis of liver effects.
While older rats appear to be more severely affected by acute doses of EGBE, continuous
exposures appear to impart a certain level of tolerance to rats and mice over time. Apparent
tolerance to EGBE-induced hemolysis in rats and mice has been seen in subchronic (Krasavage,
1986; Grant et al., 1985) and chronic (NTP, 2000) studies. Ghanayem et al. (1992, 1990)
investigated this effect in the male F344 rat. Daily gavage administration of EGBE at 125 mg/kg
for 12 days resulted initially in hemolytic anemia, which was more pronounced following the
third day, but the animals recovered gradually to near pretreatment levels by day 12.
Additionally, rats treated for 3 days at 125 mg/kg followed by a 7-day recovery period were
significantly less sensitive to subsequent treatment with EGBE at either 125 or 250 mg/kg, as
were rats that were bled and subsequently treated. Ghanayem et al. (1992, 1990) proposed that
the tolerance to hemolysis following repetitive dosing is not due to changes in EGBE
metabolism, but to the replacement of older and more susceptible erythrocytes with younger, less
susceptible cells. However, chronic studies in rats and mice (NTP, 2000) have shown that any
increased resistance imparted by these immature erythrocytes diminishes with age. Rats and
mice chronically exposed to EGBE experienced anemia that persisted with no apparent
progression or amelioration of severity for 9 months, up to the final blood collection at
12 months of age. There may be a balance in these rodents between the release of reticulocytes
to the circulation and the aging process, so that the level of susceptible cells and severity of
anemia remains relatively constant.
A number of secondary effects resulting from the hemolytic toxicity of EGBE have been
reported in studies with rats, mice, and rabbits. In the rat, the organs generally affected include
most prominently the liver (see discussions above) but also the kidneys, spleen, bone marrow,
and, to a lesser extent, the thymus (Shabat et al., 2004; NTP, 1993; Exon et al., 1991; Grant et al.,
1985). Typically, increased liver and kidney weights are observed with corresponding decreases
in body weights at doses that produce a hematotoxic response. Accompanying this is
hepatocellular degeneration, hemosiderin deposition in the liver, and congested spleens. Renal
damage is often reported, accompanied by hemosiderin accumulation, renal tubular
degeneration, and intracytoplasmic Hb. Often these effects are more pronounced in females.
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Hematopoiesis in bone marrow and spleen, increased cellularity of bone marrow, and splenic
congestion are all secondary to the hematotoxicity of EGBE and develop as a compensatory
response to hemolysis. In addition, intact erythrocytes have been observed histopathologically in
spleens from EGBE-treated rats, but not in spleens from control animals. This suggests an
increased rate of removal of damaged erythrocytes in EGBE-treated rats (Ghanayem et al.,
1987c). Mild lymphopenia and neutrophilia were observed at hemolytic doses of EGBE
(Ghanayem et al., 1987c) and were reported to be consistent with a "stress" leukogram produced
by the release of endogenous corticosteroids (Wintrobe, 1981a). Neutrophilia, commonly
associated with acute hemolysis or hemorrhage (Wintrobe, 1981b), was also observed.
In the NTP (2000) study, the incidence of hyaline degeneration of the olfactory
epithelium was significantly increased in all exposed groups of male rats and in the 62.5 and
125 ppm groups of females. Hyaline degeneration is the accumulation of intracytoplasmic
globules of highly eosinophilic material in the epithelial cells. While not unique to EGBE,
hyaline degeneration has been shown to occur in rodents exposed to other gases and vapors like
dimethylamine (Buckley et al., 1985) and pyridine (Nikula et al., 1995). It also has been shown
to develop in unexposed aged animals (St. Clair and Morgan, 1992).
In conclusion, humans are significantly less sensitive to the hemolytic toxicity of EGBE
than are typical laboratory species such as mice, rats, or rabbits, although human erythrocytes do
appear capable of responding similarly to the causative EGBE metabolites, albeit at much higher
exposures. This marked species difference in sensitivity has been demonstrated in several
laboratory studies and through the use of in vitro studies using either whole blood or washed
erythrocytes. Based on the results of in vitro testing, blood concentrations of the hemolytically
active metabolite BAA must reach levels in human blood in excess of 7.5 mM for prehemolytic
changes to occur. Comparable effects in rat blood occur at in vitro concentrations approximately
150-fold lower. In addition, blood from potentially sensitive individuals, including the elderly or
those with congenital hemolytic disorders, does not show an increased hemolytic response when
incubated with up to 2 mM BAA for 4 hours. Based on simulations from PBPK modeling,
6-hour whole-body exposure of humans to saturated atmospheres of EGBE will result in
maximum blood concentrations of BAA below those needed to produce hemolysis (Corley et al.,
2005a).
Most of the liver alterations documented throughout the EGBE exposure database are
related directly to hemolysis. Prominent among these alterations is hemosiderin deposition, a
pathological finding whose occurrence is related to hemolysis (Muller et al., 2006). Humans
also experience hemosiderin deposition in the liver, principally in hepatocytes, as a consequence
of excessive hemolysis, such as with thalassemia, a hereditary form of hemolytic anemia (Iancu
et al., 1976). Hemosiderin deposition in the liver is a pathological response that follows the
precursor hematologic effects, which, as a group, do not appear to progress with changes in
duration of exposure from subchronic to chronic.
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4.6. EVALUATION OF CARCINOGENICITY
4.6.1. Summary of Overall Weight of Evidence
Under the Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), EGBE is
deemed "not likely to be carcinogenic to humans" at environmental concentrations below or
equivalent to the RfD and RfC, based on laboratory animal evidence, mode-of-action
information, and limited human study information. The available data indicate that carcinogenic
effects from EGBE are not likely to occur in humans in the absence of the critical noncancer
effects, including hepatic hemosiderin staining and irritant effects at the portal of entry, and are
not likely to be carcinogenic to humans exposed to levels at or below the RfC and RfD values
established in this assessment. Carpenter et al. (1956) (see Section 4.1) reported that no changes
in erythrocyte osmotic fragility were found in human subjects exposed to up to 195 ppm
(942 mg/m3; -600 times the RfC) for two 4-hour periods separated by a 30-minute break. At
oral doses of 400-500 mg/kg with a one-time bolus dose (see Section 4.1), hematuria has been
noted in two human case reports. This dose is 3,000-3,500 times the RfD and would need to be
sustained for a significant period of time to produce hemosiderin deposition. This is unlikely to
occur because the primary response of humans to high oral doses of EGBE, as shown in the case
studies in Section 4.1, is metabolic acidosis, which, if not treated, can lead to shock and
eventually death. No information is available on the carcinogenic effects of EGBE via the oral
or inhalation route in humans. A 2-year inhalation bioassay with mice and rats (NTP, 2000)
reported tumors of the liver in male mice, forestomach tumors in female mice, and tumors of the
adrenal medulla in female rats. Nonneoplastic effects in rats included hyaline degeneration of
the olfactory epithelium and Kupffer cell pigmentation. Nonneoplastic effects in mice included
forestomach ulcers and epithelial hyperplasia, hematopoietic cell proliferation, Kupffer cell
pigmentation, hyaline degeneration of the olfactory epithelium (females only), and bone marrow
hyperplasia (males only).
EGBE has been tested in conventional genotoxicity tests for its potential to induce gene
mutations in vitro and cytogeneticity in both in vitro and in vivo, and the available data do not
support a mutagenic or clastogenic mechanism for EGBE. Two laboratories (Elias et al., 1996;
Hoflack et al., 1995) did report weak genotoxicity responses in vitro at high treatment
concentrations, but results were not replicated in five other labs reporting negative results.
The hypothesized MOA for the tumors observed following EGBE treatment involves
exposure to high doses for prolonged periods of time. This MOA is described in the sections that
follow. The weight of evidence indicates that EGBE is not likely to be carcinogenic to humans
at expected environmental concentrations.
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4.6.2. Synthesis of Human, Animal, and Other Supporting Evidence
NTP (2000) conducted a 2-year inhalation study on EGBE in both genders of F344/N rats
and B6C3Fi mice. Rats (50/gender/group) were exposed to concentrations of 0, 31, 62.5, and
125 ppm (0, 150, 302, and 604 mg/m3) and mice (50/gender/group) were exposed to
concentrations of 0, 62.5, 125, and 250 ppm (0, 302, 604, and 1,208 mg/m3). The NTP report
stated that the highest exposure was selected to produce a 10-15% depression in hematologic
indices, and survival was significantly decreased in male mice at 125 and 250 ppm (54.0 and
53.1%, respectively). While the NTP researchers report that no effect on survival was observed
in rats, the female rats appeared to show a trend toward decreased survival that may have been
attributable to the hematological effects. Mean body weights of rats exposed to 31 and 62.5 ppm
were similar to those of control animals. Mean body weights of the exposed mice were generally
less than those of controls, with females experiencing greater and earlier reductions. From
week 17 to the end of the study, the mean body weights of 125 ppm female rats were generally
less than those of controls.
At the end of the 2-year chronic bioassay (NTP, 2000), neoplastic effects were observed
in female rats and in male and female mice. In female rats, the combined incidence of benign
and/or malignant pheochromocytoma of the adrenal medulla was 3/50, 4/50, 1/49, and 8/49. The
incidence in the high-dose group (16%) did not represent a statistically significant increase over
the chamber control group (6%), but it exceeded the historical control (6.4 ± 3.5%; range 2-
13%) for this effect.
The low survival rate in male mice exposed to 125 and 250 ppm EGBE may have been
due to carcinogenic effects in the liver. A high rate of hepatocellular carcinomas was found in
these exposure groups (10/50 [control], 11/50, 16/50, 21/50); the increase at the high-exposure
level was statistically significant (p < 0.01). However, when hepatocellular adenomas and
carcinomas were combined, no significant increase was observed in any exposure group. The
incidence of hemangiosarcomas in males exposed to 250 ppm (8%) was also significantly
increased (p = 0.046) relative to chamber controls (0/50, 1/50, 2/49, 4/49) and exceeded the
range of historical controls (14/968; 1.5 ± 1.5%; range 0-4%). No significant increases in
benign or malignant hepatocellular tumors or hemangiosarcomas were noted in the female mice,
and the incidence of hepatocellular adenomas actually decreased significantly (p < 0.05) in
relation to the control chamber group (16/50, 8/50, 7/49, 8/49). It should be noted that in light of
the high survival rate of the exposed female mice relative to controls (29/50, 31/50, 33/50,
36/50), the high exposure of 250 ppm may not have provided the maximum tolerated dose.
Forestomach squamous cell papillomas and carcinomas, combined, were significantly
increased (trend test = 0.003) in female mice relative to the chamber control group (0/50, 1/50,
2/50, 6/50). The incidence of these tumor types (12%) at the highest exposure level was also
statistically significant and exceeded the range for the occurrence of these tumors in historical
controls (0.9 ± 1.1%; range 0—3%). The first incidence of these tumors appeared in the group
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exposed to 250 ppm at 582 days, as compared to 731 days at 62.5 and 125 ppm, indicating a
decreased latency period in the highest exposure group. While the incidence of these types of
forestomach tumors was not significantly increased over controls in male mice (1/50, 1/50, 2/50,
2/50), the incidence of squamous cell papillomas (4%) in the two highest exposure groups
exceeded the range for historical controls (0.5 ± 0.9%; range 0-2%). The increased incidence of
forestomach neoplasms in males, as in females, occurred in groups with ulceration and
hyperplasia.
The NTP (2000) study concluded that there was no evidence showing carcinogenic
activity in male F344/N rats, and equivocal evidence of carcinogenic activity in female F344/N
rats based on increased combined incidences of benign (mainly) and malignant pheochromo-
cytoma of the adrenal medulla. The researchers reported some evidence of carcinogenic activity
in male B6C3Fi mice based on increased incidences of hemangiosarcoma of the liver and an
increase in the incidence of hepatocellular carcinoma, as well as some evidence of carcinogenic
activity in female B6C3Fi mice based on increased incidence of forestomach squamous cell
papilloma (mainly) or carcinoma.
With respect to the pheochromocytomas reported in female rats, while the data showed a
positive trend (p = 0.044) and the high-dose tumor frequencies (16%) were above the upper
range of historical controls (13%), the tumor incidence data were not statistically significant.
Further, the NTP (2000) report noted that pheochromocytomas can be difficult to distinguish
from nonneoplastic adrenal medullary hyperplasia. The presence of mild-to-moderate
compression of the adjacent tissue is a primary criterion used to distinguish pheochromocytomas
from medullary hyperplasia; most tumors observed were small and not substantially larger than
the more severe grades of adrenal medullary hyperplasia. Interpretation of these tumors should
be done cautiously. Given the marginal dose response, lack of tumor evidence in any other organ
system of the rats, and reported difficulties in distinguishing pheochromocytomas from
nonneoplastic adrenal medullary hyperplasia, this tumor type was not given significant weight in
the qualitative or quantitative assessment of EGBE cancer potential.
4.6.3. Mode-of-Action Information
4.6.3.1. Hypothesized MOA for Liver Tumor Development in Male Mice
The hypothesized MOA for EGBE-induced liver tumors in male mice is believed to
involve iron accumulation and subsequent oxidative stress due to the hemolytic effects of a
metabolite of EGBE. Male mice developed hepatocellular carcinomas and hemangiosarcomas in
the liver that appeared to be exposure-related. The incidence of hemangiosarcomas was
statistically significant and increased over both concurrent and historical control groups. The
hepatocellular carcinomas were within the range of historical controls for male mice, but are also
considered in this discussion because the dose-response trend is significant and because a similar
MOA has been suggested for this tumor (Klaunig and Kamendulis, 2005). The incidences in the
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high dose group of these two types of tumors were only slightly higher than the upper end of the
range for historical controls. Furthermore, these two tumor types were not increased in other
organs (e.g., bone, bone marrow) and were not noted in either rats or female mice. The
hypothesized MOA involves the hemolysis of RBCs, the accumulation of hemosiderin, and
subsequent oxidative stress that leads to neoplasia in the two cell types believed to be the origin
for these two tumors: hepatocytes for hepatocellular carcinoma and endothelial cells for
hemangiosarcomas. Only the male mice developed these tumor types, and this is hypothesized
to be due to their susceptibility to oxidative stress. Mice are known to be more susceptible to
oxidative stress than rats because of their lower antioxidant capacity (Siesky et al., 2002;
Bachowski et al., 1997). The available studies suggest that iron accumulation from the
hemolytic effects of EGBE produces liver oxidative damage that is more severe in mice and
increased DNA synthesis in both endothelial cells and hepatocytes that may be unique to mice
(Corthals et al., 2006; Siesky et al., 2002). From this and reported differences in antioxidant
capacity1 and background rates of these tumors2 between male and female mice, it can be
hypothesized that events leading to oxidative stress could contribute to the development of
hemangiosarcomas and hepatocellular carcinomas in male mice. A series of events that may lead
to the formation of liver tumors follows:
Step event
(1) EGBE is metabolized to BAL, which is subsequently oxidized to BAA;
(2) BAA causes RBC swelling, triggering sequestration in the spleen by resident
macrophages. When the capacity of these macrophages becomes overwhelmed, the
damaged RBCs make their way into the liver;
(3) Excess Hb from damaged RBCs is taken up by phagocytic (Kupffer) cells of the liver and
stored as hemosiderin;
(4) Oxidative damage and increased synthesis of endothelial and hepatocyte DNA are
initiated by one or more of the following events:
(a) Generation of reactive oxygen species (ROS) from Hb-derived iron within
Kupffer cells and perhaps from within hepatocytes and sinusoidal endothelial cells;
and
1 While the reason for the sex difference in liver tumor susceptibility between male and female mice is not clear, it
has been shown that estrogens can be protective through their antioxidant capacities and through their modulation of
the activities of other antioxidants (Nyska et al., 2004).
2NTP has observed liver hemangiosarcomas in 105/4183 (2.51%) male versus just 35/4177 (0.84%) female
historical controls (Klaunig and Kamendulis, 2005; NTP, 2000). In addition, other chemicals reported by NTP to
cause both early onset hemosiderin buildup and liver tumors have also exhibited this male specificity (U.S. EPA,
2005c).
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(b) Activation of Kupffer cells to produce cytokines/growth factors that suppress
apoptosis and promote cell proliferation.
(5) ROS results in oxidative DNA damage to hepatocytes and endothelial cells;
(6) ROS modulates hepatocyte and endothelial cell gene expression;
(7) ROS stimulates hepatocyte and endothelial cell proliferation;
(8) ROS promotes initiation of hepatocyte and endothelial cells; and
(9) ROS promotes neoplasm formation.
The first two steps, the metabolism of EGBE to BAA and the association of BAA with
hemolytic effects, have been established in many studies, both in vitro and in vivo, with
sensitivities ranging from extreme sensitivity as in rats, mice, rabbits and dogs, to moderate to
extreme insensitivity seen in monkeys, guinea pigs, and humans (see Section 3). The third step
has been seen in both genders of rats and mice exposed to EGBE in multiple studies that
observed hemosiderin within Kupffer cells and hepatocytes after RBC breakdown (Siesky et al.,
2002; NTP, 2000; Kamendulis et al., 1999; Ghanayem and Sullivan, 1993; Ghanayem et al.,
1987a, b; Krasavage, 1986). A number of studies (Park et al., 2002; Siesky et al., 2002;
Kamendulis et al., 1999) provide support for step 4a by showing that in vivo exposure to EGBE
increases 8-OHdG levels (an indicator of oxidative damage) in mice, but not rats, and decreases
vitamin E levels in rats and mice. Using rat and mouse hepatocytes, Park et al. (2002) showed
that FeS04 produced dose-related changes in these same indicators in mouse hepatocytes, but not
in rat hepatocytes, and that treatment with EGBE or BAA did not produce changes in these
oxidative stress parameters. Additionally, Nyska et al. (2004) analyzed 130 2-year
carcinogenicity studies of B6C3Fi mice from NTP bioassays and concluded that a significantly
increased risk of inducing hepatic hemangiosarcomas in male B6C3Fi mice exists in studies with
compounds that caused increased tissue burdens of ROS. Klaunig and Kamendulis (2005) and
Corthals et al. (2006) provided support for step 4b, by showing that the activation of Kupffer
cells, either through RBC hemolytic components and/or iron accumulation in the Kupffer cells,
results in the production of cytokines such as TNFa. The comet assay has been used to assess
DNA damage to endothelial cells from ROS (step 5) (Klaunig and Kamendulis, 2005, 2004;
Reed et al., 2003). While step 6 has not been shown directly for endothelial cells or hepatocytes
exposed to EGBE, induction of oxidative damage has been shown to modify gene expression in
mammalian cells. In addition, ROS production can stimulate cell proliferation and the inhibition
of apoptosis (Klaunig and Kamendulis, 2005; Nyska et al., 2004). Siesky et al. (2002) observed
increased DNA synthesis in endothelial cells and hepatocytes in vivo in mice, but not in rats, at
doses that produced hemangiosarcomas in the mouse liver (NTP, 2000) (step 7). Steps 8 and 9
are consistent with the lack of direct genotoxicity of EGBE (see Section 4.4.4) and the high rate
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of spontaneous endothelial neoplasms in the male mouse liver relative to the rat (Klaunig and
Kamendulis, 2005). The observation of decreased antioxidant capacity and increased 8-OHdG
levels in male mice also lends support to the proposed steps of initiation and promotion of
neoplasms by ROS (Klaunig et al., 1998).
Temporal association and species specificity
Key steps in the proposed MOA (i.e., hemolysis, hepatic hemosiderin buildup, and
oxidative damage) have all been observed in subchronic or shorter-duration rat and mouse
studies of EGBE (Siesky et al., 2002; NTP, 2000; Kamendulis et al., 1999) well in advance of
tumor formation. Differences between rats and mice with respect to these responses may help to
explain the relative sensitivity of male mice to the formation of liver tumors following EGBE
exposure. In mice, Siesky et al. (2002) observed a dose-dependent increase in levels of liver
oxidative stress indicators at exposure days 7 and 90, increased endothelial cell DNA synthesis at
exposure days 7 and 14, and increased hepatocyte DNA synthesis at 90 days. No increase in
ROS or the DNA synthesis of either cell type was observed in rats at any time point.
Dose-response relationships
Five chemicals have been determined by the NTP to cause hemosiderin buildup in the
livers of mice. As shown in Table 4-8, all three studies that reported hemosiderin buildup in
Kupffer cells of male mice within 13 weeks of exposure also showed an increased incidence of
hemangiosarcomas and hepatocellular carcinomas following chronic exposure. The dose
responses for endpoints describing possible precursor effects, splenic hematopoietic cell
proliferation, and liver hemosiderin accumulation appear to be dose-related and coincident to the
formation of tumors. Dose-responses for several hemolytic effects were also observed in rats
exposed to EGBE, but liver tumors were not increased in rats at any dose. However, the high
dose used in the EGBE rat study was only half that of the high dose used in the mouse study, and
the 2-year duration of these studies represents a smaller fraction of the rat lifespan, leaving the
possibility that similar responses could have been observed in rats if higher and longer EGBE
exposures had been administered.
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Table 4-8. Incidence of liver hemangiosarcomas and hepatocellular
carcinomas in studies of NTP chemicals that caused increased hemosiderin
in Kupffer cells in male mice
Chemical (NTP TR#)
Hemosiderin
SCa
Hemangiosarcoma
Hepatocarcinoma
Type
2-Butoxyethanol
(EGBE) (TR-484)
0/50, 0/50, 8/49°,
30/49°
Yes
0/50, 1/50, 2/49, 4/49b
10/50, 11/50, 16/49,
21/49°
I
p-Chloroaniline
hydrochloride (TR-351)
0/50, 0/49, 0/50, 50/50d
Yes
2/50, 2/49, 1/50, 6/50
3/50, 7/49, 11/50b, 17/50°
G
p-Nitroaniline
(TR-418)
1/50, 1/50, 8/50b,
50/50°
Yes
0/50, 1/50, 2/50, 4/50
10/50, 12/50, 13/50, 6/50
G
C.I. Pigment Red 3
(TR-407)d
0/50, 5/50, 30/50, 41/50
No
0/50, 1/50, 1/50, 0/50
5/50, 10/50, 8/50, 4/50
F
o-Nitroanisole
(TR-416)
0/50, 0/50, 3/50, 16/50°
No
2/50, 2/50, 1/50, 0/50
7/50, 12/50, 11/50, 7/50
F
aChemicals that caused hemosiderin accumulation in Kupffer cells following subchronic (SC) exposure are
identified with a "yes" in this column.
Statistically significant difference, p < 0.05.
Statistically significant difference,/) <0.01.
Statistics not reported.
F = feed; G = gavage; I = inhalation
Of the five chemicals listed in Table 4-8, liver hemangiosarcomas were observed with
only three chemicals that induced hemosiderin buildup by week 13, but not with the other two
chemicals for which hemosiderin buildup was not observed until the end of the 2-year study
(Gift, 2005; U.S. EPA, 2005c). Two of the three chemicals that induced early liver hemosiderin
accumulation and increased hemangiosarcoma incidence, EGBE and p-chloroaniline
hydrochloride, also induced an increase in hepatocellular carcinomas. Early buildup of
hemosiderin combined with early increases in endothelial cell and hepatocyte DNA synthesis
would result in a longer exposure of cells to oxidative damage via iron-generated radicals
(step 4). This would be consistent with a mechanism involving a continuing cycle of damage
and repair and accumulation of DNA mutations (steps 5 and 6). In addition to an earlier onset of
hemosiderin buildup, mice also show evidence of a more sustained hemolytic response to EGBE
than rats.3
Biological plausibility and coherence of the database
Oxidative damage plays an important role in the pathogenesis of several diseases,
including cancer and cardiovascular disease (Djordjevic, 2004; Lesgards et al., 2002; Klaunig
et al., 1998). In support of the proposed hypothesis, increased ROS are known to accompany the
3Mice experienced an increase in liver and splenic hematopoietic cell proliferation throughout the 2-year NTP
(2000) study, while rats tended to compensate for the effects of EGBE after a few months. This increased tolerance
in rats is evidenced by a lack of induction of splenic hematopoiesis at the end of the 2-year NTP (2000) study.
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release of large amounts of iron from hemolysis (Ziouzenkova et al., 1999). If EGBE causes
oxidative stress via hemolysis, then the production of protein and DNA damage would be
expected, including the production of 8-OHdQ accompanied by a decrease in antioxidant levels,
such as Vitamin E (Houglum et al., 1997; Yamaguchi et al., 1996; Wang et al., 1995). These
effects were verified by both Siesky et al. (2002) and Kamendulis et al. (1999), who measured a
dose-dependent increase in levels of ROS indicators, including 8-OHdG and malondialdehyde,
and a decrease in vitamin E levels in the livers of mice, but not in rats, after acute and subchronic
exposure to EGBE. The fact that mice appear more susceptible than rats to ROS from EGBE
exposure is consistent with the proposed MO A, and is a reasonable explanation for why
hemangiosarcomas are not observed in rats following chronic EGBE exposure, despite the fact
that rats are at least as sensitive as mice to the hemolytic effects of EGBE.
Iron, which is known to accumulate in cells of rodent livers following EGBE exposure,
can produce hydroxyl radicals in combination with oxidative by-products via the Fenton reaction
(Kamendulis et al., 1999). The damaging effects of iron overload to liver sinusoidal cells have
been shown in rats following a single i.p. injection of 200 mg iron/kg (Junge et al., 2001). In
addition, endothelial cells appear to be relatively sensitive to oxidative stress (Spolarics, 1999;
DeLeve, 1998). Liver hemangiosarcomas develop from the endothelial cell component of the
vascular sinusoidal cells of the liver (Frith and Ward, 1979).
In vivo studies have indicated that pretreatment of rats with an ALDH inhibitor, pyrazole,
prior to a single 125 mg EGBE/kg gavage exposure protected against hemolysis (Ghanayem
et al., 1987b), presumably by blocking the production of both BAL and BAA. Pretreatment of
rats with an ADH inhibitor, cyanamide, prior to a single 125 mg EGBE/kg gavage exposure,
reduced hemolytic responses, but increased RBC swelling, increased mortality, decreased BAA
formation and excretion in the urine, and increased the urinary excretion of EGBE conjugates
with glucuronide and sulfate (Ghanayem et al., 1987b). This hematotoxicity in the presence of
cyanamide may be due to BAL, but residual BAA may also be a factor. EGBE + cyanamide
decreased BAA concentrations in rats; however, some BAA was formed and the BAA tu was
increased (Ghanayem et al., 1990), and, when Ghanayem et al. (1987b) administered a gavage
dose of 125 mg BAL/kg + cyanamide to rats, the researchers observed almost no hemolytic
activity. Furthermore, gavage administration to rats of 125 mg EGBE/kg and the molar
equivalent of BAL and BAA resulted in no significant difference between the hemolytic effects
of the three chemicals between 2 and 24 hours postexposure (Ghanayem et al., 1987b). These
data suggest that EGBE's hemolytic activity (without coexposures) is due to BAA and that the
metabolism of EGBE and BAL to BAA takes place rapidly and completely.
Relevance of the hypothesized MO A to humans
The occurrence of liver tumors in mice exposed to EGBE is hypothesized to occur
through an MOAthat requires first a dosage of EGBE that is high enough to cause sustained
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hemolysis of RBCs and, second, leads to sufficient buildup of hemosiderin in the Kupffer cells of
the liver to produce ROS and subsequent neoplasm formation.
Several studies have examined the susceptibility of RBCs to BAA-induced hemolysis and
have found a range in sensitivity from the sensitive (rats, mice, rabbits, and dogs) to the less
sensitive (monkeys, guinea pigs, and humans). Human volunteers experienced no hemolysis
from controlled laboratory acute inhalation exposures (up to 195 ppm), a dosage which caused
significant erythrocyte fragility in rats (Carpenter et al., 1956). Only mild hemolytic effects have
been observed in humans acutely exposed to oral doses of EGBE (400-1,500 mg/kg), doses that
have been shown to cause marked hemolytic effects in rats (Ghanayem et al., 1987c; Grant et al.,
1985). RBCs from populations that were potentially more sensitive to hemolysis in general—the
elderly, and individuals with sickle cell anemia and hereditary spherocytosis—were tested in
vitro and not found to exhibit hemolysis after exposure to concentrations 40-fold higher (the
highest tested in the study) than those shown to induce hemolysis in rat RBCs (Udden, 1994;
Udden and Patton, 1994). In an in vitro study of RBCs from hospitalized children and adults,
concentrations of up to 150-fold higher than those used in rat studies did not produce hemolysis
(Udden, 2002). The resistance of human RBCs to the initial event of hemolysis makes it
unlikely that they would experience the subsequent effects of increased hemosiderin deposition
through this pathway, and consequently, humans would not be at increased risk of tumor
development through this MOA.
Other possible MO As for liver tumor development in male mice
Although certain key events in EGBE's MOA for the development of liver tumors in
male mice are fairly well-described and plausible, some alternative considerations (also
supported by scientific literature) may be involved. ROS can potentially be derived from two
sources: iron overloading in the liver (through Fenton and Haber-Weiss reactions) and Kupffer
cell activation. Via either source, oxygen radicals can induce oxidative damage to DNA and
lipids as documented in the liver following EGBE treatment (Siesky et al., 2002). The activation
of Kupffer cells, through phagocytosis of RBC hemolytic components or iron in the Kupffer cell,
results in the production of cytokines, possibly including vascular endothelial growth factor that
may elicit a growth response on endothelial cells (Corthals et al., 2006). In addition to the
production of oxidative DNA damage, ROS, whether derived from Kupffer cell activation or
other biological processes, can alter gene expression (e.g., MAP kinase/AP-1 and NFkB),
resulting in stimulation of cell proliferation and/or inhibition of apoptosis (Klaunig and
Kamendulis, 2004).
Another recognized mechanism for the development of chemically-induced liver
hemangiosarcomas involves direct interaction with DNA. This MOA is recognized for vinyl
chloride and thorotrast, two agents that are known to induce hemangiosarcomas in humans. The
EGBE metabolite BAL is considered to have the greatest potential to interact with DNA, since it
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has been shown to cause in vitro SCE at concentrations ranging from 0.2 to 1 mM (Elliot and
Ashby, 1997). However, high ADH activity in the liver, as in the forestomach, is expected to
result in very short residence time and low Cmax liver tissue concentrations of BAL. Corley et al.
(2005b) extended their 1994 model (Appendix B) to include the metabolism of EGBE to BAL
via ALDH and the subsequent metabolism of BAL to BAA via ADH in both the liver and
forestomach. As shown in Figure 4-1, using rate constants derived from mouse stomach
fractions (Green et al., 2002) and making several assumptions about the use of these enzyme
activity data, Corley et al. (2005a) estimated that 250 ppm EGBE (the highest concentration used
in the NTP [2000] study) would result in peak Cmax values of 7 |iM EGBE, 0.5 |iM BAL, and
3,250 [xM BAA in liver tissue of male mice at the end of a 6-hour exposure period.
BE
BAL
BAA
CmaxfuM)
7
0.5
3,250
3,500
BE & Metabolites in Liver
50
-BE
¦ BAL
-BAA
40
30
20
10
Time fh)
BE = EGBE
Source: Adapted from Corley et al. (2005b).
6 11
Time (h)
18
Figure 4-1. Simulated concentrations of EGBE, BAL, and BAA in liver
tissues of female mice exposed via inhalation for 6 hours to 250 ppm EGBE.
Thus, the Corley et al. (2005b) PBPK model suggests that the high cytotoxic
concentrations of BAL that showed some evidence of clastogenicity may not be relevant to the
target organ where lower concentrations of BAL would exist in the presence of metabolizing
enzymes. A recent gavage study performed by Deisinger and Boatman (2004) provided support
for the Corley et al. (2005b) model and the predicted low levels of the BAL metabolite in liver
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tissue.4 In addition, as discussed in Section 4.4.4, evidence from in vivo and in vitro
genotoxicity assays does not suggest that BAL would have any significant genotoxicity in vivo.5
Furthermore, the MOA for hemangiosarcoma induction by genotoxins such as vinyl chloride and
thorotrast involves the initiation of hepatocellular and sinusoidal cell hyperplasia and sinusoidal
compression, leading to the development of fibrous septa, generally in the periportal area, out of
which eventually develop multiple areas of angiosarcomas (Foster, 2000). EGBE exposure does
not generate this same pattern of effects prior to the development of cancer in mice.
4.6.3.2. Hypothesized MOA for Forestomach Tumor Development in Female Mice
A significant increase over controls (experimental and historical) of papillomas and one
carcinoma of the forestomach (6/50; 12%) was reported by NTP (2000) in female mice exposed
for 2 years to 250 ppm EGBE by inhalation. An increase was also seen in squamous cell
papillomas in male mice, although this did not reach the level of statistical significance.
Significant increases in forestomach papillomas and carcinomas were not observed in rats of
either gender. The study also showed statistically significant, dose-dependent increases in
hyperplasia for both male and female mice, and for ulceration in female mice. Male mice
showed significant increases in ulceration at the 125 ppm exposure. The process of irritation,
hyperplasia, and ulceration is thought to be a part of the cell injury and regeneration process
elicited by irritant chemicals such as EGBE, and the study authors hypothesized that the
neoplasia may occur due to an increase in the cell replication associated with regeneration.
A hypothesized series of events that may lead to the formation of forestomach tumors is
as follows:
Step event
(1) Deposition of EGBE/BAAin the stomach and forestomach via consumption or
reingestion of EGBE laden mucus, salivary excretions, and fur material;
(2) Retention of EGBE/BAA in food particles of the forestomach long after being cleared
from other organs;
(3) Metabolism of EGBE to BAL, which is rapidly metabolized to BAA systemically and in
the forestomach;
(4) Irritation of target cells by BAA leading to hyperplasia and ulceration;
4The Corley et al. (2005b) model predicts that the concentrations of BAL in liver tissues of male and female mice
would be 17 and 29 |iM, respectively, following oral gavage exposure to 600 mg/kg EGBE. The levels of BAL
actually observed in the liver tissue of male and female mice, following oral gavage exposure to EGBE at
600 mg/kg, 3 and 4 (iM, respectively, were even lower than the predicted values (Deisinger and Boatman, 2004).
5The exposure concentrations that would be necessary to cause these effects in humans, if attainable at all, are likely
to be much higher than the RfC and RfD.
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(5) Continued injury by BAA and degeneration leading to high cell proliferation and
turnover; and
(6) High levels of cell proliferation and turnover leading to clonal growth of spontaneously
initiated forestomach cells.
There are a number of studies that have demonstrated steps 1 and 2 (i.e., the deposition
and retention of EGBE and BAA in the forestomach). Studies have shown this occurs following
whole-body exposure (Green et al., 2002; Poet et al., 2002), nose-only inhalation (Poet et al.,
2002), i.v. exposure (Green et al., 2002; Poet et al., 2002; Bennette, 2001), i.p. exposure (Poet
et al., 2002; Corley et al., 1999), s.c. exposure (Corley et al., 1999), and gavage exposures
(Green et al., 2002; Poet et al., 2002; Ghanayem et al., 1987a, b). It is of note that following i.v.
and inhalation exposures in mice, EGBE metabolites rapidly accumulate in salivary secretions
and are swallowed (Green et al., 2002; Bennette, 2001), leading to the collection and retention of
the chemical(s) in the forestomach. The same process is likely to occur through other systemic
exposures. The metabolism of EGBE to BAA (step 3) has been shown in both in vitro and in
vivo tests with rats, mice, rabbits, guinea pigs, dogs, monkeys, and humans, (see Section 3) and
is further supported by the EGBE PBPK model developed by Corley et al. (2005b). Step 4, the
irritation of target cells, has been seen in both genders of B6C3Fi mice, (Poet et al., 2003; Green
et al., 2002; NTP, 2000), with irritation and a compensatory proliferative response reported
following exposure to EGBE. Female mice were shown to have more extensive and severe
forestomach lesions than male mice and were observed in the NTP (2000) study to have
statistically significant increases in forestomach tumors. This suggests the importance of the
continued damage and high cell proliferation (step 5) that is associated with tumor formation.
Green et al. (2002) found that the number of cells in S-phase (an indication of cell turnover)
increased in a dose-dependent fashion after exposure to EGBE and BAA, even though none of
the changes were statistically significant due to the high turnover for the control group. Step 6,
high levels of cell proliferation and turnover, leads to clonal growth of spontaneously initiated
cells and is supported by the continuum of effects observed in the mice (Green et al., 2002; NTP,
2000) and the effects seen with other irritant compounds (Kroes and Wester, 1986).
Green et al. (2002) also provided relevant information regarding step 3 through
examination of the activity and localization of ALDH and ADH in the stomach tissues of mice,
rats, and a human sample. Whole body autoradiography of mice that had been exposed to
radiolabeled EGBE was also performed. Histochemical staining of stomach tissues from the
rodent species showed the dehydrogenase enzymes to be heavily concentrated in the stratified
squamous epithelium of the forestomach of both rats and mice, whereas their distribution in the
glandular stomach was more diffuse. By comparison, histochemical analysis of a human
stomach tissue sample showed both enzymes to be present and evenly distributed throughout the
epithelial cells of the entire stomach mucosa. A marked species difference in ALDH activity in
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the forestomach was observed between rats (Km = 0.29 mM; Vmax = 1.627 nmol/minute per mg
protein) and mice (Km = 46.59 mM; Vmax = 17.094 nmol/minute per mg protein) with Km values
up to one order of magnitude greater in mice compared to rats. These differences indicate that
mice forestomach tissues would have the capacity to metabolize appreciably larger amounts of
EGBE to BAL, and subsequently to BAA, than would the rat forestomach. Whole body
autoradiography of mice exposed to EGBE demonstrated selective accumulation in the
forestomach, which would provide substrate for these enzymes. Collectively, these data
demonstrate several points regarding the observed forestomach toxicity in mice following
exposure to EGBE, including: (1) the accumulation of EGBE in the target tissue, the
forestomach, of mice; (2) a high degree of localization in the forestomach (as compared to the
glandular stomach) tissues of both rats and mice of the enzymes that metabolize EGBE to the
corresponding carboxylic acid; and (3) kinetic differences in these enzymes consistent with mice
being capable of metabolizing more EGBE to BAA than rats. The data also suggest that human
stomach tissues do not have a high localization of the EGBE metabolizing enzymes; the diffuse
distribution of these enzymes in the human stomach sample examined is more similar to the
distribution seen in the glandular portions of the rodent species examined. These observations
suggest that human stomach tissues would be less capable of accumulating and localizing BAA
than rat tissues and, thus, would be less likely to be exposed to the irritating effects of BAA.
Temporal association
All of the steps in the proposed MOA have been observed to occur in female mice prior
to tumor formation. NTP (2000) reported that female mice experienced epithelial hyperplasia
(1/10, 5/10, 9/10, and 10/10) after just 13 weeks of exposure at the same exposure levels used in
the chronic study, 0, 62.5, 125, and 250 ppm. The reported incidence of a forestomach papilloma
or carcinoma in female mice was 731, 731, and 582 days in the 62.5, 125, and 250 ppm exposure
groups, respectively. This is consistent with the findings of Ghanayem et al. (1994, 1993, 1986),
who investigated the temporal relationship between the induction of this type of forestomach
lesion by another nonmutagenic irritant, ethyl acrylate (EA), and the development of squamous
cell papillomas and carcinomas. These researchers observed cell proliferation/hyperplasia in the
forestomach of all rats that received EAby gavage (200 mg/kg, 5 days/week) for 6 or 12 months.
All these potentially precancerous forestomach lesions regressed in animals treated with EA for
6 months and allowed 2 or 15 months of recovery, and no forestomach neoplasms were
observed. Although EA, an unsaturated aldehyde, is not a metabolite of EGBE, it is an analog of
BAL and a much more potent carcinogen than either EGBE or BAL (Gold et al., 1993).
For EGBE, the high incidence of forestomach hyperplasia, the relatively lower incidence
of papillomas, and the late occurrence of a single carcinoma in the high 250 ppm exposure group
are suggestive of a temporal relationship and tumor progression following EGBE exposure to
female mice. Male mice may show the beginnings of tumorigenic effects as the incidence of
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papillomas increases, but such findings have not been statistically significant compared to
concurrent or historical controls. No hyperplasia and no tumors were observed in inhalation
studies of rats (NTP, 2000) or in drinking water studies of mice (NTP, 1993), supporting the need
for these steps prior to tumor formation.
Dose-response relationships
The incidences of epithelial hyperplasia (6/50, 27/50, 42/49, 44/50) and ulceration (1/50,
7/50, 13/49, 22/50) in EGBE-exposed female mice were dose-related and significantly increased
over both concurrent and historical controls at lower dose levels than the forestomach tumors.
The hyperplasia was often associated with ulceration, particularly in the female mice.
Forestomach tumors observed by NTP (2000) (incidence 0/50, 1/50, 2/50, 6/50) increased over
control animals only at exposure levels above those that caused significant hyperplasia. The
increased incidence of the forestomach neoplasms occurred in groups with ulceration and
hyperplasia, suggesting a dose-dependent relationship between the nonneoplastic and the
neoplastic lesions.
Biological plausibility and coherence of the database
Both mutagenic and nonmutagenic chemicals have been shown to induce forestomach
tumors in rodents (NTP, 2000; Ghanayem et al., 1994, 1993, 1986; Kroes and Wester, 1986).
Some nonmutagenic substances that cause such tumors appear to require long-term contact with
the forestomach epithelium, leading to irritation, cell proliferation, and neoplasia. The
overstimulation of repair processes and enhancement of growth-promoting factors are believed
to be involved (Harrison, 1992). Promotion and other activities associated with the stimulation
of cell proliferation have been observed for many of these compounds (Ghanayem et al., 1994;
Clayson et al., 1991). High concentrations of EGBE and its BAA metabolite sequestered in the
forestomach are assumed to cause chronic irritation and the more serious damage observed in the
forestomach lining of female mice. The incidence of ulcers was significantly increased in all
exposed groups of females. NTP (2000) suggests that EGBE exposure-induced irritation caused
inflammatory and hyperplastic effects in the forestomach and that the neoplasias (papillomas and
one carcinoma) were associated with a continuation of the injury/degeneration process.
Other substances that induce forestomach hyperplasia in male and female mice following
inhalation exposure include acetonitrile, 1,3-butadiene, and chloroprene (U.S. EPA, 2005c).
Both propionic and butyric acid have been shown to induce proliferative responses in
forestomach epithelium after only 7 days, and long-term propionic acid exposure has produced
papillomas in the rat forestomach (Kroes and Wester, 1986). Since high levels of EGBE and
BAA have been observed in the stomachs of mice following i.v., i.p., gavage, and inhalation
exposures, it is apparent that the chemical partitions to the forestomach via multiple routes,
including grooming of fur, systemic blood circulation, ingestion of salivary excretions and
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respiratory tract mucus, and possibly repartitioning of the stomach contents (Poet et al., 2003;
Green et al., 2002). Because the forestomach functions as a storage organ, there is a reduced
requirement for vascularization. The planar capillary network within the epithelial layers of the
rodent forestomach contrasts strongly with the thick mucosal network of capillaries in the
glandular stomach of rodents (Browning et al., 1983). The cells of the forestomach epithelium,
especially the more superficial squamous cells, are separated from capillaries by substantial
diffusion distances (Bueld and Netter, 1993; Browning et al., 1983). In addition, the glandular
stomach contains a complex mucosal protection and buffering system necessary to withstand the
high acidity of the digestion process. As a result, irritant substances that concentrate in the
forestomach may produce hyperplasia in the forestomach, but not in the glandular stomach or
other gastrointestinal tissue (Kroes and Wester, 1986).
Relevance of the hypothesized MO A to humans
While this proposed MOA is thought to be of qualitative relevance to humans, the EGBE
exposure concentrations that would be necessary to cause hyperplastic effects and tumors in
humans, if attainable at all, are likely to be much higher than the concentrations necessary to
cause forestomach effects in mice for several reasons:
(1) The mouse forestomach serves a food storage function and the slow rate of emptying
provides a sink for EGBE where it is metabolized to BAA and remains in contact with
squamous epithelium long after EGBE has been cleared from the rest of the body. While
the human esophagus is histologically similar to the murine forestomach, the contact time
with food and other ingested substances is short, because this organ does not have a
storage function. Thus, the risk for esophageal tumors is low (Kroes and Wester, 1986).
The human stomach also has a faster rate of emptying than the rodent forestomach, and is
further protected from irritant compounds by a mucous layer that is not present in the
rodent forestomach;
(2) The localization of the enzymes needed for acid production in the human stomach tissue
is not the same as it is in the target, the rodent forestomach (Green et al., 2002);
(3) A benchmark dose (BMD) analysis (see Appendix C) indicates that the exposure
concentrations necessary to cause hyperplastic effects in humans would be much higher
than the existing RfD and RfC for EGBE.
Other possible MO As for forestomach tumor development in female mice
Though the evidence favors the hypothesis that BAA is the principal toxic metabolite of
EGBE, roles for BAL (Ghanayem et al., 1987b) and butyric acid (Harrison et al., 1991) have
been suggested. It is not likely that butyric acid plays a significant role in the toxicity of EGBE,
particularly at environmentally relevant concentrations. High concentrations of butyric acid have
caused ulceration and other preneoplastic lesions in mice (Harrison et al., 1991). However, low
concentrations of butyric acid do not appear to be harmful, since it naturally occurs in the diet
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through the fermentation of fiber and starch and as a significant portion (up to 10 mol%) of total
bovine milk fatty acid (Smith and German, 1995).
Another possible alternative MOA could exist if EGBE or one of its metabolites were to
have the capability of damaging a cell through direct interaction with its DNA. As has been
discussed in Section 4.4.4, there is very limited evidence that EGBE or BAA is genotoxic. BAL,
a short-lived metabolite of EGBE, has been found to be clastogenic in in vitro assays without
enzyme activation at concentrations ranging from 0.2 to 1 mM (Elliot and Ashby, 1997).
However, as has been discussed, in vivo and in vitro genotoxicity assays do not suggest that BAL
would have any significant genotoxicity in vivo. In addition, chemicals for which mutagenesis/
genotoxic effects play a significant role generally induce more tumors at earlier time points than
near the end of the conducted bioassays due to their ability to both initiate and promote tumor
pathogenesis. The mutagenic compound ethylene dibromide, for instance, was reported to
induce forestomach tumors in all dose groups 168-280 days from the start of exposure (U.S.
EPA, 2004). EGBE is consistent with other forestomach carcinogens that are not mutagenic,
such as EA, in that observed tumors generally did not progress to carcinoma and were not
observed until well into the study (i.e., after long periods of forestomach cell damage and repair).
The first reported incidence of forestomach papilloma or carcinoma in female mice was 731,
731, and 582 days in the 62.5, 125, and 250 ppm EGBE exposure groups, respectively (NTP,
2000).
It does not appear that EGBE, BAL, or BAA preferentially binds to stomach tissue
macromolecules (Poet et al., 2003; Green et al., 2002). Poet et al. (2003) found that high levels
of EGBE concentrate in the food content of the forestomach following i.p. exposure (Poet et al.,
2003), indicating that the observed sequestering of EGBE in the forestomach is related to its
retention in the food that remains there, not to preferential binding to proteins within
forestomach tissue.
4.6.3.3. Conclusions About the Hypothesized Modes of Action
Inhalation exposure of B6C3Fi mice to EGBE gave rise to tumors in the liver and
forestomach. The liver tumors, hemangiosarcomas, and hepatocellular carcinomas occurred in
males only and were significantly elevated over controls with a positive trend test for the
hemangiosarcomas. The forestomach tumors occurred in females only, had a positive trend test
result, and were significantly increased over controls only at the highest dose.
The MO As that have been developed for these tumors reflect the evidence for the
nonmutagenic nature of EGBE and its metabolites. For the liver tumors, the hypothesized key
steps of the MOA are metabolism of EGBE to BAA, hemolysis of RBCs with release of Hb and
hepatic hemosiderin accumulation, followed by oxidative stress, modulation of gene expression,
cell proliferation, promotion, and neoplasm, leading to the formation of liver tumors. For the
forestomach tumors, the hypothesized steps are metabolism to BAA, followed by tissue irritation
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and subsequent cytotoxicity, compensatory proliferation, and the induction of forestomach
tumors. No other viable MO As have been identified that adequately explain the existing
laboratory animal and human observations.
Both of these MO As have some degree of qualitative significance for humans since the
principal biological components supporting them are all present and the processes can occur in
humans. Collectively, however, the evidence presented in this assessment for these MO As
suggests that both MO As have only limited quantitative significance to humans, principally due
to kinetic/dynamic differences from the rodents. In the case of the liver tumors, in vitro data
suggest there is a 40- to 150-fold difference in the dose that produces hemolytic changes in the
RBCs of humans as compared to rodents. This difference is supported by the Carpenter et al.
(1956) study in which no changes in erythrocyte fragility were measured in humans at the
highest tested concentration, 195 ppm, but increased erythrocyte fragility was measured in
coexposed rats. Further, PBPK model simulations performed by Corley et al. (2005a) predict
that given the vapor pressure of EGBE, the maximum blood level of BAA that can be obtained
from inhalation exposure would be lower than the predicted concentrations from bolus exposures
that have not resulted in hemolytic effects, and lower than concentrations that have been shown
to produce an effect on human RBCs in vitro (Udden, 2002). In the case of the forestomach
tumors, the primary difference between mice and humans is in the degree of kinetics in the
metabolizing enzymes involved in the production and clearance of BAA. Thus, the hypothesized
key events in the MO As for the animal tumors (liver and forestomach) are not likely to occur in
humans, especially at low doses.
Based on the preceding analysis, EGBE is deemed not likely to be carcinogenic to
humans at the calculated RfC and RfD values presented in this document when examining it on
its physical-chemical properties, toxicokinetic and dynamic factors, and MOA information.
4.7. SUSCEPTIBLE POPULATIONS
The hemolytic effect of EGBE is presumed to be caused by its primary metabolite, BAA,
interacting with the RBC membrane. Potentially susceptible subpopulations would include
individuals with enhanced metabolism or decreased excretion of BAA. As discussed in
Section 4.7.1, older rats have reduced ability to metabolize the toxic metabolite BAA to C02 and
a diminished ability to excrete BAA in the urine (Ghanayem et al., 1990, 1987c). However, the
relevance of this finding to the possible susceptibility of elderly humans is uncertain; as
discussed in Section 3, humans may have conjugation pathways for the excretion of BAA, such
as BAA-glutamine and BAA-glycine, which are not present in rats.
It would also be expected that individuals whose RBC membranes are more susceptible
to the lysis caused by BAA would be more sensitive to effects from EGBE exposure. However,
RBCs from normal, aged, sickle-cell anemia, and hereditary spherocytosis patients were no more
sensitive to the hemolytic effects of BAA than RBCs from healthy volunteers when tested in
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vitro (Udden, 1994). As work in this area continues, further information on the metabolic or
structural differences that result in the lower sensitivity of human RBCs compared to rat RBCs
may eventually identify characteristics in humans that may indicate increased susceptibility. For
instance, it is unknown if a genetic predisposition to hemolytic anemia from other causes, such
as G6PD deficiency, would lead to increased susceptibility to EGBE-induced hemolysis. G6PD
deficiency appears in approximately 400 variants, thus describing a genetically heterogeneous
disorder. The label has been applied to all types of moderate to severe enzyme deficiency with
intermittent, induced hemolytic episodes with chronic hemolytic anemia. To date, complete
deficiency of this enzyme has not been identified. The clinical result is the reduced ability to
produce nicotinamide adenine dinucleotide phosphate, an enzyme required for reactions of
various biosynthetic pathways, as well as for the stability of catalase and the maintenance of
GSH levels. Catalase and glutathione peroxidase are the primary enzymes in the detoxification
of hydrogen peroxide. Thus, cells are dependent on G6PD for this pathway; without it, they are
vulnerable to oxidative damage. RBCs are sensitive to this loss of enzyme: they rely on this
system for their antioxidant defenses. Other human risk factors for anemia include ingestion of
certain therapeutic drugs, infections, family history, diet, and systemic illnesses (Berliner et al.,
1999).
Individuals with hereditary hemochromatosis (HH) represent a population potentially
susceptible to increased release of iron from any source. It is reported that 5/1,000 persons of
northern European descent are homozygous for the gene or genes that cause hemochromatosis,
although it is unknown what proportion of this population will go on to develop the HH
phenotype (Pietrangelo, 2004). Individuals with this disorder are not able to reduce their
absorption of iron in response to increasing iron levels in the body. Iron stores in the body
continue to increase. The iron stores normally start out as ferritin, then eventually become
aggregates of a breakdown product of ferritin called hemosiderin. HH is a condition
characterized by excessive iron deposition in the form of hemosiderin, found in the liver, heart,
skin, joints, pancreas, and other endocrine organs. It is unknown whether individuals with this
condition would be susceptible to the effects of EGBE exposure. There is no indication in the
literature, however, that RBCs in individuals with HH are more fragile, and it is therefore
unlikely that HH would increase the risk of hemolysis or additional hemosiderin deposition from
EGBE exposure. Studies have shown differences in the localization of iron between HH patients
and rodents. In HH patients, iron appears to accumulate preferentially in the parenchymal cells
of the liver, early on as ferritin and later as hemosiderin, due to increased iron absorption from
the duodenum; late in the disease, iron storage is seen in Kupffer cells and reticular endothelial
cells of the bone marrow (Knutson and Wessling-Resnick, 2003; Valberg et al., 1975). In
contrast, EGBE-induced toxicity in mice and rats results in initial and preferential accumulation
of hemosiderin in Kupffer cells by phagocytosing senescent RBCs (NTP, 2000). While it is clear
that macrophages and other cells can in fact contain hemosiderin, the relative level compared to
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hepatocytes is much less; staining in these cells is typically seen in late stages of the disease
(Kwittken and Tartow, 1966). The human course of developing hepatocellular carcinomas as a
consequence of HH is also quite different compared to the development in the mouse model of
hemangiosarcoma. Human cases of HH-induced hepatocellular carcinoma are typically
observed in the presence of cirrhosis of the liver, another long-term process that reflects the
chronic nature of the disease progression in humans (Harrison and Bacon, 2005).
Hemangiosarcomas, the tumor type of concern in the male mice, have not been associated with
HH in the literature.
4.7.1. Possible Childhood Susceptibility
A number of factors may differentially affect children's responses to toxicants. The only
human toxicity information available on the toxicity of EGBE to children is from the case study
by Dean and Krenzelok (1991). They observed 24 children, aged 7 months to 9 years, after oral
ingestion of at least 5 mL of glass window cleaner containing EGBE in the 0.5-9.9% range
(potentially 25-1,500 mg EGBE exposures). Two children who had taken >15 mL did well after
gastric emptying or lavage and observation in the hospital. The remainder were watched at home
after receiving diluting oral fluids. No symptoms of EGBE poisoning or hemolysis were
observed. While the effects reported in adult poisonings have been more severe than those
reported in these children, the adults tended to consume larger volumes and different
concentrations of EGBE, making it extremely difficult to assess the correlation of toxic effects
with age sensitivity.
As discussed above, there are numerous risk factors for anemia that might predispose an
individual to, or compound the effects of, EGBE-induced hemolysis. It is generally recognized,
however, that children do not share the same risk factors for anemia as adults for a number of
reasons, including: (1) a higher rate of RBC turnover; (2) lower incidence of neoplastic disease
in childhood as either a direct or indirect cause of anemia (<7,000 of the 1,000,000 new cases of
cancer each year in the U.S. occur in individuals <15 years of age); (3) the fact that iron
deficiency is almost always secondary to nutritional factors in children; (4) the relative rarity of
alcoholism and its related liver disease; (5) a much lower incidence of anemia associated with
thyroid disease; and (6) a rarity of cardiovascular disease other than congenital heart diseases,
with the result that valve replacement, malignant hypertension, and the use of certain drugs are
not usually a factor (Berliner et al., 1999; Hord and Lukens, 1999).
Anemia in children is usually associated with an abnormality of the hematopoietic system
(Berliner et al., 1999; Hord and Lukens, 1999). Studies of the osmotic fragility and
deformability of RBCs exposed to BAA, the toxic metabolite of EGBE (Udden, 1994), suggest
that certain patients with abnormal hematopoietic systems, such as sickle-cell anemia or
hereditary spherocytosis, are not more sensitive to the hemolytic effects of EGBE than normal
adults. Other studies suggest that the RBCs of children may be pharmacodynamically less
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sensitive to hemolysis than those of adults. RBCs of neonates and children up to 6 months of
age differ from normal adult RBCs in that they are larger and have higher levels of Hb F versus
adult Hb A (Lewis, 1970). Frei et al. (1963) showed that the larger calf erythrocytes containing
Hb F were osmotically more resistant than smaller, adult erythrocytes containing Hb A and
suggested that, as fetal erythrocytes are replaced by postnatal erythrocytes, the total population
of RBCs becomes more susceptible to lysis.
The effect of age on EGBE-induced hematotoxicity was studied in male F344 rats by
Ghanayem et al. (1990, 1987c). These studies also demonstrated the time course for the onset
and resolution of the hematological and histopathologic changes accompanying hemolysis.
Adult male F344 rats were significantly more sensitive to the hemolytic effects of EGBE than
were young (4-5 week) male rats following administration of a single gavage dose of EGBE.
Concurrent metabolism studies found increased blood retention of EGBE metabolite BAA (as
measured by increased Cmax, AUC, and tu) in young rats and that these rats eliminated a
significantly greater proportion of the administered EGBE dose as exhaled CO2 or as urinary
metabolites. The rats also excreted a greater proportion of the EGBE conjugates, glucuronide
and sulfate, in the urine. These researchers suggested that a reduced ability of older rats to
metabolize the toxic metabolite BAA to CO2 along with a diminished ability to excrete BAA in
the urine may explain the age-dependent toxicity of EGBE.
NTP (2000) also found that young mice eliminated BAA 10 times faster than aged mice
following a 1-day exposure to EGBE. This difference was not as apparent after 3 weeks of
exposure, suggesting that factors other than age may be involved (Dill et al., 1998).
Available in vitro information suggests that children are no more and are possibly less
sensitive to the hemolytic effects of BAA than adults. Udden (2002) compared the in vitro
responses of erythrocytes (percent hemolysis and MCV alterations) obtained from hospitalized
adults (n = 29-40) and hospitalized children (n = 25-46) to 0 or 10 mM BAA for 4 hours. BAA
(10 mM) produced comparable significant increases in MCV in both adults (from 87.1 to
89.2 [j,m3; 2%) and children (92.8-95.2 [j,m3; 3%). In the case of hemolysis in response to BAA,
the response was noted as being significantly increased for hospitalized adults (0.54-0.69%;
28%) but not for hospitalized children (0.68-0.75%; 10%).
Relatively minor developmental effects due to maternal toxicity related to hematologic
effects of EGBE exposure were found in studies using rats, mice, and rabbits dosed orally, by
inhalation, or dermally (NTP, 1993; Heindel et al., 1990; Sleet et al., 1989; Wier et al., 1987;
Hardin et al., 1984; Tyl et al., 1984). No teratogenicity was noted in any studies. It can be
concluded that EGBE is not significantly toxic to developing fetuses of laboratory animals.
4.7.2. Possible Gender Differences
Gender differences have been noted in a number of animal and human studies: females
were more susceptible to effects from EGBE exposure. In the NTP (1993) 2-week drinking
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water studies with EGBE, the absolute and relative thymus weights in female F344 rats at the
highest exposure level (265 mg/kg-day) were slightly reduced. In the 13-week studies, male rats
in the highest three dose groups and females in all dose groups suffered mild (males) to moderate
(females) anemia.
Gender differences have also been noted in some studies that observed the hemotoxic
effects of dermal administration of EGBE. Repeated application of EGBE either neat or as a
dilute aqueous solution (occluded) to male or female NZW rabbits at exposure levels of 18, 90,
180, or 360 mg/kg (6 hours/day, nine applications) produced hemoglobinuria in males at
360 mg/kg and in females at 180 or 360 mg/kg (Tyler, 1984). Only female rabbits showed
decreased RBC counts, Hb concentrations, and MCHC along with increased MCH at the highest
treatment level. Recovery was noted following a 14-day observation period.
A number of secondary effects resulting from the hemolytic toxicity of EGBE, such as
effects on the rat liver, kidneys, spleen, bone marrow, and, to a lesser extent, the thymus, were
more pronounced in females. In the NTP drinking water studies (1993), liver lesions in females,
but not males, included cytoplasmic alterations, hepatocellular degeneration, and pigmentation.
These effects were most pronounced in the three highest dose groups.
Carpenter et al. (1956) reported female rats to be more sensitive than males to the
hemolytic effects of EGBE. In dogs, slight increases in erythrocyte osmotic fragility in both the
male and female basenji hybrids were seen, but RBC counts and Hb concentrations were slightly
decreased in the female. Overall, the effects were seen in both genders, but appeared to develop
more slowly in the males. In monkeys, occasional rises in erythrocyte osmotic fragility were
recorded during the exposure period more frequently in the female than in the male.
In the process of studying and comparing the metabolic and cellular basis of EGBE-
induced hemolysis of rat erythrocytes in vitro with human erythrocytes, Ghanayem (1989)
observed that the blood from male and female human volunteers was unaffected by 4-hour
incubations with BAA at concentrations of up to 4.0 mM. At 8 mM, only slight but significant
hemolysis of human blood was observed, with blood from females showing a slightly greater
sensitivity.
The NTP 2-year inhalation bioassay (NTP, 2000; Dill et al., 1998) also reported evidence
of gender specificity in mice and rats, particularly with respect to the elimination of BAA in rats.
Female rats eliminated BAA more slowly from the blood, as indicated by a smaller elimination
rate constant, longer elimination ty2, and larger AUC. In addition, the Cmax of BAA was greater
for females at each concentration and time point. It has been suggested because a smaller
amount of BAA was excreted in the urine of female rats, that higher blood concentrations of
BAA accumulated in the females (Dill et al., 1998). Mouse data from the NTP (2000) study also
suggest a slightly increased hematologic effect among female mice; however, while female mice
tended to have higher blood concentrations of BAA, they excreted more BAA in urine than male
mice.
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5. DOSE-RESPONSE ASSESSMENTS
5.1. INHALATION REFERENCE CONCENTRATION (RFC)
In general, the RfC is an estimate, with uncertainty spanning perhaps an order of
magnitude, of a daily exposure to the human population—including susceptible subgroups—that
is likely to be without an appreciable risk of adverse health effects over a lifetime. It is derived
from a lower confidence limit on the BMD, a NOAEL, a LOAEL, or another suitable point of
departure (POD), with uncertainty/variability factors applied to reflect limitations of the data
used. The RfC is expressed in terms of mg/m3 of exposure to an agent and is derived by a
methodology similar to the RfD. Ideally, studies with the greatest duration of exposure and
conducted via the inhalation route of exposure give the most confidence for derivation of an RfC.
5.1.1. Choice of Principal Study and Critical Effect—with Rationale and Justification
There are no studies reported in which humans have been exposed sub chronically or
chronically to EGBE by the inhalation route of exposure. The animal studies considered for
selection as principal studies include the 14-week and 2-year inhalation studies by NTP (2000) in
rats and mice, the developmental toxicity study by Tyl et al. (1984) in rats and rabbits, the
developmental toxicity study by Nelson et al. (1984) in rats, and the subchronic study by Dodd et
al. (1983) in rats. The NTP (2000) study was selected as the principal study because it was
conducted in two species and provides data for different durations and for more dose groups than
the other studies. The developmental toxicity studies identified effects at doses higher than the
doses associated with the critical effects identified in the NTP (2000) study and were not used for
quantitative purposes. While the subchronic study by Dodd et al. (1983) was well-conducted, the
NTP (2000) study contained more dose groups, more animals per group, and a longer duration of
exposure. Thus, Dodd et al. (1983) was not used for quantitative purposes. Two endpoints from
the NTP (2000) study—the hemolytic endpoint from the 14-week inhalation study and the
hemosiderin deposition endpoint from the 2-year inhalation study—were used for the critical
effect. The hemolytic endpoints in the 1999 EGBE Toxicological Review were used to derive the
reference values (see Section 5.1.5), but were not used to derive the values in this updated
assessment. New MOA information published since the 1999 EGBE Toxicological Review is
included in this document, and this information supports the hemosiderin deposition endpoint as
an important key event in the proposed MOA. A comparison of the NOAELs and LOAELs for
the candidate studies are summarized in Table 5-1.
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Table 5-1. Results of candidate studies
Species
(strain)
Number/
Effect levels (ppm)
Reference
Gender
dose group
Duration/effect
NOAEL
LOAEL
NTP (2000)
Rat (F344)
M
9-10
50
14 wk, hematologic
2 yr, hematologic,
hemosiderin (liver)
31
62.5
31
F
9-10
50
14 wk, hematologic
2 yr, hematologic,
hemosiderin (liver)
-
31
31
NTP (2000)
Mouse
(B6C3F0
M
50
2 yr, histopathology of the
forestomach
—
62.5
50
2 yr, hematologic,
hemosiderin (liver)
62.5
125
F
50
50
2 yr, histopathology of the
forestomach
2 yr, hematologic,
hemosiderin (liver)
-
62.5
62.5
Tyl etal. (1984)
Rat (F344)
F
36
GD 6-15, hematologic
50
100
Nelson et al.
(1984)
Rat (Sprague-
Dawley)
F
15
GD 7-15, hematologic
150
200
Dodd et al.
(1983)
Rat (F344)
M, F
16
13 wk, hematologic
25
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The primary effects of EGBE exposure were hematological effects and were observed in
both species and genders tested. Female rats (NTP, 2000) appeared to be most sensitive among
animals studied. A mild-to-moderate regenerative anemia was observed in females exposed to
all concentrations, with a LOAEL of 31 ppm identified for hematological effects in male and
female rats and no NOAEL. Exposure-related trends were noted for reticulocyte count, RBC
count, MCV, Hb concentration, and Hct. The hematological endpoints were considered for the
derivation of the RfC; however, they presented a number of difficulties. It was not clear which
of the hematological endpoints (changes in RBC count, reticulocyte count, MCV, Hb
concentration, and Hct) observed in EGBE-exposed animals should be used to derive an RfC. In
the case of BMD analysis, the proper benchmark response (BMR) level for the BMD derivation
was uncertain. In addition, while these hematologic effects were observed in both the subchronic
and chronic studies and persisted with exposure duration, they did not progress in severity in the
subchronic-to-chronic study (see Tables 4-3, 4-6, and 5-2). Further, better model fits were
obtained from the BMD analysis of the subchronic study, which used two more exposure
concentrations than the chronic study. For these reasons, the hematologic responses from the
14-week subchronic study were chosen for use in the BMD analyses of this endpoint (see
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1 Appendix C). Selection of the most appropriate hematologic endpoints for use in the BMD
2 analysis also required consideration of EGBE's MOA for hemolysis.
3
Table 5-2. Female and male rat and mouse liver hemosiderin staining
incidence and RBC from subchronic and chronic EGBE inhalation studies
Effect/time
Control
31 ppm
62.5 ppm
125 ppm
250 ppm
500 ppm
Female rat
Hemosiderin
14 wk
2 yr
0/10
15/50
0/10
19/50
10/10°
36/50b
10/10°
47/50°
9/9°
NT
5/5°
NT
RBC count3
14 wk
1 yr
8.48 ±0.05
7.81 ±0.05
8.08 ± 0.07° (95)
NT
7.70 ± 0.08° (91)
7.42 ± 0.06° (95)
6.91 ±0.05° (82)
6.75 ± 0.05° (86)
6.07 ± 0.04° (72)
NT
4.77 ±0.15° (56)
NT
Male rat
Hemosiderin
14 wk
2 yr
0/10
23/50
0/10
30/50
0/10
34/50b
7/10°
42/50b
10/10°
NT
10/10°
NT
RBC count3
14 wk
1 yr
9.05 ±0.08
8.88 ±0.08
8.71 ±0.14b (96)
NT
8.91 ±0.06 (94)
8.39 ± 0.15° (94)
8.01 ±0.08° (89)
7.43 ± 0.20° (84)
7.10 ±0.07° (78)
NT
5.97 ± 0.05° (66)
NT
Female mouse
Hemosiderin
14 wk
2 yr
0/10
0/50
0/10
NT
0/10
5/50°
0/10
25/493
10/103
44/503
6/63
NT
RBC count3
14 wk
1 yr
9.72 ±0.05
9.32 ±0.09
9.55 ± 0.06b (98)
NT
9.51 ± 0.06b (98)
9.14 ±0.08 (98)
9.18 ±0.05° (94)
8.50 ±0.12° (91)
8.57 ± 0.06° (88)
8.08 ± 0.09° (87)
7.35 ± 0.07° (76)
NT
Male mouse
Hemosiderin
14 wk
2 yr
0/10
0/50
0/10
NT
0/10
0/50
0/10
8/493
0/10
30/493
6/63
NT
RBC count3
14 wk
1 yr
9.71 ±0.22
9.58 ±0.07
10.04 ±0.08
(103) NT
9.77 ±0.10 (101)
9.73 ±0.49 (102)
9.47 ± 0.06b (98)
9.36 ± 0.32b (98)
8.90 ± 0.07° (92)
8.33 ±0.10° (87)
7.21 ±0.23° (74)
NT
3Mean ± standard error with RBC counts expressed as 106/|iL: percent of control in parentheses.
Statistically significant difference, p < 0.05.
Statistically significant difference,/) <0.01.
NT = not tested
Source: NTP (2000).
4
5 The suggested MOA of EGBE hemolysis is based on data indicating that BAA, an
6 oxidative metabolite of EGBE and the first hypothesized event in the MOA, is likely to be the
7 causative agent in hemolysis (Ghanayem et al., 1990, 1987b; Carpenter et al., 1956). The second
8 event in the MOA is erythrocyte swelling and cell lysis, which is believed to be preceded by an
9 increase in the osmotic fragility and a loss of deformability of the erythrocyte (Udden, 1995b,
10 1994; Udden and Patton, 1994; Ghanayem, 1989). This results in decreased values for RBC
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count, Hb, and Hct and in response, an increase in the production of immature RBCs
(reticulocytes) by the bone marrow.
Although changes in reticulocyte and nucleated erythrocyte counts sometimes represent
the largest measurable differences between exposed animals and unexposed control animals, this
parameter is highly variable and does not always exhibit a dose-dependent trend (NTP, 2000,
1993). While these endpoints can be indirect markers of RBC lysis, they are governed by
multiple feedback control processes that can be both very sensitive and variable. Therefore, a
change in reticulocyte or nucleated erythrocyte count is not considered a suitable endpoint for
deriving the RfC or RfD.
Until more is known about the molecular interaction between BAA and specific cellular
molecules, it is reasonable to assume that changes in MCV and RBC count are measurements of
precursor events in response to both oral and inhalation EGBE exposure. Therefore, dose-
response information on MCV and RBC count are key endpoints used in the BMD analyses and
were considered for derivation of the RfC and RfD for EGBE.
While the toxicokinetic data suggest that MCV should theoretically be the earlier
indicator of hemolytic effects from EGBE exposure, recent studies suggest that the relationship
between the rate of MCV increase and RBC count decrease may not be consistent across
exposure protocols. In the gavage studies of Ghanayem et al. (1987c) and the inhalation studies
of NTP (2000), Hct, a measure of RBC volume relative to blood volume, tended to decrease
along with RBC count and Hb at all exposure levels for which a hematologic effect was
observed. However, Hct did not change as RBC count and Hb decreased following drinking
water exposures (NTP, 1993). Thus, the loss of erythrocytes in the drinking water studies
(reduced RBC count) may have been offset by a concurrent increase in the size of the individual
cells (increased MCV). This was not the case in the gavage and inhalation studies. For these
reasons, greater weight is given to reduced RBC count, as opposed to increased MCV.
While the hemolytic effects appeared to be among the earliest effects from EGBE
exposure, the hemosiderin deposition endpoint was selected as the critical effect. This effect was
found to occur in both species and genders of animals tested, with rats being the more sensitive
species; the effect also occurred in the 14-week subchronic NTP inhalation study. The suggested
MOA of EGBE-induced liver effects is based on the observation that the hemolytic effects led to
compensatory erythropoiesis and significant increases in blood degradation products, including
an increased accumulation of hemosiderin in the liver Kupffer cells of EGBE-exposed animals.
The hemosiderin accumulation seen in the Kupffer cells was found to increase in severity with
increasing dose and exposure duration (Table 5-2), unlike the hemolytic endpoints, such as
decreased Hct, which did not progress from 3 to 12 months (Table 4-6). Thus, hemosiderin
deposition in Kupffer cells in the rat liver is believed to be a sequela to the hematologic effects.
Because of the progression of this effect with chronic exposure, hemosiderin is deemed to be the
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most sensitive effect. ANOAEL was not identified, while a LOAEL of 31 ppm was identified in
both male and female rats.
The 2-year chronic inhalation study by the NTP (2000) observed forestomach ulcers in
female mice at all exposure levels, but this effect has not been observed in any other species,
including mice exposed orally to EGBE (NTP, 1993). Though the incidence of this lesion
increased with exposure, severity of the lesion did not increase with increasing dose. While this
effect was not considered a critical effect for the derivation of an RfC, Appendix C contains the
BMD analysis for this endpoint for comparison purposes.
5.1.2. Methods of Analysis—Including Models (PBPK, BMD, etc.)
PODs for the RfC derivation in terms of the human equivalent concentrations (HECs)
have been calculated via the application of PBPK modeling and the use of internal dose metrics
published by Dill et al. (1998) to obtain NOAEL and benchmark concentration, 95% lower
bound (BMCL) estimates. Details of the various POD derivation approaches that were
considered are provided in Appendix C. The selected approach is described in Section 5.1.2.1
below.
The PBPK models developed for EGBE are briefly summarized in Table 5-3. Shyr et al.
(1993) and Johanson (1986) do not address BAA distribution and are only parameterized for
humans and rats, respectively. In the 1999 EGBE Toxicological Review, the model described by
Lee et al. (1998) is the most appropriate model for the estimation of rat and mouse internal doses
following inhalation exposure. Since the 1999 Toxicological Review, Corley et al. (2005a)
published a revision to the Lee et al. (1998) model for rats and mice where several assumptions
used by Lee et al. (1998) were replaced with measured values (e.g., protein binding, partition
coefficients, metabolism rate constants for multiple pathways, and renal clearance) as a function
of species, gender, and age. As is described in Appendix C, that model was used to estimate the
Cmax of BAA in rat blood for the derivation of reference values from hematologic endpoints, and
the human PBPK model of Corley et al. (1997, 1994) was used to obtain estimates of human
inhalation exposure concentrations associated with the rat BAA Cmax estimates. In the analysis
of the hemosiderin effect in rats and mice described in Section 5.1.2.1 below, the human PBPK
model of Corley et al. (1997, 1994) was used to obtain estimates of human inhalation exposure
concentrations associated with the BMDs derived from rat BAA AUC levels reported by Dill et
al. (1998).6 Established U.S. EPA (2006c) methods and procedures were used to review, select
and apply these chosen PBPK models.7
6The basic components of the Corley model are summarized in Appendix B.
7EPA notes that a review of the PBPK models was conducted prior to their use in the 1999 EGBE toxicological
review.
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Table 5-3. Summary of PBPK models
Model
Species
Routes of exposure
Comments
Johanson (1986)
Human
Inhalation
BAA not addressed
Shyretal. (1993)
Rat
Inhalation, oral, dermal
BAA excretion
Corley et al. (1997,
1994)
Rat and human
Inhalation, oral, dermal
BAA distribution and excretion; male rats
only
Lee et al. (1998)
Rat and mouse
Inhalation
BAA distribution and excretion; males and
females
Corley et al. (2005a)
Rat and mouse
Inhalation, oral,
dermal, i.p., i.v.
Age-dependent BAA distribution, metabolism
and excretion, males and females
Franks et al. (2006)
Human
Inhalation and dermal
Extended Corley et al. (1997) model to
include bladder compartment for human
biomonitoring studies
5.1.2.1. BMD Approach Applied to Hemosiderin Staining Data
For the purposes of deriving an RfC for EGBE, hemosiderin staining data were evaluated
in male and female rats from the 2-year chronic study by NTP (2000). The current BMD
technical guidelines (U.S. EPA, 2000b) suggest the use of 10% extra risk as a BMR level for
quantal data, as this is at or near the limit of sensitivity in most cancer bioassays and in some
noncancer bioassays as well. Because the hemosiderin staining endpoint was observed in control
animals and a 10% increase in incidence was within the observable range of the data, 10% extra
risk was considered an appropriate BMR and a BMCLio an appropriate POD for derivation of
the RfC (U.S. EPA, 2000b, 1995). All BMD assessments in this review were performed using
U.S. EPA benchmark dose software (BMDS) version 1.4.1. Graphical figures and text output
files for selected benchmark concentration (BMC) analyses are provided in Appendix C.
The AUC was selected as the appropriate dose metric due to the nature of the endpoint,
hemosiderin deposition. This endpoint increased in severity with increased duration (subchronic
to chronic) and is believed to be the result of the cumulative exposure to EGBE as opposed to a
peak event. Table 5-4 reports AUC BAA blood concentrations measured at 12 months8
published by Dill et al. (1998) in both genders of B6C3Fi mice and F344 rats exposed to the
same concentrations used in the NTP (2000) chronic studies of these test animals.
8Dill et al. (1998) also reported 18 month data, but due to the smaller number of animals and higher variability in
this data, the 12 month data were used for the purposes of this analysis.
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Table 5-4. AUC BAA blood concentrations measured at 12 months in both
genders of B6C3Fi mice and F344 rats
Exposure concentration (ppm)
Gender
AUCbaa (jtmol-hr/L)3
n
Mean
SE
Rats
31.2
Male
7
358.3
16.6
Female
5
638.8
18.7
62.5
Male
6
973.0
86.2
Female
9
1,128.9
50.9
125
Male
9
2,225.6
71.1
Female
12
3,461.8
154.8
Mice
62.5
Male
10
1,206.6
205.6
Female
12
1,863.6
112.4
125
Male
9
2,819.8
685.1
Female
6
5,451.6
508.9
250
Male
10
17,951.5
1,770.4
Female
11
18,297.1
609.7
"Authors reported AUC values in terms of (ig/min and g, which were converted to units consistent with the PBPK
model of (imol-hour/L by dividing by 60 minute/hour and 132.16 g/mol and multiplying by 1,060 g/L.
Source: Dill et al. (1998).
The fit statistics and BMC information derived from the dichotomous models available in
the BMD software as applied to the male and female rat hemosiderin staining data versus AUC
BAA are shown in Table 5-5. All models were fit using restrictions and option settings suggested
in the U.S. EPABMD technical guidance document (U.S. EPA, 2000b). The best model fit to
these data, as determined by visual inspection, examination of low dose model fit (i.e., scaled
residual for the dose group closest to the BMD), and comparison of overall fit (i.e., Akaike
information criterion [AIC] values), was obtained using a multistage model (1st degree) for the
male response data and a Log-Logistic model for the female response data. The male rat BMCio
was 196 [j,mol-hour/L and the BMCLio was determined to be 133 [j,mol-hour/L, using the 95%
lower confidence limit of the dose-response curve expressed in terms of the AUC for BAA in
blood. The BMC io and BMCLio values for the female rat were determined to be 425 and
244 [j,mol-hour/L, respectively. Assuming continuous exposure (24 hour/day), the Corley et al.
(1997) PBPK model was used to back-calculate HECs of 3.4 ppm (16 mg/m3) from the male rat
data and 4.9 ppm (24 mg/m3) from the female rat data.
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Table 5-5. Comparison of BMC/BMCL values for male and female rat liver
hemosiderin staining data from inhalation chronic study using measured
blood AUC (12 months) of the EGBE metabolite BAA as a common dose
metric
Model
BMCio (jtmol-hr/L)
BMCLio (nmol-hr/L)
/7-Value
AICa
Scaled residualb
Male rats
Multistage-lst degree0
196.252
133.141
0.8680
247.234
0.441
Gamma0
196.253
133.141
0.8680
247.234
0.441
Logistic
259.296
192.773
0.7692
247.476
0.526
Log-logistic
166.376
69.3279
0.5623
249.283
0.313
Probit
271.525
205.882
0.7450
247.54
0.517
Log-probit
368.336
241.992
0.6309
247.876
0.765
Weibull0
196.253
133.141
0.8680
247.234
0.441
Female rats
Multistage-lst degree
122.166
214.555
0.0698
218.868
-1.945
Gamma
316.635
134.02
0.0554
219.229
-1.238
Logistic
273.693
221.689
0.0993
218.188
-1.294
Log-logisticc
424.527
243.69
0.1533
217.526
-0.896
Probit
291.017
241.206
0.0683
218.985
-1.260
Log-probit
427.728
248.683
0.1238
217.884
-0.965
Weibull
266.515
130.801
0.0454
219.58
-1.377
AIC = -2L + 2P, where L is the log-likelihood at the maximum likelihood estimates for the parameters and P is the
number of model degrees of freedom (usually the number of parameters estimated).
b%2 residual (measure of how model predicted response deviates from the actual data) for the dose group closest to
the BMC scaled by an estimate of its SD. Provides a comparative measure of model fit near the BMC. Residuals
that exceed 2 in absolute value should cause one to question model fit in this region.
°Model choice based on adequate p-valuc (>0.1), visual inspection, low AIC, and low (absolute) scaled residual.
The Multistage (1st degree) is referred to as the chosen model for male rats, though equivalent fit was obtained by
the restricted Gamma and Weibull models.
Likewise, the fit statistics and BMC information for male and female mouse hemosiderin
staining data versus AUC BAA are shown in Table 5-6. All models were fit using restrictions
and option settings suggested in the U.S. EPABMD technical guidance document (U.S. EPA,
2000b). The best model fit to these data, as determined by visual inspection, examination of low
dose model fit (i.e., scaled residual for the dose group closest to the BMD), and comparison of
overall fit (i.e., AIC values), was obtained using a log-probit model for both the male and female
response data. The male mouse BMCio was 3,077 [j,mol-hour/L and the BMCLio was
determined to be 2,448 [^mol-hour/L using the 95% lower confidence limit of the dose-response
curve expressed in terms of the AUC for BAA in blood. The BMCio and BMCLio values for the
female mouse were determined to be 1,735 and 1,322 [j,mol-hour/L, respectively. Assuming
continuous exposure (24 hour/day), the Corley et al. (1997) PBPK model described in Appendix
B was used to back-calculate HECs of 36 ppm (174 mg/m3) from the male mouse data and
20 ppm (97 mg/m3) from the female mouse data.
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Table 5-6. Comparison of BMC/BMCL values for male and female mouse
liver hemosiderin staining data from inhalation chronic study using
measured blood AUC (12 months) of the EGBE metabolite BAA as a
common dose metric
Model
BMC10
(nmol-hr/L)
BMCL10
(nmol-hr/L)
/7-Value
AICa
Scaled residualb
Male mice
Multistage-1st
degree
2,100.07
1,613.9
0.3067
117.571
-1.766
Gamma
2,725.35
1,702.27
0.1452
118.559
1.358
Logistic
6,605.45
5,333.72
0.0022
127.326
2.789
Log-logistic
2,616.51
1,628.48
0.1882
118.02
1.193
Probit
5,917.06
4,825.09
0.0031
126.405
2.734
Log-probitc
3,076.8
2,448.3
0.1290
116.614
1.946
Weibull
2,689.76
1,687.09
0.1445
118.712
-1.448
Female mice
Multistage-1st
degree
946.491
769.879
0.3680
142.669
-1.583
Gamma
1,402.92
818.367
0.3420
143.288
-0.817
Logistic
2,897.15
2,341.03
0.0002
162.338
-0.942
Log-logistic
1,705.75
1,121.43
0.8223
141.501
-0.343
Probit
2,860.03
2,364.52
0.0002
161.681
-0.829
Log-probitc
1,734.53
1,322.06
0.8237
141.498
-0.315
Weibull
1,282.82
804.234
0.2958
143.631
-0.988
AIC = -2L + 2P, where L is the log-likelihood at the maximum likelihood estimates for the parameters and P is the
number of model degrees of freedom (usually the number of parameters estimated).
b%2 residual (measure of how model predicted response deviates from the actual data) for the dose group closest to
the BMC scaled by an estimate of its SD. Provides a comparative measure of model fit near the BMC. Residuals
that exceed 2 in absolute value should cause one to question model fit in this region.
°Model choice based on adequate p-valuc (>0.1), visual inspection, low AIC, and low (absolute) scaled residual.
The Log-probit model provides a slightly better fit than other models for both genders.
5.1.2.2. Selection of the POD
Consideration of the available data has led to the selection of the 2-year inhalation study
(NTP, 2000) and increased hemosiderin staining in the liver of male F344 rats as the principal
study and critical effect for deriving the chronic RfC for EGBE. This is a high-quality study and,
when coupled with information on the MO A, U.S. EPA concluded that this is a precursor to an
adverse effect and is appropriate for use in deriving the RfC. ABMCLio of 133 [j,mol-hour/L for
hemosiderin staining in liver of male rats chronically exposed to EGBE (NTP, 2000) was used as
the POD to calculate the RfC. A human PBPK model (Corley et al., 1997) was used to back-
calculate to a HEC of 16 mg/m3 (3.4 ppm) for the BMCLrec-
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5.1.3. RfC Derivation—Including Application of Uncertainty Factors (UFs)
UFs are applied to account for recognized uncertainties in extrapolation from
experimental conditions to the assumed human scenario (e.g., chronic exposure over a lifetime
when subchronic studies are used for deriving a toxicity value). The application of UFs may
1/2
include the use of a partial UF of 10 (3.2) based on the assumption that the actual values for
the UFs are log-normally distributed. Application of these factors in the assessments is such that,
when a single partial UF is applied, the factor is rounded to 3—for instance, the total factor for
individual UFs of 3 and 10 would be 30 (3 x 10). When two partial UFs are evoked, however,
they are not rounded, such that a UF of 3, 3, and 10 would result in a total uncertainty factor
of 100 (actually 101/2 x 101/2 x 101) (U.S. EPA, 2002). UFs applied for this RfC assessment and
the justification for their use are as follows.
A factor of 10 was selected to account for the uncertainty associated with the variability
of the human response (UFH) to the effects of EGBE. Potentially susceptible subpopulations
include individuals with enhanced metabolism or decreased excretion of BAA and individuals
whose RBC membranes are more susceptible to the lysis caused by BAA, the precursor step to
developing hemosiderin staining in the liver. Human in vitro studies suggest that the elderly and
patients with fragile RBCs would not be more sensitive to the hemolytic effects of EGBE than
normal adults. Laboratory animal studies suggest that older animals are more sensitive than
neonates and that females are more sensitive than males. While developmental studies do not
reveal increased susceptibility in infants, none of the developmental studies examined fetal or
infant blood for signs of effects from prenatal exposure to EGBE. Additionally, human
responses to EGBE have not been observed under a broad range of exposure conditions (e.g.,
repeated or long-term exposures) and potentially sensitive subjects (e.g., individuals predisposed
to hemolytic anemia or infants).
A factor of 1 was selected to account for the uncertainty associated with interspecies
variability resulting from toxicodynamic and toxicokinetic differences between animals and
humans (UFA). Traditionally, these components (toxicodynamic and toxicokinetic) are
individually represented by partial UFs of 3 for a total UF of 10 in the absence of chemical-
specific information; thus, application of a full UF of 10 would depend on two areas of
uncertainty (i.e., toxicokinetic and toxicodynamic uncertainties). In this assessment, the
toxicokinetic uncertainty is addressed by the determination of an HEC, using a combination of
measured internal blood levels in the test animals and PBPK modeling. A value of 1 was
selected for the toxicokinetic portion of the UFA. Regarding toxicodynamics, in vivo (Carpenter
et al., 1956) and in vitro (Udden, 2002; Udden and Patton, 1994; Ghanayem and Sullivan, 1993)
studies indicate that humans may be significantly less sensitive than rats to the hematological
effects of EGBE. Because epidemiologic studies are often limited in their ability to explore
outcomes related to workplace or environmental exposures, it is typically impossible to rule out
the relevance of an effect seen in a particular rodent tissue unless there is detailed mechanistic
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information on why humans would not be affected (NRC, 2008; IARC, 2006). Therefore, the in
vivo human response to EGBE cannot be accurately determined without some degree of
speculation. For this reason, a value of 1 was selected for the toxicodynamic portion of the UFA.
A factor to account for extrapolation from subchronic to chronic exposure (UFS) was not
needed because the RfC was derived from a chronic inhalation study.
A factor to account for the extrapolation from a LOAEL to a NOAEL (UFL) was not
applied because the current approach is to address this extrapolation as one of the considerations
in selecting a benchmark response (BMR) for BMD modeling. In this case, EPA concluded a
10% increase in hemosiderin staining, indicating a precursor to an adverse effect, is appropriate
for use in deriving the RfC under the assumption that it represents a minimal biologically
significant change.
A factor of 1 was selected to account for deficiencies in the database (UFD). Chronic and
subchronic studies are available for two species (rats and mice), and several reproductive and
developmental studies, including a two-generation reproductive toxicity study. There are also
limited human studies available following short-term inhalation exposure.
A total UF of 10 (10 for UFH, 1 for UFA, and 1 for UFD) was used in the derivation of the
RfC. The combined PBPK and BMC modeling method using hemosiderin as an endpoint was
used to derive the RfC. In addition, MOA information was used to inform the choice of the
critical effect. The RfC for EGBE based on hemosiderin deposition in the liver was calculated as
follows:
RfC = BMCLrec - UF
= 16 mg/m3 10
= 1.6 mg/m3
Thus, the RfC is 1.6 mg/m3.
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5.1.4. RfC Comparison Information
For comparison purposes, Figure 5-1 presents the POD, applied UFs, and derived
reference values, including the RfC, for the effect endpoints discussed. BMC modeling was
done using U.S. EPABMDS version 1.4.1 (U.S. EPA, 2000b), and results are provided in
Appendix C. This comparison is intended to provide information on alternative endpoints
associated with EGBE exposure. The selected RfC value is circled; BMC analyses are provided
in Appendix C.
~ PODhe ~ UFh
UFa
m
uf.
RfC
1000
100 -
10 -
0.1
NTP (2000); 2 yr
male and female
rat study;
hematological
effects; LOAEL
NTP (2000);
3 mo female rat
study; RBC
effects;
BMCLos(HEC)
NTP (2000); 2 yr
female rat study;
hemosiderin
staining;
BMCL10(HEC)
NTP (2000); 2 yr
male rat study;
hemosiderin
staining;
BMCL10(HEC)
NTP (2000); 2 yr
female mouse
study;
hemosiderin
staining;
BMCL10(HEC)
NTP (2000); 2 yr
male mouse
study;
hemosiderin
staining;
BMCL10(HEC)
NTP (2000); 2 yr
female mouse
study;
forestomach
hyperplasia;
BMCL10(HEC)
Figure 5-1. PODs for selected endpoints with corresponding applied UFs
and derived RfC.
Figure 5-1 shows PODs and comparison reference values (including the RfC) that could
be derived from the various endpoints to allow a comparison with the chosen critical effect and
the resultant RfC for the critical effect. Hemolytic effects and effects related to hemolysis (i.e.,
hemosiderin deposition) are the most sensitive endpoints for identification of a NOAEL or a
BMCL in the subchronic and chronic studies available; these endpoints have been considered as
the critical endpoint for derivation of the RfC. The BMC Los for RBC count changes in female
rats was 133 [iM, using Cmax at 3 months, and was converted to an inhalation HEC (BMCLHec)
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of 225 mg/m3 using the U.S. EPA model. Though adequate model fit per U.S. EPABMD
technical guidance (U.S. EPA, 2000b) could not be obtained for the NTP (2000) 3-month male
rat RBC response, BMCs derived for hemolytic endpoints in male rats of this study were
approximately twofold higher than for female rats (data not shown). For the hemosiderin
endpoint, both male and female data from the chronic study were considered. The BMCLio for
hemosiderin staining in male rats was 133 [xM-hour/L using the AUC for BAA in arterial blood
at 12 months and was converted to a BMCLHec of 16 mg/m3 using the Corley et al. (1997, 1994)
human PBPK model. The BMCLio for hemosiderin staining in female rats was 244 [xM-hour/L
using the AUC for BAA in arterial blood at 12 months and was converted to a BMCLrec of
24 mg/m3 using the Corley et al (1997, 1994) human PBPK model.
5.1.5. Previous Inhalation Assessment
The previous IRIS assessment for EGBE was entered into the database on December 31,
1999; it contains an inhalation RfC of 13 mg/m3. The RfC was based on the BMCos(hec) of
380 mg/m3 for changes in RBC count in female F344 rats following a 14-week inhalation
exposure (NTP, 1998). A total UF of 30 was used to account for human variability and
extrapolation from an adverse effect level.
5.2. ORAL REFERENCE DOSE (RfD)
In general, the RfD is an estimate, with uncertainty spanning perhaps an order of
magnitude, of a daily exposure to the human population—including susceptible subgroups—that
is likely to be without an appreciable risk of adverse health effects over a lifetime. It is derived
from a statistical benchmark dose, 95% lower bound (BMDL), a NOAEL, a LOAEL, or another
suitable POD, with uncertainty/variability factors applied to reflect limitations of the data used.
The RfD is expressed in terms of mg/kg-day of exposure to an agent and is derived by a similar
methodology to the RfC. Ideally, studies with the greatest duration of exposure and conducted
via the oral route of exposure give the most confidence for derivation of an RfD. The database
of oral studies for EGBE is more limited than the database of inhalation studies. For this reason,
a PBPK model for EGBE has been applied to the inhalation data for derivation of an RfD.
5.2.1. Choice of Principal Study and Critical Effect—with Rationale and Justification
No studies have been reported in which humans have been exposed subchronically or
chronically to EGBE by the oral route of exposure, and thus would be suitable for derivation of
an oral RfD. No chronic oral laboratory animal studies are currently available for EGBE. The
results of the only two available subchronic 91-day drinking water studies in rats and mice (NTP,
1993) are summarized in Table 5-7.
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Table 5-7. Subchronic 91-day drinking water studies in rats and mice
Reference
Species
(strain)
Gender
Animals/dose
Effect
Effect levels
(mg/kg-d)
NOAEL
LOAEL
NTP (1993)
Rat
(F344)
M
10
Hepatocellular changes
-
54.9a
F
10
Hematological
-
58.6a
NTP (1993)
Mouse
(B6C3F0
M
10
Body weight
223
553b
F
10
Body weight
370
676b
aDoses were calculated using water consumption rates and body weights measured during the last week of exposure
and, therefore, differ slightly from those reported by the authors and presented in Section 4.2.
bThe LOAEL in mice was based on reduced body weight and body weight gain.
Based on a comparison of NOAELs and LOAELs for hematological and liver effects, rats
are clearly more sensitive to the noncancer effects of EGBE than mice. As discussed in
Section 4.2, hematologic and hepatocellular changes were noted in both rat genders. In females,
both hematologic and hepatocellular changes were noted at the low-dose level (58.6 mg/kg-day,
using water consumption rates and body weights measured during the last week of exposure).
Only hepatocellular cytoplasmic changes were observed in low-dose male rats (54.9 mg/kg-day,
using water consumption rates and body weights measured during the last week of exposure).
In the female rat study (NTP, 1993), groups of 10 F344 rats were exposed to 0, 82, 151,
304, 363, and 470 mg/kg-day EGBE via drinking water for 13 weeks. Body and organ weights
were measured, and clinical, hematological, gross, and histopathologic examinations were
conducted. Decreases in body weight were observed in female rats exposed to the two highest
dose levels. Hematologic changes were observed at all dose levels after 13 weeks and were
indicative of mild-to-moderate anemia. These changes included reduced RBC count, Hb, and
Hct, as well as increased reticulocyte count and MCV. Liver hemosiderin pigmentation was
noted in the cytoplasm of Kupffer cells in both genders. In females it was noted in 0/10 controls
in 82 mg/kg-day treated animals, 2/10 with a severity grade of 1 (minimal) at 151 mg/kg-day,
and 10/10 in the three highest dose levels, with the severities noted as increasing from a
numerical grade of 1.2 in the 304 mg/kg-day group to 1.9 in both of the upper two dose groups.
In males the pigmentation was noted in animals exposed to the highest dose only (452 mg/kg-
day) at an incidence of 7/10 and a severity rating of 1 (minimal). No hepatic pigmentation was
reported in the mice exposed for 13 weeks.
Hematological effects leading to RBC lysis and organ accumulation of iron in the form of
hemosiderin accumulation appear to be the most sensitive of the effects caused by EGBE in
laboratory animals. Less clear, however, is whether one of the hematological endpoints (changes
in RBC count, reticulocyte count, MCV, HCT, and Hb) or incidence of hemosiderin pigmentation
observed in EGBE-exposed animals is the most appropriate basis for an RfC/RfD (see
Section 5.1.1).
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ABMD analysis has also been performed on the hemosiderin pigmentation endpoint
observed in the NTP (2000) chronic EGBE inhalation study (Section 5.1.2), and PBPK models
have been applied to extrapolate this POD to a human equivalent oral exposure (Section 5.2.2).
As discussed, hematological endpoints indicative of hemolysis do not progress with increasing
duration of inhalation exposure, whereas the incidence of hemosiderin pigmentation did progress
considerably with chronic exposure (Table 5-2). Furthermore, not enough is known about the
mechanism of action of EGBE to make a biologically-informed determination regarding which
hematological endpoints—changes in RBC count, reticulocyte count, MCV, Hct, and Hb—
observed in EGBE-exposed animals should be used to derive an RfC and, in the case of BMD
analysis, what a proper BMR level should be for the BMD derivation. Finally, hematologic
effects signified by changes in RBC count, reticulocyte count, MCV, HCT, and Hb are
considered precursor effects to the pathological finding of hemosiderin deposition (Section 4.5).
For these reasons, and because the hemolytic effects of EGBE appear to be consistent between
oral and inhalation routes of exposure, an RfD has been derived via the application of a PBPK
model to perform a route-to-route extrapolation from the incidence of hemosiderin pigmentation
observed in the NTP (2000) chronic inhalation study of EGBE. BMD/NOAEL analyses of
hematologic endpoints and hemosiderin pigmentation observed in the oral NTP (1993)
subchronic study are provided below for comparison purposes.
Another issue that needs to be addressed with respect to the NTP (1993) study is the lack
of reported forestomach lesions at even the highest drinking water doses administered in this
study relative to the considerable incidence of forestomach hyperplasia and ulceration observed
at all exposure levels in the NTP (2000) subchronic and chronic inhalation studies of mice. This
is difficult to explain, considering the lowest dose in the subchronic drinking water study is
predicted to result in similar, if not higher, Cmax blood levels of the EGBE metabolite BAA (the
presumed irritant) compared to the lowest exposure concentration in the NTP (2000) subchronic
inhalation study. There is no clear explanation. It has been suggested that oral nonbolus dosing
of EGBE does not result in high enough local concentrations of EGBE and BAA (Poet et al.,
2003). Studies with other forestomach carcinogens that are not mutagenic have demonstrated
that forestomach effects are dependent not only on the dose but also on the chemical
concentration in the dosing solution (Ghanayem et al., 1985), and other effects of EGBE appear
to be highly dependent on the concentration attained (Ghanayem et al., 2001, 2000; Long et al.,
2000; Nyska et al., 1999a, b). In addition, first-pass liver metabolism of orally administered
EGBE may affect the extent to which EGBE reaches the forestomach via the route that has been
proposed following i.v. injection: distribution to salivary glands followed by the swallowing of
EGBE-laden saliva (Poet et al., 2003; Green et al., 2002). In any case, since forestomach
irritation was not reported in rats or mice in the NTP (1993) drinking water study, this is not
considered a sensitive endpoint, and route-to-route extrapolation of this endpoint from inhalation
data is not considered appropriate for use in the RfD derivation.
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5.2.2. Methods of Analysis—Including Models (PBPK, BMD, etc.)
PODs for the RfD derivation in terms of the human equivalent doses (HEDs) have been
calculated via the application of PBPK modeling to NOAEL and BMDL estimates. Details of
the various POD derivation approaches that were considered are provided in Appendix C. The
selected approach is described below.
5.2.2.1. BMD Approach Applied to Hemosiderin Endpoint
Due to the limited oral database, U.S. EPA concluded that a route-to-route extrapolation
will increase the confidence in the determination of the RfD POD. Inhalation studies considered
for derivation of the RfC are used to supplement the oral database using the route-to-route
extrapolation as described below.
5.2.2.2. Route-to-Route Extrapolation from Inhalation Data
A route-to-route extrapolation was performed on the data used in the derivation of the
RfC from the NTP (2000) chronic inhalation study because of the lack of a chronic oral study for
EGBE. As with the animal to human extrapolation used in the development of the RfC, the dose
metric used for interspecies (rat to human) and route-to-route (inhalation to oral) extrapolation
was the AUC of BAA at 12 months in arterial blood. This dose metric was used for dose-
response modeling of chronic inhalation data (Section 5.1.2) to derive the POD of 133 [j,mol-
hour/L, expressed as a BMDL. The BMDL was then back-calculated using the human PBPK
model (Corley et al., 1997, 1994) to obtain an equivalent human oral drinking water dose
(BMDLhed) of 1.4 mg/kg-day. As for the alternative HED estimations described in Section C.2
of Appendix C, a simplifying assumption was used that the entire dose of drinking water EGBE
was consumed over a 12-hour period each day.
5.2.2.3. Selection of the POD
The BMCL chosen for the RfC is used to determine the POD for the RfD. This value is
based on a more comprehensive chronic data set and is below the range of estimates from
available oral data of shorter duration of exposure. Hemosiderin deposition in male rat liver is
the critical effect chosen for derivation of the RfC. New MO A information (see Section 4.6.3.1)
supports the hemosiderin deposition endpoint as an important key event in the proposed MOA.
The BMCL for the RfC (AUC of 133 [xM-hour/L BAA in arterial blood at 12 months) is
converted using the Corley et al. (1997, 1994) model to an oral HED (BMDLHed) of 1.4 mg/kg-
day. This extrapolated oral value is consistent with and slightly lower than the LOAELred of
18 mg/kg-day and the BMDLred of 6.8 mg/kg-day estimated from the subchronic oral (NTP,
1993) study (see Appendix C).
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5.2.3. RfD Derivation—Including Application of Uncertainty Factors (UFs)
UFs are applied to account for recognized uncertainties in extrapolation from
experimental conditions to the assumed human scenario (e.g., chronic exposure over a lifetime
when subchronic studies are used for deriving a toxicity value). The application of UFs may
1/2
include the use of a partial UF of 10 (3.2) based on the assumption that the actual values for
the UFs are log-normally distributed. Application of these factors in the assessment is such that,
when a single partial UF is applied, the factor is rounded to 3 (e.g., the total factor for individual
UFs of 3 and 10 would be 30 [3 x 10]). When two partial UFs are evoked, however, they are not
rounded, such that a UF of 3, 3, and 10 would result in a total uncertainty of 100 (actually
101/2 x 101/2 x 101). UFs applied for this RfD assessment and the justification for their use
follow.
A factor of 10 was selected to account for the uncertainty associated with the variability
of the human response (UFH) to the effects of EGBE. Potentially susceptible subpopulations
include individuals with enhanced metabolism or decreased excretion of BAA and individuals
whose RBC membranes are more susceptible to the lysis caused by BAA, the precursor step to
developing hemosiderin staining in the liver. Human in vitro studies suggest that the elderly and
patients with fragile RBCs would not be more sensitive to the hemolytic effects of EGBE than
normal adults. Laboratory animal studies suggest that older animals are more sensitive than
neonates and that females are more sensitive than males. While developmental studies do not
reveal increased susceptibility in infants, none of the developmental studies examined fetal or
infant blood for signs of effects from prenatal exposure to EGBE. Additionally, human
responses to EGBE have not been observed under a broad range of exposure conditions (e.g.,
repeated or long-term exposures) and potentially sensitive subjects (e.g., individuals predisposed
to hemolytic anemia or infants).
A factor of 1 was selected to account for the uncertainty associated with interspecies
variability resulting from toxicodynamic and toxicokinetic differences between animals and
humans (UFA). Traditionally, these components (toxicodynamic and toxicokinetic) are
individually represented by partial UFs of 3 for a total UF of 10 in the absence of chemical-
specific information; thus, application of a full UF of 10 would depend on two areas of
uncertainty (i.e., toxicokinetic and toxicodynamic uncertainties). In this assessment, the
toxicokinetic uncertainty is addressed by the determination of an HEC, using a combination of
measured internal blood levels in the test animals and PBPK modeling. A value of 1 was
selected for the toxicokinetic portion of the UFA. Regarding toxicodynamics, in vivo (Carpenter
et al., 1956) and in vitro (Udden, 2002; Udden and Patton, 1994; Ghanayem and Sullivan, 1993)
studies indicate that humans may be significantly less sensitive than rats to the hematological
effects of EGBE. Because epidemiologic studies are often limited in their ability to explore
outcomes related to workplace or environmental exposures, it is typically impossible to rule out
the relevance of an effect seen in a particular rodent tissue unless there is detailed mechanistic
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information on why humans would not be affected (NRC, 2008; IARC, 2006). Therefore, the in
vivo human response to EGBE cannot be accurately determined without some degree of
speculation. For this reason, a value of 1 was selected for the toxicodynamic portion of the UFA
A factor to account for extrapolation from subchronic to chronic exposure (UFS) was not
needed because the RfD was derived from a chronic inhalation study.
A factor to account for the extrapolation from a LOAEL to a NOAEL (UFL) was not
applied because the current approach is to address this extrapolation as one of the considerations
in selecting a benchmark response (BMR) for BMD modeling. In this case, EPA concluded a
10% increase in hemosiderin staining, indicating a precursor to an adverse effect, is appropriate
for use in deriving the RfD under the assumption that it represents a minimal biologically
significant change.
A factor of 1 was selected to account for deficiencies in the database (UFD). While no
chronic oral studies or adequate human data are available for EGBE, PBPK models allow for
deriving a BMDL from the chronic inhalation study using measured internal dose metrics and
then extrapolating it back to an equivalent human oral dose. The database for inhalation exposure
includes chronic and subchronic studies in two species (rats and mice), and several reproductive
and developmental studies, including a two-generation reproductive toxicity study.
A total UF of 10 (10 for UFH, 1 for UFA, and 1 for UFD) was used in the derivation of the
RfD. The RfD for EGBE based on hemosiderin deposition in the liver was calculated as follows:
RfD = BMDLhec - UF
= 1.4 mg/kg-day -M0
= 0.1 mg/kg-day
Thus, the RfD is 0.1 mg/kg-day.
5.2.4. RfD Comparison Information
For comparison purposes, Figure 5-2 presents the POD, applied UFs, and derived
reference values, including the RfD, for the effect endpoints discussed. BMC modeling was
done using U.S. EPABMDS version 1.4.1 (U.S. EPA, 2000b), and results are provided in
Appendix C. This comparison is intended to provide information on alternative endpoints
associated with EGBE exposure. The selected RfD value is circled; BMD analyses are provided
in Appendix C.
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100
E
o
'•4—>
CB
CD
O
C
o
O
10
0.1
0.01
~ PODhed I I UFh
(mg/kg-d) 1 1
UF,
NTP 1993; 91 d
oral study;
hematological
changes;
LOAEL(HED)
NTP 2000; 2 yr
Inhalation study;
hemosiderin
staining;
BMDLio(HED)
RfD
(mg/kg-d)
NTP 1993; 91 d
oral study;
RBC changes;
BMDL05(HED)
Figure 5-2. PODs for selected endpoints with corresponding applied UFs
and derived RfD.
Figure 5-2 shows PODs and comparison reference values (including the RfD) that could
be derived from the various endpoints to allow a comparison with the chosen critical effect and
the resultant RfD for the critical effect. Hematological effects and effects related to hemolysis
(i.e., hemosiderin deposition) are the most sensitive endpoints for identification of a NOAEL or a
BMDL in the subchronic and chronic studies available; these endpoints have been considered as
the critical endpoint for derivation of the RfD.
5.2.5. Previous Oral Assessment
The previous IRIS assessment for EGBE was entered into the database on December 30,
1999 and contains an oral RfD of 0.5 mg/kg-day. The RfD was based on the BMD05(hed) of
5.1 mg/kg-day for changes in MCV in female F344 rats following a 91-day drinking water
exposure (NTP, 1993). A total UF of 10 was used to account for human variability. This
assessment was conducted prior to the adoption of the current BMD technical guidance
document (U.S. EPA, 2000b).
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1 5.3. UNCERTAINTIES IN THE DERIVATION OF THE INHALATION REFERENCE
2 CONCENTRATION (RfC) AND ORAL REFERENCE DOSE (RfD)
3 The following is a more extensive discussion of uncertainties associated with the RfC and
4 RfD for EGBE beyond the quantitative discussion in Sections 5.1.2, 5.1.3, 5.2.2, and 5.2.3. A
5 summary of these uncertainties, along with uncertainties specific to the Section 5.4 cancer
6 analysis, is presented in Table 5-8.
7
Table 5-8. Summary of uncertainty in the EGBE noncancer and cancer risk
assessments
Consideration
Potential impact
Decision
Justification
Choice of low-
dose
extrapolation
method
A linear low-dose
extrapolation would likely
drive up the risk
estimation when
combined with site-
specific exposure data.
Nonlinear approach; RfC
and RfD are considered
protective of the key
events leading to cancer.
Proposed key events in the two MOAs
(forestomach irritation and hemolysis
leading to increased liver hemosiderin) are
not likely to occur in humans at the RfC or
RfD.
Choice of
endpoint
Use of forestomach
endpoint could increase
RfC by up to 18-fold (see
Section 5.3).
RfC is based on the most
sensitive endpoint,
increased liver
hemosiderin.
Chosen endpoint is considered most
relevant due to forestomach toxicokinetic
and exposure issues.
Choice of dose
metric
Alternatives could
increase or decrease
RfC/RfD (e.g., use of Cmax
BAA increases RfC by
two to threefold).
AUC for BAA in arterial
blood.
Evidence for a causal association between
the proposed key steps of BAA and
hemolysis leading to increased liver
hemosiderin and increased liver tumors.
AUC chosen because hemosiderin
increased with cumulative exposure to
EGBE/BAA.
POD derivation
method
RfC/RfD threefold lower
than for NOAEL.
BMD method used.
Advantages include capacity to account for
sample size and to provide confidence
bounds on dose.
Choice of model
for BMCL
derivation
Alternative models could
increase RfC up to
threefold (see
Section 5.3).
Multistage (1st degree)
model chosen.
The best-fitting model was chosen based
on U.S. EPA (2000b) BMD technical
guidance.
Choice of animal
to human
extrapolation
method
Alternatives could
increase or decrease
RfC/RfD (e.g., default
would increase RfC by
twofold) (see Section 5.3).
A PBPK model was used
to extrapolate animal to
human concentrations.
Use of a PBPK model reduces uncertainty
associated with the animal to human
extrapolation. AUC blood levels of BAA
are an appropriate dose metric, and a peer-
reviewed and verified PBPK model exists
that estimates this metric.
Statistical
uncertainty at
POD
POD would be -40%
higher if BMD were used.
BMDL used per U.S.
EPA BMD guidance
(U.S. EPA, 2000b).
Limited size of bioassay results in
sampling variability; lower bound is 95%
CI on administered exposure.
Choice of
bioassay
Alternatives could
increase or decrease
RfC/RfD.
NTP (2000) study.
Alternative bioassays were inadequate.
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Table 5-8. Summary of uncertainty in the EGBE noncancer and cancer risk
assessments
Consideration
Potential impact
Decision
Justification
Choice of species/
gender
RfC would be increased if
based on another
species/gender.
RfC is based on the most
sensitive endpoint,
increased liver
hemosiderin, in the most
sensitive species and
gender, male rats.
Male mice are the only species/gender to
experience an increase in liver tumors
from EGBE exposure. If one relies on the
MO A, however, liver hemosiderin increase
in male rats is the appropriate key event
and species to utilize for the RfC. Female
mice exhibit forestomach papilloma but at
higher doses.
Human relevance
of rodent tumor
data
If MO As for tumors from
EGBE were deemed not
relevant, cancer descriptor
would be "suggestive of
human carcinogenic
potential."
MO As for liver and
forestomach tumors are
used.
Although EGBE has some evidence for
liver tumors in male mice and forestomach
tumors in female mice, the MOAs describe
rationales as to why humans are unlikely to
experience appreciable risk at
environmental concentrations.
Human
population
variability
Low-dose toxicity would
increase to an unknown
extent.
10-fold UF applied to
derive the RID and RfC
values.
10-fold UF is applied principally because
of limited data on human variability or
potential susceptible subpopulations.
5.3.1. Choice of Endpoint
Comparison RfC values were calculated (see Section 5.1.4) and are intended to provide
information on alternative health effects associated with EGBE exposure. The comparison RfCs
ranged from 1 to 23 mg/m3, depending on whether irritation (forestomach), hematologic effects,
or hemosiderin deposition data were used to derive the POD, with the latter endpoint
representing the lower end of the RfC range.
5.3.2. Choice of Dose Metric
The AUC for BAA in arterial blood was selected as the appropriate measure of dose due
to evidence for a causal association between BAA and hemolysis, between hemolysis and the
accumulation of hemosiderin in the liver, and between hemosiderin accumulation in the liver and
increased incidence of liver hemangiosarcoma. AUC is considered to be a more appropriate
response measure because hemosiderin pigmentation increases in incidence and severity with
increased duration (subchronic to chronic) and is believed to be the result of the cumulative
exposure to EGBE/BAA as opposed to a peak exposure. The corresponding aldehyde of BAA,
BAL, was also considered as a choice for an internal dose measure. As discussed in Section 4.6,
BAL is the EGBE metabolite considered to have the greatest potential to interact directly with
DNA. However, high ADH activity in the liver and forestomach is expected to result in very
short residence time and in very low tissue concentrations of BAL; this scenario has been
demonstrated in simulations using the Corley et al. (2005b) PBPK model. Also, the Corley et al.
(2005b) PBPK model along with the gavage study of Deisinger and Boatman (2004) suggest that
the conditions of in vitro assays showing BAL to be clastogenic (e.g., no metabolic activation;
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high cytotoxic concentrations of BAL) are considered to be of little relevance to the expected
target organ (liver) environment (e.g., high metabolic activity; low concentrations of BAL). Use
of an alternate measure of internal dose, for instance, a parent compound or Cmax, would be more
difficult to justify based on available empirical information. However, for comparative purposes,
a BMD analysis was done on the critical endpoint to determine the impact that choosing Cmax of
BAA in blood rather than AUC of BAA in blood would have had on the BMCLio(hec) derivation.
If Cmax had been used as the dose measure, the BMCLi0(hec) value would have been 109 mg/m3,
approximately 6.8-fold higher than the 16 mg/m3 BMCLio(hec) value derived using AUC as the
dose measure.
5.3.3. Use of BMC Approach
Utilization of the BMC approach has advantages over other approaches to dose-response
analysis, such as the NOAEL/LOAEL approach. These advantages include the capacity of the
BMC approach to accommodate study sample size and reflect this in providing confidence
bounds to the lower limit on dose. As shown in Figures 5-1 and 5-2, use of the BMC approach
on the incidence of chronic hemosiderin deposition resulted in RfC and RfD values lower than
RfC and RfD values that would have been derived via the NOAEL/LOAEL approach.
5.3.4. Choice of Model for BMCL Derivations
BMCLio estimates used in the derivation of the RfC, which formed the basis for both the
noncancer and cancer quantitative assessments, spanned a twofold range (60-124 |imol-hour/L)
for female rats and threefold range (40-126 (j.mol-hour/L) for male rats. All of the models fit
reasonably well (p-values above 0.1). Hence, this range of results can be considered a reflection,
in part, of model uncertainty.
5.3.5. Choice of Animal to Human Extrapolation Method
APBPK model (Corley et al., 1997) was used to extrapolate animal to human
concentration. An AUC blood level of BAA associated with a 10% increase in male mice with
hemosiderin pigmentation of 69.6 [j,mol-hour/L was estimated using the mouse PBPK model; the
human PBPK model was used to convert back to a human equivalent exposure concentration, or
a BMCLio(hec), of 12 mg/m3. If no PBPK models were available, the BMCLio(hec) would have
been derived by dividing the BMCLio for external exposure concentration of 75 mg/m3 by the
threefold pharmacokinetic portion of the animal to human default adjustment factor (U.S. EPA,
1994b), resulting in a BMCLi0(hec) of 25 mg/m3. This default value would have been twofold
higher than the value derived using the PBPK model.
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5.3.6. Route-to-Route Extrapolation
To estimate an oral dose POD for chronic hemosiderin deposition, a route-to-route
extrapolation was performed on the inhalation exposure POD used to derive the RfC using a
PBPK model and assuming, as discussed above, that the metric most closely associated with the
effect seen is the AUC measure of BAA in blood. One way of characterizing the uncertainty
associated with this approach is to compare dose levels (BMCL values) using this metric for
hemolytic effects—assumed to be associated with chronic hemosiderin deposition—derived from
(1) an existing oral subchronic NTP (1993) study; or (2) from a model estimating this metric
from an existing inhalation subchronic NTP (2000, 1998) study. This analysis was performed
(data not shown) and the values derived using the second procedure were slightly but
consistently lower than those derived using the former, suggesting that estimates using the
procedure employed for route-to-route extrapolation would tend to overestimate the toxicity
value and result in lower RfD estimates.
5.3.7. Statistical Uncertainty at the POD
Parameter uncertainty can be assessed through CIs. Each description of parameter
uncertainty assumes that the underlying model and associated assumptions are valid. For the
linear multistage model applied to the male mouse hemosiderin data, there is a reasonably small
degree of uncertainty at the 10% excess incidence level (the POD for derivation of the RfC),
with the 95%, single-sided lower confidence limit (BMCL) being about 30% below the
maximum likelihood estimate of the BMC.
5.3.8. Choice of Bioassay
The NTP (2000) inhalation study was used for development of the RfC and RfD. This
was a well-designed study, conducted in both genders in two species with an adequate number of
animals per dose group and with examination of appropriate toxicological endpoints in both
genders of rats and mice. Alternative comparable bioassays were unavailable.
5.3.9. Choice of Species/Gender
The RfC was based on the incidence of liver hemosiderin pigmentation in male rats, the
species and gender most sensitive to this effect (NTP, 2000). This event also occurs in female
rats and in mice, and is thought to be a precursor to the observed increase in liver tumors in male
mice. The cancer assessment was based on the more sensitive hemosiderin response in the male
rat, not the male mouse, because there is no evidence to suggest that the proposed MOA for
tumor formation would not be relevant to rats, and because the lack of an observed tumor
response in rats may have been due to the fact that rats were exposed to lower doses for a shorter
portion of their average lifespan (see further discussion in Sections 4.6 and 4.7.2).
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If the RfC had been based on increased incidence of liver hemosiderin in another
species/gender, such as male mice, a higher RfC value would have been derived. Similarly, the
RfC would also have been higher had it been based on forestomach irritation, an effect less
sensitive than hemosiderin deposition but considered to be a precursor event leading to the
increase in incidence of forestomach squamous cell papillomas and one high-dose carcinoma
observed by NTP (2000) in female mice.
5.3.10. Human Relevance of Noncancer Responses Observed in Mice
The effects of hemosiderin deposition and forestomach irritation may both have
qualitative relevance to humans. However, for reasons discussed in Section 4.5, the exposure
concentrations that would be necessary to cause these effects in humans, if attainable at all, are
likely to be much higher than the RfC/RfD and well above concentrations necessary to cause
these effects in mice.
5.3.11. Human Population Variability
The extent of interindividual variability associated with EGBE metabolism is not well
characterized in humans. As is discussed in Section 4.7, the hemolytic effect of EGBE is
presumed to be caused by the interaction of its primary metabolite, BAA, with the RBC
membrane. Potentially susceptible subpopulations or life stages would include individuals with
enhanced metabolism or decreased excretion of BAA. In addition, individuals whose RBC
membranes are more susceptible to the lysis caused by BAA could be more sensitive to EGBE.
However, RBCs from normal, aged, sickle-cell anemia, and hereditary spherocytosis patients
were no more sensitive to the hemolytic effects of BAA than RBCs from healthy volunteers
(Udden, 1994). As is discussed further in Section 4.7, some potentially susceptible
subpopulations or life stages have not been tested, and, when combined with the lack of
understanding about EGBE's mechanism of hemolytic action, this represents a considerable
source of uncertainty and forms the principal basis for the 10-fold UF applied to derive the RfD
and RfC values.
5.4. CANCER ASSESSMENT
In accordance with the Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a),
the method used to characterize and quantify cancer risk from a chemical depends on what is
known about the MOA of carcinogenicity and the shape of the cancer dose-response curve for
that chemical. An assumption of linearity is appropriate when evidence supports an MOA of
gene mutation due to DNA reactivity or an MOA that is anticipated to be linear. The linear
approach is used as a default option if the MOA of carcinogenicity is not known. A nonlinear
approach "can be used for cases with sufficient data to ascertain the MOA and to conclude that it
is not linear at low doses..(U.S. EPA, 2005a). Alternatively, the MOA may theoretically have
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a threshold; that is, the carcinogenicity may be a secondary effect of toxicity that is itself a
threshold phenomenon.
In the case of EGBE, the MO A of carcinogenicity for hepatic hemangiosarcoma,
hepatocellular adenoma and carcinoma, and forestomach tumor formation in animals is
reasonably well understood. An RfC and RfD approach has been used for EGBE because "When
adequate data on MOA provide sufficient evidence to support a nonlinear MOA for the general
population and/or any subpopulations of concern, a different approach—a RfD/RfC that assumes
nonlinearity—is used" (U.S. EPA, 2005a). It is recognized, however, that while this approach
fits this case, other nonlinear approaches may be appropriate in other settings. Available data
indicate that EGBE is not likely to be mutagenic and that it is not expected to produce rodent
tumors via a mutagenic MOA. Rather, there is evidence that carcinogenic responses observed in
animals are associated with erythrocyte hemolysis (leading to oxidative damage, increased
hepatocyte and endothelial cell proliferation, and initiation for the liver tumors) and with the
formation of BAA, an acidic metabolite, that leads to regenerative hyperplasia in response to
irritation for the forestomach tumors. Because cytolethality occurs only at exposure levels above
some critical dose, a nonlinear (threshold) approach is considered to be the most appropriate
method for characterizing the cancer risk from EGBE.
As discussed in Sections 4.1 and 4.6, there are currently no human studies addressing the
potential carcinogenicity of EGBE. A 2-year inhalation bioassay in mice and rats (NTP, 2000)
reported some evidence of carcinogenic activity in male B6C3Fi mice based on increased
incidences of hemangiosarcoma of the liver and an increase in the incidence of hepatocellular
carcinoma, as well as some evidence of carcinogenic activity in female B6C3Fi mice based on
increased incidences of forestomach squamous cell papilloma or carcinoma (mainly papilloma).
The study also reported no evidence of carcinogenic activity in male F344/N rats and equivocal
evidence of carcinogenic activity in female F344/N rats based on increased combined incidences
of benign and malignant pheochromocytoma (mainly benign) of the adrenal medulla. The
hypothesized MOA for the induction of hepatic hemangiosarcomas and hepatocellular
carcinomas reported in male mice exposed to EGBE may comprise a sequence of events that are
dose-dependent. First, EGBE is metabolized to the carboxylic acid, which then hemolyzes
RBCs. This event leads to the release of excess iron that in turn could result in iron-induced
formation of ROS and subsequent oxidative damage to target tissues within the liver. Induction
of cell proliferation and neoplasm formation follows. This sequence of events is considered
necessary for the formation of the observed neoplasms. Strategies intended to control or omit
any of these key events, including the initial hemolytic event, would interrupt the process and
prevent formation of neoplasms. For these reasons, formation of neoplasms in humans would
likely be prevented by establishing levels of EGBE exposure in the dose range that does not
result in the initial hemolytic events and subsequent ROS-mediated cellular damage. Thus, for
the assessment of human cancer risk associated with the formation of hemangiosarcomas and
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hepatocellular adenomas and carcinomas in animals, the RfD and RfC derived in Sections 5.1
and 5.2 should be considered protective, as the carcinogenicity may be a secondary effect of
toxicity that is itself a threshold phenomenon.
Available data suggest that the MOA for the induction of forestomach tumors reported in
female mice exposed to EGBE is dependent on the initial formation of the irritating metabolite
BAA. This acidic metabolite produces chronic cytotoxicity, which results in compensatory
epithelial cell regeneration. Chronic cell proliferation in preneoplastic cells is in turn associated
with the formation of forestomach papillomas and carcinomas. In addition, the available data
indicate that BAA may be preferentially formed in the forestomach due to the presence of levels
of ADH that are higher than those found in the human stomach or esophagus. BMD modeling of
the dose response for epithelial hyperplasia in female mice forestomachs was performed as a
comparison to reference values derived for precursor effects in the liver as discussed above and
is presented in Appendix C. The analysis in Appendix C shows that, had hyperplastic effects in
female mice been used as a POD, the resultant RfD and RfC values would have been much
higher than the RfD and RfC values derived in Sections 5.1 and 5.2 using hemosiderin
deposition as the critical effect for EGBE exposure. Thus, the analysis indicates that the RfD
and RfC based on liver effects would also be protective of forestomach toxicity. The exposure
concentrations that would be necessary to cause these effects in humans, if attainable at all, are
likely to be much higher than the RfC and RfD.
Female rats reported a marginally significant trend for pheochromocytomas, and the high
dose frequencies reported in the rats were only slightly different from the upper range of
historical controls (see Section 4.6). In addition, the histopathologic data indicated that there
was difficulty distinguishing pheochromocytomas from nonneoplastic adrenal medullary
hyperplasia. Thus, these lesions are interpreted with caution as tumors. Given the marginal dose
response, lack of tumor evidence in any other organ system of the rats, and reported difficulties
in distinguishing pheochromocytomas from nonneoplastic adrenal medullary hyperplasia, this
tumor was not given significant weight in the qualitative or quantitative assessment of EGBE
cancer potential.
5.4.1. Quantification for Oral and Inhalation Cancer Risk
Following the U.S. EPA (2005b) Guidelines for Carcinogen Risk Assessment, a nonlinear
approach to dose-response assessment is taken for agents, such as EGBE, for which the most
plausible mode of action at low doses is consistent with nonlinearity. The RfC of 1.6 mg/m3 and
RfD of 0.1 mg/kg-day derived in Sections 5.1 and 5.2 represent the outcome of nonlinear
assessments based on hemolytic effects (i.e., hemosiderin depostition)associated with both oral
and inhalation exposures to EGBE. Doses (or concentrations) of EGBE below the RfC (or RfD)
would not be expected to produce hemolytic effects (i.e., hemosiderin deposition) and therefore
are not expected to produce any increase in cancer risk.
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5.4.2. Uncertainties in Cancer Risk Assessment
The cancer assessment of EGBE is based largely on the premise that key events in the
MO As proposed for mice (forestomach irritation and hemolysis leading to increased hemosiderin
deposition) are not likely to occur in humans at concentrations at or below the RfC and RfD
values. Uncertainties in the RfC and RfD derivations are addressed in part in Sections 5.1.2,
5.1.3, 5.2.2, 5.2.3, and 5.3. This section will discuss additional uncertainties relative to the
human relevance of cancer responses observed in mice and the extrapolation method employed
for the estimation of low-dose cancer risk. All of these uncertainties are summarized in
Table 5-8.
5.4.2.1. Choice of Low-Dose Extrapolation Method
The MOA is a consideration in estimating risks. A linear low-dose extrapolation
approach was not considered to be optimal for the estimation of human carcinogenic risk
associated with EGBE exposure because of the physicochemical properties of EGBE,
toxicokinetic limitations, specific MO As articulated earlier, and limitations in data to
parameterize appropriate models. It should be noted that the demonstration that a chemical is not
mutagenic is insufficient, alone, to postulate a nonlinear dose response. Key events in the MO As
proposed for mice (forestomach irritation and hemolysis leading to increased hemosiderin
deposition) are not likely to occur in humans at concentrations at or below the RfC and RfD
values. This assumes that the proposed MO As in mice are reasonably correct. If, for instance,
hemosiderin accumulation in the mouse liver is not the result of increased hemolysis but of
EGBE interaction with other cell types, a more direct, linear MOA for the observed increase in
male mouse hemangiosarcomas might be hypothesized. In order to illustrate the predicted
cancer risk under such a scenario, the cancer risks associated with the tumor types that were
increased following EGBE exposure were calculated using the default approach of low-dose
linear extrapolation outlined in the cancer guidelines (U.S. EPA, 2005a). The results of this
linear analysis are presented in Table 5-9.
Table 5-9. Illustrative potency estimates for tumors in mice, using a linear
analysis approach
BMDLiohec
(mg/m3)a
Slope factor
0.1/BMDLiohec
(risk/mg-m3)
Hepatocellular carcinoma (males)
208
4.8 x 10-4
Hemangiosarcomas (males)
575
1.7 x 10-4
Papilloma or carcinoma of the forestomach (females)
544
1.8 x 10-4
"BMDLiohec values were calculated using AUC as the dose metric.
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For illustrative purposes, an estimate of the increased cancer risk at the RfC (1.6 mg/m3)
was calculated, using the inhalation unit risk for hepatocellular tumors in male mice, that was
derived using a default linear low-dose extrapolation approach. The estimated increased cancer
risk at the RfC in this comparison exercise would be 1.6 mg/m3 x 4.8 x 10"4 = 7.6 x 10"4 This
value is only for illustrative purposes and indicates the differences in the two approaches. It
should not be misconstrued as an estimate of the cancer risk at the RfC since the default linear
approach is not recommended.
5.4.2.2. Human Relevance of Cancer Responses Observed in Mice
The hypothesized MO As for EGBE-induced liver tumors observed in male mice and
EGBE-induced forestomach tumors observed in female mice may both have qualitative
relevance to humans. However, for reasons discussed in Section 4.6, the exposure
concentrations that would be necessary to cause these effects in humans, if attainable at all, are
likely to be much higher than the RfC and RfD.
5.5. POTENTIAL IMPACT OF SELECT UNCERTAINTIES ON THE RFC
In this assessment, the RfC forms the basis for the RfD. The range of possible results
associated with some of the areas of uncertainty in the RfC derivation can be estimated (see
Sections 5.3 and 5.4). Figure 5-3 graphically illustrates the change in the RfC that would result
had particular choices, other than those presented in this assessment, been made (see summary in
Table 5-8). These specific areas were presented in this illustration because there are data
available that could be used to quantify their contribution to the uncertainty in the noncancer and
cancer assessments. The "Cancer approach" illustrates the dose that would represent a 10"6
increased cancer risk if a default linear low-dose extrapolation approach were used. This dose is
expressed as X-fold change in the RfC.
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Areas of Uncertainty - X-fold Change from Alternative Choices
OJ
U)
c
TO
_c
0
1
¦?
X
20
15
10
5
0
-5
-10
-15
-50
-55
18
6.8
~ . ~
18
2
~
1.4
~
56
1 Nonlinear —t Linear
2 Hemosiderin —» Forestomach
3 /\UC —
4 Low BMDL —» NOAEL
5 Low — High BMDL
6 PBPK —~ Default
7 BMDL — BMD
8 Rat —~ Mouse
Cancer
Approach1
Endpoint7
Dose
Metric3
POD
Method"
BMD
Model5
HEC
Derivation6
Benchmark
POD7
Species
Figure 5-3. Potential impact of select uncertainties on the RfC for EGBE.
The area of uncertainty that has the potential to have the greatest impact on the EGBE
assessment is that which is associated with the approach used to extrapolate human cancer risk
from the available rodent studies; that is, if human cancer risk had been estimated from a linear
extrapolation of responses observed in rodent studies. Figure 5-3 provides a graphic of the areas
of uncertainty, described in Table 5-8, for which there is quantitative information and impact on
the RfC can be estimated. The "Cancer Approach" value in Figure 5-3 represents the difference
in the dose that was selected for the RfC and the dose that would represent a 10"5 increased
cancer risk if a default linear low-dose extrapolation approach were to be used. In this instance,
the comparison indicated that the dose associated with a 10"6 cancer risk, estimated via the linear
dose-response method using the inhalation unit risk for hepatocellular tumors in male mice as
indicated in Table 5-9, is 56-fold lower than the RfC. Note that the default linear approach is not
recommended.
These discussions and illustrative tools should not be viewed as a comprehensive analysis
of all possible uncertainty considerations; rather, they exemplify what are believed to be
important areas of uncertainties for which data are available. This characterization is presented
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1 in an effort to make apparent the limitations of the assessment and to aid and guide the risk
2 assessor in the ensuing steps of the risk assessment process.
3
4
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6. MAJOR CONCLUSIONS IN Till CHARACTERIZATION OF HAZARD AND DOSE
RESPONSE
6.1. HUMAN HAZARD POTENTIAL
EGBE is a clear, miscible solvent used in formulating cleaning products and protective
coatings. It is metabolized primarily to the carboxylic acid, BAA, the proximate toxicant, in
both humans and animals. BAA and its conjugates are readily excreted in the urine.
Hemosiderin deposition in the liver of male rats, as a consequence of hemolysis, is
identified as the critical endpoint of concern in toxicological studies on EGBE. Toxicologically,
this effect increases with both exposure concentration and duration, is not compensated for (as
with hemolysis), and is considered to be a pathological finding. Mechanistically, this effect is at
the interface between noncancer and cancer effects from exposure to EGBE.
Observations regarding the potential relevance of EGBE toxicity to humans include the
insensitivity of human RBCs to the hemolytic effects of EGBE and its metabolite, BAA. While
it is established that humans can experience hemosiderin deposition in the liver as a consequence
of excessive hemolysis (Iancu et al., 1976), the relative insensitivity of human blood to the
effects of EGBE have been demonstrated in numerous in vitro studies through the use of either
whole blood or washed erythrocytes (e.g., Udden, 2002; Ghanayem and Sullivan, 1993).
Humans appear significantly less sensitive to the hemolytic toxicity of EGBE than are typical
laboratory species, such as mice, rats, or rabbits, with reports from analyses of isolated RBCs
demonstrating that human RBCs are 40- to 150-fold less sensitive than rat RBCs (Udden, 2002).
These observations are inclusive of human RBCs from individuals with hereditary spherocytosis
and sickle cell anemia, conditions characterized by RBC sensitivity to hemolysis. Available in
vivo information with human exposure supports this species disparity in sensitivity to the
hemolytic effects of EGBE. Male rats in one study (NTP, 1993) experienced mild liver effects at
a drinking-water dose lower than that which caused observable hemolytic effects. Available
human toxicity data show that the primary effects after acute oral ingestion of large doses of
EGBE (most often combined with other solvents) are reversible metabolic acidosis from the
production of BAA and some hematological changes. Occupational exposure to low levels of
EGBE were not reported to cause adverse changes in hepatic, renal, or hematologic parameters
(Haufroid et al., 1997).
Due to the known reproductive toxicity (i.e., toxicity to male testes and sperm) of two
other glycol ethers, EGME and EGEE, the reproductive toxicity of EGBE has been studied in a
variety of well-conducted oral studies (NTP, 1993; Exon et al., 1991; Heindel et al., 1990; Foster
et al., 1987; Grant et al., 1985; Nagano et al., 1984, 1979) and inhalation studies (NTP, 2000;
Nachreiner, 1994; Doe, 1984; Dodd et al., 1983) using rats, mice, and rabbits. In addition,
several developmental studies have addressed EGBE's toxicity from conception to sexual
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maturity, including toxicity to the embryo and fetus, following oral exposures (Sleet et al., 1989;
Wier et al., 1987), inhalation exposures (Nelson et al., 1984; Tyl et al., 1984), and dermal
exposures (Hardin et al., 1984) to rats, mice, and rabbits. EGBE did not cause biologically
significant effects in any reproductive organ, including testes, in any study. In a two-generation
reproductive toxicity study, fertility was reduced in mice only at very high maternally toxic doses
(>1,000 mg/kg). Maternal toxicity related to the hematologic effects of EGBE and relatively
minor developmental effects have been reported in developmental studies. No teratogenic
effects were noted in any of the studies. It can be concluded from these studies that EGBE is not
significantly toxic to the reproductive organs (male or female) of either parents or to the
developing fetuses of laboratory animals.
No reliable human epidemiological studies are available that address the potential
carcinogenicity of EGBE. The NTP (2000) performed a 2-year inhalation bioassay with rats and
mice and found no evidence of carcinogenic activity in male F344/N rats and equivocal evidence
of carcinogenic activity in female F344/N rats based on increased combined incidences of benign
and malignant pheochromocytoma (mainly benign) of the adrenal medulla. The researchers
reported some evidence of carcinogenic activity in male B6C3Fi mice based on an increased
incidence of hemangiosarcoma of the liver and an increase in the incidence of hepatocellular
carcinoma that may have been exposure related. They also reported some evidence of
carcinogenic activity in female B6C3Fi mice based on an increased incidence of forestomach
squamous cell papilloma or carcinoma (mainly papilloma). Based on its physical-chemical
properties, toxicokinetic and dynamic factors, and MOA information, under existing U.S. EPA
guidelines (U.S. EPA, 2005a), EGBE is judged not likely to be carcinogenic to humans at
expected environmental concentrations (see Section 4.6). The MO As presented for the animal
tumors indicate that both high doses and sustained periods of exposure are necessary for the
carcinogenic response. The available human exposure/response information indicates that these
conditions are unlikely to occur because the primary response of humans to high oral doses of
EGBE, as shown in the case studies in Section 4.1 is metabolic acidosis, which, if not treated,
can lead to shock and eventually death. Further, based on simulations from PBPK modeling, the
maximum blood concentrations of BAA that could be produced in humans following exposure to
a saturated atmosphere of EGBE would be below those needed to produce hemolysis (Corley et
al., 2005a). Evidence from the only human inhalation exposure study available showed that,
while nasal and ocular irritation were reported in research subjects exposed to up to 195 ppm, no
changes in erythrocyte fragility were observed (Carpenter et al., 1956). For a more complete
discussion of the carcinogenic potential of EGBE, see Section 4.6.
6.2. DOSE RESPONSE
6.2.1. Noncancer—Inhalation
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Studies have not been reported in which humans were exposed subchronically or
chronically via inhalation to EGBE. After consideration of the available animal inhalation
studies with EGBE, the NTP (2000) study was selected as the principal study because it was
conducted in two species and provides data for different durations and for multiple dose groups
compared to other available studies.
For derivation of the RfC, the most sensitive endpoint for dose-response assessment is the
effect of inhalation exposure on hemosiderin staining in Kupffer cells, as reported in the study by
NTP (2000). The concentration of BAA in the blood was used as an internal dose metric for
EGBE exposure. The human PBPK model of Corley et al. (1997, 1994) was used to obtain
estimates of human inhalation exposure concentrations associated with the BMCs derived from
rat BAA AUC levels reported by Dill et al. (1998). The RfC is based on the human equivalent
BMCL io of 16 mg/m , which was back-calculated from rat data using the BMD and PBPK
approach.
The use of uncertainty factors used to derive the RfC are as follows: a factor of 10 was
selected to account for the uncertainty associated with the variability of the human response
(UFh) to the effects of EGBE, a factor of 1 was selected to account for the uncertainty associated
with interspecies variability resulting from toxicodynamic and toxicokinetic differences between
animals and humans (UFA), and a factor of 1 was selected to account for deficiencies in the
database. A total UF of 10 (10 for UFH, 1 for UFA, and 1 for UFD) was used in the derivation of
the RfC. Thus, the RfC is 16 mg/m3 -M0 = 1.6 mg/m3.
The overall confidence in the RfC assessment is medium to high. Higher confidence is
placed in the RfC values derived from internal dose measures (PBPK method and combined
PBPK/BMC method) because pharmacokinetic differences between rats and humans were
accounted for using PBPK models (Corley et al., 2005a, 1997; Lee et al., 1998) and actual
measurements of internal blood concentrations in test animals of interest (Dill et al., 1998).
Higher confidence is placed on the NTP (2000) study because it was a chronic study, it employed
both male and female rats and mice, it had a wide range of exposure levels, and animals were
observed twice daily. Medium-to-high confidence is placed on the database because data are
available for a variety of animal species, including humans. While the database lacks long-term
human studies, the available short-term human controlled studies and case reports, as well as
laboratory animal and in vitro studies provide evidence to suggest that, with respect to the
hemolytic effects of EGBE, long-term human exposures would be no more adverse than long-
term rat exposures. Additionally, the selection of medium-high confidence is supported because
the potential for effects in humans from repeat, long-term exposures has not been investigated.
6.2.2 Noncancer—Oral
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Studies have not been reported in which humans have been exposed subchronically or
chronically to EGBE by the oral route of exposure, and thus would be suitable for derivation of
an oral RfD. No chronic oral laboratory animal studies are currently available for EGBE.
Hematological effects leading to RBC lysis and organ accumulation of iron in the form of
hemosiderin accumulation appear to be the most sensitive of the effects caused by EGBE in
laboratory animals. Because the hemolytic effects of EGBE appear to be consistent between oral
and inhalation routes of exposure, an RfD has been derived via the application of a PBPK model
to perform a route-to-route extrapolation from the incidence of hemosiderin pigmentation
observed in the NTP (2000) chronic inhalation study of EGBE. As with the animal to human
extrapolation used in the development of the RfC, the dose metric used for interspecies (rat to
human) and route-to-route (inhalation to oral) extrapolation was the AUC of BAA at 12 months
in arterial blood. This dose metric was used for dose-response modeling of chronic inhalation
data (Section 5.1.2) to derive the POD of 133 [j.mol-hour/L, expressed as a BMDLio. The
BMDLio was then back-calculated using the human PBPK model (Corley et al., 1997, 1994) to
obtain an equivalent human oral drinking water dose (BMDLHed) of 1.4 mg/kg-day.
The use of uncertainty factors used to derive the RfD are as follows: a factor of 10 was
selected to account for the uncertainty associated with the variability of the human response
(UFh) to the effects of EGBE, a factor of 1 was selected to account for the uncertainty associated
with interspecies variability resulting from toxicodynamic and toxicokinetic differences between
animals and humans (UFA), and a factor of 1 was selected to account for deficiencies in the
database. A total UF of 10 (10 for UFH, 1 for UFA, and 1 for UFD) was used in the derivation of
the RfD. Thus, the RfD is 1.4 mg/kg-day -M0 = 0.1 mg/kg-day.
The overall confidence in the RfD assessment is medium to high. The RfD value has
been calculated for EGBE using a route-to-route extrapolation from the inhalation PBPK/BMC
method used to derive the RfC. A higher confidence is placed in the RfD values derived from
this combined method, since pharmacokinetic differences between rats and humans were
accounted for using a validated PBPK model (Corley et al., 1997, 1994). High confidence is
placed on the NTP (2000) study because it was a chronic study, it employed both male and
female rats and mice, it had a wide range of exposure levels, and animals were observed twice
daily. Medium-to-high confidence is placed on the database, because data are available for a
variety of animal species, including humans. While the database lacks long-term human studies,
the available short-term human controlled studies, case reports, laboratory animal, and in vitro
studies provide ample evidence to suggest that, with respect to the hemolytic effects of EGBE,
long-term human exposures would be no more adverse than long-term rat exposures.
Additionally, the selection of medium-high confidence is supported because the potential for
effects in humans from repeat, long-term exposures has not been investigated.
6.2.3. Cancer—Oral and Inhalation
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Information regarding the reported liver and forestomach tumors observed in laboratory
animals exposed to EGBE indicates that the MO As underlying these lesions are nonmutagenic
and include intermediate processes that have nonlinear dose-response characteristics. Control or
omission of these intermediate events would likely be sufficient to prevent the occurrence of
such tumors in humans, including potentially sensitive subpopulations such as children.
Application of nonlinear quantitative assessment methods indicate that the noncancer
RfD (0.1 mg/kg-day) and RfC (1.6 mg/m3) values developed for EGBE are considered protective
of these key events and would serve to prevent the occurrence of carcinogenic effects in
humans.9 The exposure concentrations that would be necessary to cause these effects in humans,
if attainable at all, are likely to be much higher than the RfC and RfD.
The estimation of uncertainty in this analysis is based on the alternative approaches for
estimating the dose response that are discussed and shown in Section 5.1.4 and Appendix C.
These alternatives include using the NOAEL approach combined with the measure of the internal
dose estimated with the PBPK model and the BMC approach combined with the same PBPK
model. Sections 5.1.3 and 5.2.3 summarize the uncertainty associated with each of these
approaches for the RfC and RfD, respectively. Other uncertainties associated with the noncancer
and cancer assessments presented in this ToxicologicalReview are summarized in Table 5-8 and
in Sections 5.4.1 and 5.5.
9These analyses are consistent with the nonlinear assessment approach described in the 2005 cancer guidelines (U.S.
EPA, 2005a).
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70(66): 17765-18717. Available online at http://www.epa.gov/cancerguidelines (accessed September 21, 2009).
U.S. EPA (Environmental Protection Agency). (2005b) Supplemental guidance for assessing susceptibility from
early-life exposure to carcinogens. Risk Assessment Forum, Washington, DC; EPA/630/R-03/003F. Available
online at http://www.epa.gov/cancerguidelines (accessed September 21, 2009).
U.S. EPA (Environmental Protection Agency). (2005c) An evaluation of the human carcinogenic potential of
ethylene glycol butyl ether, Final Report. National Center for Environmental Assessment, Office of Research and
Development, February 2005. EPA 600/R-04/123.
U.S. EPA (Environmental Protection Agency). (2006a) Science policy council handbook: peer review. 3rd edition.
Office of Science Policy, Office of Research and Development, Washington, DC; EPA/100/B-06/002.
U.S. EPA (Environmental Protection Agency). (2006b) A framework for assessing health risk of environmental
exposures to children. National Center for Environmental Assessment, Washington, DC; EPA/600/R-05/093F.
Available online at http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=158363 (accessed September 21, 2009).
U.S. EPA (Environmental Protection Agency). (2006c) Approaches for the application of physiologically-based
pharmacokinetic (PBPK) models and supporting data in risk assessment. National Center for Environmental
Assessment, Washington, DC; EPA/600/R-05/043F. Available online at
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Udden, MM. (1994) Hemolysis and decreased deformability of erythrocytes exposed to butoxyacetic acid, a
metabolite of 2-butoxyethanol: II. Resistance in red blood cells from humans with potential susceptibility. J Appl
Toxicol 14:97-102.
Udden, MM. (1995) Effects of butoxyacetic acid on human red cells. Occup Hyg 2:283-292.
Udden, MM. (2000) Rat erythrocyte morphological changes after gavage dosing with 2-butoxyethanol: a
comparison with the in vitro effects of butoxyacetic acid on rat and human erythrocytes. J Appl Toxicol 20:381-
387.
Udden, MM. (2002) In vitro sub-hemolytic effects of butoxyethanol acid on human and rat erythrocytes. Toxicol
Sci 69:258-264.
Udden, MM; Patton, CS. (1994) Hemolysis and decreased deformability of erythrocytes exposed to butoxyacetic
acid, a metabolite of 2-butoxyethanol. I. Sensitivity in rats and resistance in normal humans. J Appl Toxicol 14:91—
96.
Udden, MM; Patton, CS. (2005) Butoxyacetic acid-induced hemolysis of rat red blood cells: effect of external
osmolality and cations. Toxicol Lett 156(1):81—93.
Valberg, LS; Simon, JB; Manley, PN; et al. (1975) Distribution of storage iron as body stores expand in patients
with hemochromatosis. J Lab Clin Med 86(3):479-489.
Wang, YJ; Ho, YS; Lo, MJ; et al. (1995) Oxidative modification of DNA bases in rat liver and lung during chemical
carcinogenesis and ageing. Chem-Biol Interact 94:135-145.
Werner, HW; Nawrocki, CZ; Mitchell, JL; et al. (1943a) Effects of repeated exposure of rats to vapors of monoalkyl
ethers of ethylene glycol. J Ind Hyg Toxicol 25:374-379.
Werner, HW; Mitchell, JL; Miller, JW; et al. (1943b) Effects of repeated exposure of dogs to monoalkyl ethylene
glycol ether vapors. J Ind Hyg Toxicol 25:409-414.
Wier, PJ; Lewis, SC; Traul, KA. (1987) A comparison of developmental toxicity evident at term to postnatal growth
and survival using ethylene glycol monoethyl ether, ethylene glycol monobutyl ether, and ethanol. Teratog
Carcinog Mutag 7:55-64.
Wintrobe, MM. (1981a) Variations of leukocytes in disease. In: Wintrobe, MM; ed. Clinical hematology.
Philadelphia, PA: Lea & Febiger; pp. 1284-1323.
Wintrobe, MM. (1981b) The normocytic, normochromic anemia. In: Wintrobe, MM; ed. Clinical hematology.
Philadelphia, PA: Lea & Febiger; pp. 677-697.
Yamaguchi, R; Hirano, T; Asami, S; et al. (1996) Increased 8-hydroxyguanine levels in DNA and its repair activity
in rat kidney after administration of a renal carcinogen, ferric nitrilotriacetate. Carcinogenesis 17:2419-2422.
Yoshizawa, K; Kissling, GE; Johnson, JA; et al. (2005) Chemical-induced atrial thrombosis inNTP rodent studies.
Toxicol Pathol 33(5):517-532.
Zeiger, E; Anderson, B; Haworth, S; et al (1992) Salmonella in mutagenicity tests. V. Results form the testing of
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Ziouzenkova, O; Asatryan, L; Sevanian, A. (1999) Oxidative stress resulting from hemolysis and formation of
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APPENDIX A. SUMMARY OF EXTERNAL PEER REVIEW AND PUBLIC
COMMENTS AND DISPOSITION
The Toxicological Review of Ethylene Glycol Monobutyl Ether has undergone formal
external peer review performed by scientists in accordance with U.S. EPA guidance on peer
review (U.S. EPA, 2006a). The external peer reviewers were tasked with providing written
answers to general questions on the overall assessment and on chemical-specific questions in
areas of scientific controversy or uncertainty. A summary of significant comments made by the
external reviewers and U.S. EPA's responses to these comments arranged by charge question
follow. In many cases the comments of the individual reviewers have been synthesized and
paraphrased for development of Appendix A. Public comments that were received are
summarized and addressed following the peer-reviewers comments and disposition.
On April 10, 2008, U.S. EPA introduced revisions to the IRIS process for developing
chemical assessments. As part of the revised process, the disposition of peer reviewer and public
comments, as found in this Appendix, and the revised IRIS Toxicological Review was provided
to the external peer review panel members for their comment on May 15, 2009. Any additional
peer review panel comments received as part of this second review and U.S. EPA's responses are
included at the end of this Appendix.
General Charge Questions:
1. Is the Toxicological Review logical and clear? Has EPA accurately, clearly and
objectively represented and synthesized the scientific evidence for noncancer and cancer
hazard?
Comments:
In general, reviewers found the Toxicological Review to be logical and clear. In addition,
they found that the review accurately, clearly, and objectively represented and synthesized the
scientific evidence for noncancer and cancer hazard.
More specifically, several reviewers provided suggestions to improve the clarity of the
document and one reviewer questioned the quality of analytical procedures used in the critical
study.
Two reviewers stated that the PBPK modeling of EGBE and BAA was not well
illustrated or adequately described in the document and one reviewer felt that Figure 4-2 was not
very informative.
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One reviewer commented that the low renal clearance of BAA relative to GFR needs to
be discussed in more depth. The same reviewer posed several questions: is there data on plasma
albumin binding of BAA and does it vary between rodents and humans? If the binding is less
than -95%, it is unlikely to be the explanation of the restricted renal elimination. The alternate
explanation of active reuptake by a carrier mechanism is amenable to study: is there relevant
information in the literature? Resolution of this question is deemed crucial in understanding the
factors that determine both the Cmax and AUC of BAA in the various species/sexes of rodents as
well as the determination of HECs.
One reviewer commented that the toxicokinetic discussion (Section 3.2) is, perhaps of
necessity, somewhat diffuse and suggested a table incorporating some of the relevant kinetic
information.
Response:
The suggestions for document improvement were incorporated into the Toxicological
Review wherever possible and appropriate. Additional text has been added in Section 4.5 to
clarify the effects of EGBE on RBCs and the possible pathophysiological mechanisms involved.
Comments concerning the quality of the critical study were considered, but do not negate the
positive findings nor preclude the use of study dose-response data.
A brief summary of the existing PBPK models and rationale for use have been included in
Section 3.2 as well as in Appendix B. We have eliminated Figure 4-2 as PBPK simulations for
peak (Cmax) concentrations of EGBE, BAL, and BAA in the liver of male mice following 6-hour
inhalation exposures to 250 ppm (the highest concentration used in the NTP bioassay) were
adequately described in the text.
The PBPK modeling paper by Corley et al. (1994) describes the basis for inclusion of
plasma protein binding and renal clearance of BAA for male F344 rats and humans. The
parameters for protein binding were assumed to be similar to phenolsulfophthalein and renal
clearance optimized from in vivo pharmacokinetic studies that indicated that the elimination of
BAA in urine by male F344 rats was saturable, leading to nonlinear increases in blood BAA Cmax
and AUC values at high dose levels. Parameters developed from rats were successfully scaled to
humans as demonstrated by Franks et al. (2006), Jones et al. (2003), and Corley et al. (1997,
1994), for 'low-dose' inhalation and dermal exposures, and Gualtieri et al. (2003, 1995) for 'high
dose' oral exposures (suicide attempt). In their 1998 publication, Lee et al. (1998) updated the
model parameters using the in vivo toxicokinetic data for male and female, young and old, rats
and mice. As identified by the reviewer, these processes are experimentally accessible. Thus,
Corley et al. (2005a) conducted a series of studies to determine the species, gender, and age-
dependent tissue:blood partitition coefficients for BAA, plasma protein binding of BAA,
metabolism of EGBE and BAA, and renal active transport of BAA. These parameters were used
to update the PBPK model of Lee et al. (1998). The toxicokinetic summary (Section 3.2) has
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1 been revised to include additional information from Corley et al. (2005a) who determined plasma
2 protein binding, partition coefficients, and renal elimination of BAA in young vs. aged male and
3 female rats and mice.
4 A table of relevant toxicokinetic information has been included in Section 3.2 to
5 summarize the information presented in the text.
6
7 2. Please identify any additional studies that should be considered in the assessment of the
8 noncancer and cancer health effects of EGBE.
9
10 Comments:
11 Most reviewers did not identify any additional studies that should be considered in the
12 assessment of the health effects. However, a few reviewers did provide suggestions for
13 supporting references and expanded discussions that are listed below.
14
15 a.
16 b.
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18 c.
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20 d.
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22 e.
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24 Response:
25 a. The subject of thrombosis and infarction is briefly discussed in Section 4.4.1. Additional
26 text has been added in Section 4.5 of the Toxicological Review including the suggested
27 reference as well as others. The added references are as follows:
28
Bevers, EM; Comfurius, P; van Rijn, JL; et al. (1982) Generation of prothrombin-
converting activity and the exposure of phosphatidylserine at the outer surface of
platelets. Eur JBiochem 122(2):429-436.
Connor, J; Bucana, C; Fidler, IJ; et al. (1989) Differentiation-dependent expression of
phosphatidylserine in mammalian plasma membranes: quantitative assessment of outer-
leaflet lipid by prothrombinase complex formation. Proc Natl Acad Sci USA.
86(9):3184-3188.
Ezov, N; Levin-Harrus, T; Mittelman, M; et al. (2002) A chemically induced rat model of
hemolysis with disseminated thrombosis. Cardiovasc Toxicol 2:181-194.
Nyska, A; Maronpot, RR; Long, PH; et al. (1999b) Disseminated thrombosis and bone
infarction in female rats following inhalation exposure to 2-butoxyethanol. Toxicol
Pathol 27(3):287-294.
References on the subjects of thrombosis and infarction.
References on the subjects of hematology instrument methodology differences and iron
overload.
References on the subject of the role of Kupffer cells in hepatotoxicity and
carcinogenicity.
While no specific references were supplied, one reviewer suggested an increased
discussion on the MOA of EGBE on the RBC and its fate.
An increased discussion on the subject of olfactory hyaline membrane degeneration.
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Ramot, Y; Lewis, DA; Ortel, TL; et al. (2007) Age and dose sensitivities in the
2-butoxyethanol F344 rat model of hemolytic anemia and disseminated thrombosis. Exp
Toxicol Pathol 58(5):311-322.
Yoshizawa, K; Kissling, GE; Johnson, JA; et al. (2005) Chemical-induced atrial
thrombosis in NTP rodent studies. Toxicol Pathol 33(5):517—532.
b. All of the references cited in the Toxicological Review pertaining to MCV used the
impedence type analyzers (the more sensitive of the two types commented on).
Therefore, no changes were made to the document. The references suggested on iron
overload were reviewed and found to be inappropriate for incorporation into the
document based on the high levels of iron compared to what we see with EGBE as
presented in the review.
c. Additional text and the suggested reference have been added to the Toxicological Review
on the role of Kupffer cells in hepatotoxicity and carcinogenicity in Section 4.5. The
following reference was added:
Roberts, RA; Ganey, PE; Ju, C; et al. (2007) Role of the Kupffer cell in mediating hepatic
toxicity and carcinogenesis. Toxicol Sci 96(1):2—15.
d. Additional text has been added to the Toxicological Review discussing the MOA of
EGBE on the RBC and its fate in Section 4.5.
e. Additional text and references have been added to the Toxicological Review discussing
the effect of olfactory hyaline membrane degradation and the U.S. EPA's rationale for its
exclusion as a critical effect. The references added are as follows:
Buckley, LA; Morgan, KT; Swenberg, JA; et al. (1985) The toxicity of dimethylamine in
F344 rats and B6C3Fi mice following a 1-year inhalation exposure. Fundam Appl
Toxicol 5(2):341-352.
Nikula, KJ; Novak, RF; Chang, IY; et al. (1995) Induction of nasal carboxylesterase in
F344 rats following inhalation exposure to pyridine. Drug Metab Dispos 23(5):529-535.
St. Clair, MBG; Morgan, KT. (1992) Changes in the upper respiratory tract. In Mohr, U;
Dungworth, DL; Capen, CC; eds. Pathobiology of the Aging Rat, Vol. 1, ILSI Press;
Washington, DC: pp. 111-127.
3. Please discuss research that you think would be likely to increase confidence in the
database for future assessments of EGBE.
Comments:
All reviewers provided research suggestions that may increase confidence in the database for
future EGBE assessments. The following is a brief summary of these comments and
suggestions:
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• Measure iron levels in the liver;
• Do dose/response for oxidant-induced liver cell initiation and promotion;
• Define the age of erythrocyte susceptibility to BAA;
• Definition of the MOA of the toxic insult to the erythrocyte with definition of dose
metrics;
• Determine the loss of sidedness in the phospholipid composition of the outer leaflet of the
red cell membrane to the premature loss of red cells in the circulation and the possibility
that this may induce microvascular thrombosis and infarction;
• Determine the role of the spleen vs. the liver in the removal of damaged red cells and
effects on Hp and hemopexin levels and saturation during chronic exposure to EGBE;
• Determine the time and dose/response relationships for hemosiderin deposition in the
liver as a whole and in specific cell types;
• Determine the MOA and dose/response relationships for hemosiderin in
hemangiosarcoma development;
• Experiments designed to determine the fate of BAA damaged RBCs and the role of
spleen and liver macrophages;
• Measure the phosphatidylserine levels on RBCs following exposure to BAA;
• Experiments designed to better define the cellular effects of BAA exposure and their role
in hemosiderin deposition and ROS toxicity;
• Determination of the dose and temporal relationships between EGBE-induced
hemosiderin and its induction of hemolysis;
• Experiments designed to define the nature and dose response of effects seen in humans
following moderate exposures;
• Experiments designed to characterize the species differences in RBC membrane
physiology;
• Long term human inhalation and oral studies of EGBE;
• Experiments designed to address the mechanism(s) by which BAA induces RBC
hemolysis in various species;
• Experiments designed to investigate the role of ROS in modulation of endothelial and
hepatic cell gene expression;
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• In vivo studies designed to quantify the internal threshold doses that must be met to
progress from one key precursor event in the mechanistic sequence to another; and
• Experiments designed to determine the relative susceptibility of endothelial cells and
hepatocytes to oxidative damage.
Response:
A number of research suggestions were received, and U.S. EPA agrees further research
may enhance future risk assessments of EGBE. No additional changes to the Toxicological
Review are indicated at this time.
4. Please comment on the identification and characterization of sources of uncertainty in
Sections 5 and 6 of the assessment document. Please comment on whether the key sources
of uncertainty have been adequately discussed. Have the choices and assumptions made in
the discussion of uncertainty been transparently and objectively described? Has the impact
of the uncertainty on the assessment been transparently and objectively described?
Comments:
All of the reviewers found Sections 5 and 6 to be well written and complete. A number
of reviewers had comments on the selection of the factors that account for uncertainty (UFs).
Additional suggestions included the following: (1) an alternative endpoint of olfactory
epithelial degeneration; (2) Sections 5.1.3 and 5.2.3 be expanded for clarity; and (3) further
analyses (Hazard Quotients and Hazard Indices) be performed for Section 6.
Response:
Specific comments on the values of the UFs are addressed in the response to charge
question A5.
The potential use of olfactory epithelial degeneration as an endpoint is discussed in
Section 5. Additional references have been added that support the position that olfactory
epithelial degeneration is not an adverse effect but an adaptive response.
Substantial improvements were made to Section 5 that address the suggestion for
improved clarity. Supporting text for alternative derivations has been moved to Appendix C to
improve the readability of the document.
Computing Hazard Quotients and Hazard Indices would require an exposure analysis,
and while an portent component of risk assessment, is beyond the scope of this human health risk
assessment.
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Chemical-Specific Charge Questions:
(A) Inhalation reference concentration (RfC) for EGBE
Al. The 2-year inhalation study by the National Toxicology Program (NTP, 2000) was
selected as the basis for the chronic inhalation RfC. Please comment on whether the
selection of this study as the principal study has been scientifically justified. Has this study
been transparently and objectively described in the document? Please identify and provide
the rationale for any other studies that should be selected as the principal study.
Comments:
All reviewers found the selection of the NTP (2000) study to be scientifically justified
and transparently and objectively described in the document. No additional studies were
identified for use as the principal study. A number of editorial comments were provided to
improve the document.
Response:
Based on the favorable comments from the reviewers, no substantial changes were made
to the Toxicological Review with respect to the selection of the principal study for the RfC.
Suggested editorial comments have been reviewed and revisions were made to the Toxicological
Review where appropriate.
A2. The incidence of hemosiderin staining in the liver of male rats was selected as the
critical effect because it is considered by EPA to be a precursor to an adverse effect. Please
comment on whether the selection of this critical effect has been scientifically justified. Are
the criteria and rationale for this selection transparently and objectively described in the
document? Please provide a detailed discussion. Please identify and provide the rationale
for any other endpoints that should be considered in the selection of the critical effect.
Comments:
Five of the seven reviewers agreed with the conclusion that hemosiderin staining in the
liver of male rats was the most appropriate choice for the critical effect. However, two of those
five reviewers communicated reservations regarding this conclusion and discussed the
deficiencies of its use. It was suggested that both hemosiderin and Hct, as parameters for POD
purposes, be discussed and illustrated in the review.
Two reviewers thought that alternate endpoints were more appropriate, such as RBC
counts, Hct levels, MCV, or hyaline degeneration.
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Response:
RfC and RfD derivations from all endpoints were considered and are now illustrated in
Appendix C as well as summarized in Sections 5.1.4 and 5.2.4 of the assessment. The dose-
response assessments used in the derivation of the RfC and RfD are reviewed in Sections 5.1.2
and 5.1.4, respectively. Based on the most recent information on the MOA for EGBE,
hemosiderin deposition was selected as the critical effect. While the U.S. EPA recognizes that
other hematological endpoints are biologically relevant, hemosiderin provides the best overall
choice based on pathological considerations, dose-response relationships, and model fit. The
support and rationale for the choice of the critical effect can be found in Sections 5.1, 5.2, 6.1,
and 6.2.
A3. Benchmark dose (BMD) modeling was applied to incidence data for hemosiderin
staining in male rat liver to derive the point of departure (POD) for the RfC. Please
provide comments with regard to whether BMD modeling is the best approach for
determining the POD. Has the BMD modeling been appropriately conducted and
objectively and transparently described? Has the benchmark response (BMR) selected for
use in deriving the POD (i.e., 10% extra risk of hemosiderin staining in the liver) been
scientifically justified, and transparently and objectively described? Please identify and
provide the rationale for any alternative approaches for the determination of the POD and
discuss whether such approaches are preferred to EPA's approach.
Comments:
All reviewers agreed that BMD modeling was the best approach for determining the
POD. However, alternatives were suggested for the choice of species and gender. One reviewer
commented that male mice should have been used since they have shown correlation between the
critical effect and the adverse effect (tumor). Two other reviewers commented that female rats
should have been used since they have been shown to be more sensitive to the hemolytic effects
and were used in the previous IRIS assessment. Lastly, one reviewer commented on the choice
of 10% response level. This reviewer felt that a 5% response level is scientifically supported
because it approximates the NOAEL.
Response:
Based on the majority of comments agreeing with the choice of gender and endpoint to
model, no substantial changes to the Toxicological Review are indicated. Rationale for the
selection of the species and gender are discussed in Section 5.3.9. Briefly, the male rat was
chosen based on the NTP (2000) report showing male rats were the most sensitive species and
gender with respect to the critical effect (hemosiderin deposition). In addition, the modeling of
the data for male rats provided a much better fit than the modeling for the female rat data. As for
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the selection of the response rate, utilizing U.S. EPA guidance on BMD modeling of qualitative
data 10% was chosen because it was within the range of experimental responses and appropriate
for the power of the study. Additional justification has been added regarding the fact that this
tumor response may not have been observed in the rats because the rats were exposed to lower
doses and for a shorter duration of their average lifespan than the mice.
A4. PBPK modeling was used to extrapolate the POD from rats to humans. Please
comment on whether the PBPK modeling for interspecies extrapolation is scientifically
justified, and transparently and objectively described in the document. Does the model
properly represent the toxicokinetics of the species under consideration? Was the model
applied properly? Are the model assumptions, parameter values, and selection of dose
metrics clearly presented and scientifically supported?
Comments:
All reviewers that provided comments agreed that the PBPK modeling for EGBE was
scientifically justified and transparently and objectively described in the text. Two reviewers
commented that a comparison between the 1999 assessment and this one would be helpful.
Response:
Based on the comments, no change in the Toxicological Review is indicated.
Comparisons to the previous assessment are made throughout the document in the appropriate
sections.
A5. Please comment on the selection of all of the uncertainty factors applied to the POD
for the derivation of the chronic RfC. For instance, are they scientifically justified, and
transparently and objectively described in the document? An UF of 10 for extrapolation
from animals to humans (UFA) is generally applied when data are not available to inform
potential pharmacokinetic (PK-UF) and pharmacodynamic (PD-UF) differences. In this
assessment, an UFA of 1 was applied.
• A PBPK model was used to inform pharmacokinetic differences and a PK-UF of 1 was
selected. Please comment on whether this selection is scientifically justified. Is the
rationale transparently and objectively described? Please comment on whether there are
sufficient scientific data and support for the use of this PBPK model to estimate
interspecies toxicokinetic differences and to replace the default interspecies factor for
1/2
toxicokinetic differences (i.e., 10 ).
• Evidence from human and animal in vitro and in vivo studies were used to inform
pharmacodynamic differences and a PD-UF of 1 was selected. Please comment on
whether this selection is scientifically justified. Is the rationale transparently and
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objectively described? Please comment on whether a higher value for the PD-UF should
be used (e.g., to account for the limited information available on the potential for effects
in human cell types other than red blood cells) or alternatively, should a lower (i.e.,
fractional) PD-UF be used (e.g., to account for the 40-150-fold difference in the
concentrations that cause prehemolytic effects in human red blood cells (RBCs),
including RBCs from potential susceptible populations such as the elderly, and patients
suffering from anemia and RBC disorders that weaken the cellular membrane such as
hereditary spherocytosis).
Please identify and provide the rationale for any alternative approaches for the selection of
the uncertainty factors.
Comments:
Several reviewers commented that the selection of factors to account for uncertainty
(UFs) was not supported by the available science.
a. Two reviewers conluded that the UFA should be <1.
b. Two reviewers commented that the UFH should be <10.
c. One reviewer conluded that the UFD was inconsistent with the confidence level.
d. Most reviewers agreed that the selection of the UF values should be more clearly
justified.
Response:
The current UFa value in the ToxicologicalReview is 1. The UFa accounts for a
reduction from 3 for the toxicokinetic differences between animals and humans through
the use of a PBPK model for extrapolation of doses. The toxicodynamic portion,
likewise was reduced from 3 to 1. The use of a 1 for the toxicodynamic portion of the
UFa represents the lowest reduction described by the current guidance (U.S. EPA,
1994b). The toxicological effect in question being the deposition of hemosiderin is
identified as a key event in the MOA for the development of liver hepatomas in mice. To
implement a further reduction in the UFA (i.e., fractional) would logically indicate that a
preponderance of data are available to describe this key event (toxicodynamics) in both
animals and humans to the extent of describing why humans are less sensitive to the
hemolytic effects leading to hemosiderin deposition. This is not the case, although in
vitro data (Udden and Patton, 2005; Udden, 2002, 2000; Ghanayem, 1989) do suggest
humans are less sensitive than rodents to the hemolytic effects of EGBE. Likewise, the
few human studies (Haufroid et al., 1997; Carpenter et al., 1956) indicate the same
finding. However, these studies (Udden and Patton, 2005; Udden, 2002, 2000; Haufroid
et al., 1997; Ghanayem, 1989; Carpenter et al., 1956) do not characterize hemosiderin
deposition, the key event for the MOA In addition, little is known of the long-term or
repeated exposure responses in humans to EGBE.
Lastly, part of the definition of a UF is that it is represented by a number that is generally
within an order of magnitude (i.e., 10). Implementation of UFs is described by dividing
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15 A6. Please comment specifically on the database uncertainty factor of 1 applied in the RfC
16 derivation. Are the criteria and rationale for the selection of the database uncertainty
17 factor transparently and objectively described in the document? Please comment on the
18 body of information regarding the hemato and hepatic toxicity of EGBE and the use of the
19 toxicokinetic data in the determination of the database uncertainty factor. Please comment
20 on whether the selection of the database uncertainty factor for the RfC has been
21 scientifically justified. Has this selection been transparently and objectively described in
22 the document?
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24 Comments:
25 The majority of reviewers commented that the UFD value was appropriate. One reviewer
26 commented that the value was inconsistent with the U.S. EPA's confidence level of medium to
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29 Response:
30 Based on the reviewers comments, no change in the document is indicated. The value
31 assigned for the database UF is based on the completeness of the database in terms of
32 toxicological studies assessing the range of likely potential effects including reproductive and
33 developmental effects as well as information from more than one species.
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(B) Oral reference dose (RfD) for EGBE
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36 Bl. A conclusion was reached that the available oral toxicity data are inadequate to
37 support derivation of a chronic oral RfD value. Is the rationale for not developing an RfD
38 from the available database of oral studies transparently and objectively described? If
39 other oral studies are identified that would be suitable for the derivation of the RfD, please
40 identify and provide the rationale for their use.
the POD by a factor of 1, 3, or 10 for each of the defined UFs to calculate a human health
toxicity values. The use of a fractional UF would represent a deviation from the current
guidance and is beyond the scope of this Toxicological Review.
There is very limited information on the sensitivity of various human subpopulations to
the hemolytic effects of EGBE. In addition, the long-term and repeated exposure effects
in potentially sensitive populations remain to be determined. Therefore, the default UFH
of 10 is appropriate. This subject is discussed in Section 5.2.3.
Response to this comment can be found in question A6 below.
Additional text has been added to Sections 5.1.3 and 5.2.3 for improved justification of
the chosen UFs; the sections were reformatted for improved clarity.
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Comments:
All reviewers commented that the 91-day drinking water study should be considered in
the RfD development and that the justification for not using it to at least derive comparative RfD
estimates was inadequate.
Response:
Currently, there are no chronic oral studies of EGBE and very limited sub-chronic
studies. The larger database of inhalation studies, including chronic timepoints, supports the use
of a route-to-route extrapolation using the chronic inhalation study and PBPK modeling to
produce an RfD estimate. The 91-day drinking water study has been included in Chapter 5 for
comparison purposes. Alternative endpoints including the endpoint selected for the previous
assessment are presented graphically and discussed in a new section (5.2.4).
B2. A route-to-route extrapolation was performed to derive the chronic RfD, using the
chronic inhalation study and PBPK modeling. The Human Equivalent Concentration
(HEC) was based on a continuous oral exposure to EGBE in drinking water that would
yield the same AUC for the metabolite BAA (in the arterial blood over three months) as
that estimated for the rat following an external inhalation exposure to EGBE at the level of
the proposed POD (i.e., the BMCLio). Please comment on whether the PBPK model is
adequate for use to conduct a route-to-route extrapolation for EGBE to derive an RfD in
the absence of adequate oral animal or human dose-response data to derive the RfD
directly. Was the extrapolation correctly performed and objectively and transparently
documented?
Comments:
All comments found the extrapolation was correctly performed and objectively and
transparently presented in the document.
Response:
Based on the reviewer's comments, no change in the ToxicologicalReview is indicated.
B3. Please comment specifically on the database uncertainty factor of 1 applied in the RfD
derivation. Are the criteria and rationale for the selection of the database uncertainty
factor transparently and objectively described in the document? Measured internal doses
in rats and a human PBPK model were used to perform a route-to-route extrapolation to
derive the RfD. Please comment on the use of the PBPK model and the inhalation database
in the determination of the database uncertainty factor for the RfD. Please comment on
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whether the selection of the database uncertainty factor for the RfD has been scientifically
justified. Has this selection been transparently and objectively described in the document?
Comments:
General agreement among the reviewers found that the selection of the database UF for
the RfD was scientifically justified and transparently and objectively described in the document.
Response:
Based on the reviewer's comments, no change in the ToxicologicalReview is indicated.
(C) Carcinogenicity of EGBE
CI. Under the EPA's 2005 Guidelines for Carcinogen Risk Assessment
(www.epa.gov/iris/background.htm), the Agency concluded that EGBE is not likely to be
carcinogenic to humans at expected exposure concentrations. Please comment on the
scientific justification for the cancer weight of evidence characterization and describe the
basis for your view. Has the scientific justification for the weight of evidence descriptor
been sufficiently, transparently and objectively described?
Comments:
All reviewers agreed with the Agency's conclusion regarding the cancer descriptor for
EGBE. Two reviewers suggested the phrase "at expected exposure concentrations" should either
be deleted or better defined with exposure concentrations encountered. These reviewers
provided a possible revision as follows: "at the calculated RfC and RfD values presented in this
document."
All reviewers commented that the scientific justification for the cancer weight of
evidence descriptor had been sufficiently, transparently, and objectively described.
Response:
Because all reviewers agreed with the Agency conclusion that EGBE is not likely to be
carcinogenic to humans at expected exposure concentrations, qualitatively this conclusion was
not revised. As suggested by two of the reviewers on the External Peer Review Panel, the
specific text to the weight of evidence descriptor was revised to account for their suggested
language. Lastly, all reviewers agreed that the scientific justification for the cancer weight of
evidence descriptor had been sufficiently, transparently, and objectively described; thus,
revisions were not requested for the document.
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C2. EPA has proposed a mode of action (MOA) for male mouse liver cancer involving
metabolism, hemolysis of RBCs, hemosiderin deposition in the liver, oxidative damage and
proliferation leading to tumor induction as key events best supported by the data. Please
provide detailed comments on whether this analysis regarding the MOA for liver cancer is
scientifically sound, and transparently and objectively described in the Toxicological
Review. Considerations include the scientific support regarding the plausibility for the
hypothesized MOA and the characterization of uncertainty regarding this MOA.
Comments:
In general, the reviewers found the analysis regarding the MOA for liver cancer
scientifically sound, and transparently and objectively described in the review, although several
reviewers provided scientific considerations regarding the MOA, which should be considered for
the completion of the human health risk assessment. Most notably, remarks regarding
hemosiderin deposition include: (1) Why does hemosiderin deposition not occur in the spleen?;
(2) Tumor induction from initiated cells is speculation; (3) If male mice represent the most
sensitive species for tumor formation due to exposure to EGBE, why is a more marked dose
response for hemosiderin deposition observed in female mice?; and (4) Two reviewers
commented that hemosiderin deposition was a biomarker of exposure, not effect.
Response:
Responses to this charge question will be addressed in the order in which they were
presented in the comment overview (above): (1) One reviewer inquired as to why hemosiderin
deposition does not occur in the spleen. NTP (2000) reported hemosiderin deposition does occur
in the spleen for which these data are presented in Table 4-5; (2) Tumor induction from initiated
cells is not speculated within the Toxicological Review. With respect to the hypothesized MOA
for forestomach tumors, the MOA is the same as that proposed for hepatocellular tumors
observed in male mice (NTP, 2000). Within this MOA, the occurrence of initiated cells is not
specifically addressed but could occur from two of the key events including oxidative stress and
cytotoxicity. Although data do not currently exist for EGBE, the possibilities also include
epigenetic pathways that could lead to the formation of initiated cells. A key point regarding the
hypothesized MOA is that the key events represent obligatory (rate limiting) steps or biomarkers
of such events leading to the formation of tumors, but do not represent a complete description of
the biologic mechanisms that lead to tumor development; (3) Hemosiderin deposition observed
was indeed greater in female compared to male mice (NTP, 2000; see Table 4-5). NTP has
observed liver hemangiosarcomas in 105/4,183 (2.51%) male versus just 35/4,177 (0.84%)
female historical controls (Klaunig and Kamendulis, 2005; NTP, 2000). In addition, other
chemicals reported by NTP to cause both early onset hemosiderin buildup and liver tumors have
also exhibited this male specificity (U.S. EPA, 2005c). While the reason for the sex difference in
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liver tumor susceptibility between male and female mice is not clear, it has been shown that
estrogens can be protective through their antioxidant capacities and through their modulation of
the activities of other antioxidants (Nyska et al., 2004; Section 4.6.3). Given the significant
mortality of the male mice in the study, the doses received by the female mice may not have
represented a sufficient dose to induce tumors (i.e., maximum tolerated dose); and
(4) Hemosiderin deposition is a key event in the MOA for the formation of liver tumors in male
mice. As stated previously, a key event represents an obligatory (rate limiting) step or a
biomarker of such an event. This would not preclude the use of hemosiderin deposition as a
biomarker of such an event, as long as a dose-response and temporal relationship are observed to
indicate it as an obligatory step in the MOA.
C3. EPA has proposed a MOA for female mouse forestomach tumors involving
metabolism, irritation and regenerative proliferation leading to tumor induction as key
events best supported by the data. Please provide detailed comments on whether this
analysis regarding the MOA for forestomach tumors is scientifically sound, and
transparently and objectively described in the ToxicologicalReview. Considerations
include the scientific support regarding the plausibility for the hypothesized MOA and the
characterization of uncertainty regarding this MOA.
Comments:
All reviewers found the analysis regarding the MOA for forestomach tumors
scientifically sound, and transparently and objectively described in the review and generally
agreed with the overall conclusions. A few reviewers provided comments to refine the
supporting text and conclusions for the MOA.
One reviewer felt that the discussion on the MOA was too speculative and that steps 5
and 6 should be deleted.
Response:
The overall conclusions regarding the MOA for forestomach tumors were not modified
within the Toxicological Review because comments were not received that indicated such a
change. Editorial and clarification changes to refine the supporting text were made to the
document.
While step 5 of the proposed sequence of events leading to forestomach tumors seems
redundant, it is necessary to illustrate that the development of tumors is not a single exposure
scenario but requires a repeated cycle of injury and repair. Regarding step 6, EGBE-specific
evidence of ROS modulation in endothelial cells and hepatocytes is a biologically plausible step
within the hypothesized MOA, even though it is not definitively supported. It is important to
note that key events within the MOA may have different levels of supporting scientific evidence.
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C4. EPA has not proposed a MOA for the female rat pheochromocytomas of the adrenal
medulla. NTP rated the female rat pheochromocytomas as providing equivocal evidence of
carcinogenic activity and the pathology report expressed concern as to whether the
observed tumors met the criteria used to diagnose pheochromocytomas. For these reasons,
this tumor was not given significant weight in the qualitative or quantitative assessment of
EGBE cancer potential. Please provide detailed comments on whether this analysis
regarding the female rat pheochromocytomas is scientifically sound, and transparently and
objectively described in the Toxicological Review. Please comment on whether and the
extent to which the female rat pheochromocytomas are adequate to support alternative
analyses of qualitative and quantitative cancer risks to humans and discuss approaches to
consider if such analyses are warranted.
Comments:
All comments communicated that the analysis regarding female rat pheochromocytomas
was scientifically sound, and transparently and objectively described in the review. No
dissenting scientific opinions were submitted that would warrant alternative analysis. One
reviewer did suggest a possible extension of the NTP conclusion that pheochromocytomas
represented equivocal findings. This reviewer also suggested the possibility of including a linear
low-dose extrapolation for this tumor type as an academic exercise to explore the range of
possibilities but did not recommend its use for determining risk.
Response:
As stated in the charge question, U.S. EPA has not proposed a MOA for the female rat
pheochromocytomas of the adrenal medulla. NTP concluded that the data for the female rat
pheochromocytomas provided equivocal evidence of carcinogenic activity and the pathology
report expressed concern as to whether the observed tumors met the criteria used to diagnose
pheochromocytomas. For these reasons, this tumor was not given significant weight in the
qualitative or quantitative assessment of EGBE cancer potential.
Thus, linear low-dose extrapolation for pheochromocytomas was not included in the
Toxicological Review.
C5. Please comment on the choice of the nonlinear threshold approach for the quantitative
assessment of the carcinogenic potential of EGBE. Please comment on whether this
approach is scientifically sound, and transparently and objectively described. Please
comment on whether the example calculations using linear low-dose extrapolation for
cancer as discussed in Section 5.4.1 represent useful characterizations of the potential
quantitative uncertainty associated with exposure to EGBE. Please comment on whether
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the linear analysis should be presented as an alternative to the threshold approach
considering the Agency conclusion that EGBE is not likely to be carcinogenic to humans at
expected exposure concentrations.
Comments:
All reviewers supported the choice of the nonlinear threshold approach commenting that
this approach was scientifically sound, and transparently and objectively described in the review.
Most reviewers found the presentation of the linear low-dose extrapolation to be informative for
comparative purposes, but two reviewers felt it should not be presented as an alternative to the
threshold approach.
Response:
Based on the majority of comments, no change in the Toxicological Review is indicated.
Public Comments
Comment:
The use of hemosiderin staining as the critical effect for RfC and RfD development is not
biologically appropriate and is inconsistent with the available dose-response data.
Response:
The use of hemosiderin staining as the critical effect for RfC and RfD development has
been peer reviewed by a panel of experts and found to be appropriately used. While some
increase in hemosiderin accumulation can be expected with advancing age, excessive
accumulation beyond that contributed by age is considered pathological (Muller et al., 2006).
While the hemolytic effects appeared to be among the earliest effects from EGBE exposure, the
hemosiderin deposition endpoint was selected as the critical effect. This effect was found to
occur in both species and genders of animals tested, with rats being the more sensitive species;
the effect also occurred in the 14-week subchronic NTP inhalation study. The suggested MOA
of EGBE-induced liver effects is based on the observation that the hemolytic effects led to
compensatory erythropoiesis and significant increases in blood degradation products, including
an increased accumulation of hemosiderin in the liver Kupffer cells of EGBE-exposed animals.
The hemosiderin accumulation seen in the Kupffer cells was found to increase in severity with
increasing dose and exposure duration (Table 5-2), unlike the hemolytic endpoints, such as
decreased Hct, which did not progress from 3 to 12 months (Table 4-6). Thus, hemosiderin
deposition in Kupffer cells in the rat liver is believed to be a sequela to the hematologic effects.
Because of the progression of this effect with chronic exposure, hemosiderin is deemed to be the
most sensitive effect (Section 5.1.1). Additional support for the use of hemosiderin, as opposed
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to the various hematological parameters, is the location of the effect. Kupffer cell accumulation
of hemosiderin in the liver is consistent with the pathological findings of liver
hemangiosarcomas and hepatocellular carcinomas found in EGBE-exposed male mice.
Comment:
The use of male rat data as the POD is inappropriate because the available data
convincingly demonstrate that female rats are more sensitive to the hemolytic effects of EGBE.
Response:
U.S. EPA agrees that female rats are more sensitive to the hemolytic effects of EGBE.
However, hemolysis is one step in a complex process leading to the critical endpoint,
hemosiderin accumulation. Male rats have been shown to be more sensitive to hemosiderin
accumulation than females (NTP, 2000). The fit statistics and BMC information derived from
the dichotomous models available in the BMD software as applied to the male and female rat
hemosiderin staining data versus AUC BAA are shown in Table 5-5. All models were fit using
restrictions and option settings suggested in the U.S. EPA BMD technical guidance document
(U.S. EPA, 2000b). The best model fit to these data, as determined by visual inspection,
examination of low dose model fit (i.e., scaled residual for the dose group closest to the BMD),
and comparison of overall fit (i.e., AIC values) was obtained using a multistage model
(1st degree) for the male response data and a Log-Logistic model for the female response data.
The male rat BMCio was 196 [j,mol-hour/L and the BMCLio was determined to be 133 [jmol-
hour/L, using the 95% lower confidence limit of the dose-response curve expressed in terms of
the AUC for BAA in blood. The BMCio and BMCL io values for the female rat were determined
to be 425 and 244 [j,mol-hour/L, respectively. Assuming continuous exposure (24 hours/day),
the Corley et al. (1997) PBPK model was used to back-calculate HECs of 3.4 ppm (16 mg/m3)
from the male rat data and 4.9 ppm (24 mg/m3) from the female rat data (Section 5.1.2.1). If you
combine the males and females for the BMC analysis, the result is <10% of the original BMCLio
for males alone. This indicates that there is an inconsistency with at least one of the data sets in
the female study. Based on all the information, the use of male rats in the assessment is
warranted.
Comment:
The choice of dose metric, and the application of PBPK modeling in the derivation of
candidate RfCs based on hematological data should be re-examined. The selection of an internal
dose metric to be used in the dose-response assessment using the incidence of hemosiderin
staining is not adequately justified.
The Lee et al. (1998) model has been incorrectly applied in the derivation of the Cmax
BAA values reported for female rats in Table 5.
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Similar discrepancies (current draft vs. 1999 document) are evident in the BMD
modeling based on hematological endpoints.
Response:
The AUC for BAA in arterial blood was selected as the appropriate measure of dose due
to evidence for a causal association between BAA and hemolysis, between hemolysis and the
accumulation of hemosiderin in the liver, and between hemosiderin accumulation in the liver and
increased incidence of liver hemangiosarcoma. AUC is considered to be a more appropriate
response measure because hemosiderin pigmentation increases in incidence and severity with
increased duration (subchronic to chronic) and is believed to be the result of the cumulative
exposure to EGBE/BAA as opposed to a peak exposure. However, for comparative purposes, a
BMD analysis was done on the critical endpoint to determine the impact that choosing Cmax of
BAA in blood rather than AUC of BAA in blood would have had on the BMCLio(hec) derivation.
If Cmax had been used as the dose measure, the BMCLi0(hec) value would have been 39 mg/m3,
approximately 2.4-fold higher than the 16 mg/m3 BMCLi0(hec) value derived using AUC as the
dose measure (Section 5.3.2). The choice of hemosiderin as the critical effect warrants the use of
AUC as the dose metric.
The reviewer is correct; the arterial blood BAA Cmax estimates for female rats reported in
Table 5-4 were in error. Table 5-4 was revised using the most recent PBPK model of Corley et
al. (2005a), which replaced several of the assumptions utilized by Lee et al. (1998) with
measured values for protein binding, partition coefficients, metabolism, and renal clearance. For
comparison, the Lee et al. (1998) simulations along with the simulations using the Corley et al.
(2005a) model are as follows:
Exposure
Female rat
concentration
body weight
Cm** BAA in arterial blood ([JVf)
(ppm)
(g)
Lee et al. (1998)
Corley et al. (2005a)
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As the reviewer pointed out, these revised Cmax estimates are now consistent with the
peak blood concentrations of BAA observed following 3 months of exposure to the three
concentrations used in the NTP chronic inhalation bioassay (31.25, 61.5, and 125 ppm) reported
by Dill et al. (1998). Table 5-3 and Appendix A were also revised to include a description of the
Corley et al. (2005a) PBPK model.
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The BMD calculations (Table 5-5) were revised using the Corley et al. (2005a) estimates
of arterial blood BAA Cmax values for female rats as described above. The HEC calculations
were again based upon the Corley et al. (1997) model.
Comment:
The use of inhalation data to derive the RfD is inappropriate. The current IRIS Review
of EGBE (U.S. EPA, 1999) selects a 91-day drinking water study in rats as the principal study
for the derivation of the RfD. Because information available from chronic inhalation studies
with EGBE indicates that the primary effect, hematological changes, does not become more
severe with prolonged exposures, the 91-day drinking water study in rats remains the appropriate
POD for deriving the RfD. Use of this study would preclude the need for a route-to-route
extrapolation from chronic inhalation data, with the substantial errors that can be associated with
this type of analysis (Pauluhn, 2003).
In evaluating the results of the route-to-route extrapolation, a comparison to what is
observed in the available oral studies is critical to determine whether the modeling results are
realistic or whether, instead, the modeling results contribute additional uncertainty to the derived
RfD. The route-to-route extrapolation in the draft IRIS Review is based on the observation of
male rat liver hemosiderin staining from the NTP (2000) inhalation study. With that data set as
the POD, PBPK modeling is used to derive the proposed new RfD of 0.14 mg/kg-day. In
comparing this RfD to the results from the 13-week oral study in rats (NTP, 1993), a significant
increase in hemosiderin staining was not observed in male rats until doses of 452 mg/kg-day
were achieved or in female rats until a dose of 281 mg/kg-day was achieved (Table 4-1). In
addition, significant changes in hematological endpoints were observed at lower doses
(69 mg/kg-day). The doses associated with significant increases in hemosiderin staining are
orders of magnitude above the POD of 1.4 mg/kg-day, suggesting that the use of the route-to-
route endpoint is not appropriate. Additional analyses are needed to determine why a difference
in response by route of exposure would be observed.
While route-to-route extrapolation is a valuable tool for chemicals for which no adequate
study is available for a selected route of exposure, in the case of EGBE, an adequate oral study is
available and should be used to derive the RfD.
Response:
Currently, there are no chronic oral studies of EGBE and very limited subchronic studies.
The larger database of inhalation studies including chronic timepoints supports the use of a
route-to-route extrapolation using the chronic inhalation study and PBPK modeling to produce a
more accurate RfD estimate. The 91-day drinking water study has been included in Chapter 5
for comparison purposes. Alternative endpoints including the endpoint selected for the previous
assessment are presented graphically and discussed in a new section (5.2.4). Inhalation studies
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considered for derivation of the RfC were used to supplement the oral database using the route-
to-route extrapolation (Section 5.2.2.1). As with the interspecies extrapolation used in the
development of the RfC, the dose metric used for interspecies (rat to human) and route-to-route
(inhalation to oral) extrapolation was the AUC of BAA at 12 months in arterial blood. This dose
metric was used for dose-response modeling of chronic inhalation data (Section 5.1.2) to derive
the POD of 133 [j,mol-hour/L, expressed as a BMDL. The BMDL was then back-calculated
using the human PBPK model (Corley et al., 1997, 1994) to obtain an equivalent human oral
drinking water dose (BMDLhed) of 1.4 mg/kg-day (Section 5.2.2.2). While the proposed RfD
(0.14 mg/kg-day) is lower than the RfD from the previous assessment (0.5 mg/kg-day; 1999),
this difference can be attributed to the use of hemosiderin accumulation as apposed to
hematological parameters as the critical endpoint.
Comment:
The intrahuman and interspecies UFs applied in the derivation of the RfC and RfD are
greatly overprotective and should be reevaluated. The draft IRIS Review repeatedly (e.g., pages
50, 59, and 109) mentions that humans are much less sensitive to the toxic effects of EGBE than
are the rodents that provide the basis for derivation of the RfC and RfD. These differences in
sensitivity are not simply due to pharmacokinetic differences (which are addressed in the PBPK
modeling), but differences in inherent sensitivity (pharmacodynamics) as illustrated by the
marked differences in sensitivity of human and rodent blood cells in vitro to EGBE and BAA.
For example, the work of Udden (2002) demonstrates about a 150-fold greater sensitivity of rat
blood cells than human blood cells to the effects of BAA on RBC deformability, osmotic
fragility, and hemolysis. Even potentially hypersensitive subgroups of the human population
(the elderly, and patients with sickle cell disease or hereditary spherocytosis) show similar
resistance to these effects of BAA (Udden, 2002).
Given the consistent, substantial difference in sensitivity between human and rats and the
data that is the basis for the RfD and RfC, there is no scientific justification for using a
pharmacodynamic UF as large as 1. The weight of the evidence supports the use of a fractional
value of perhaps 0.01, or even less.
Furthermore, the default UF of 10 for human variability is not needed. This factor is
typically applied to account for variations in human sensitivity or to be protective of sensitive
subpopulations. In this case, however, typical hypersensitive subgroups, such as the elderly, also
show resistance to the hematological effects of BAA, as do individuals with disease conditions
(patients with sickle cell disease or hereditary spherocytosis) that might be expected to make
them more sensitive (Udden, 2002). While animal studies suggest that older animals are more
sensitive than neonates, and females are more sensitive than males, these have been shown to be
a reflection of differences in pharmacokinetics, not pharmacodynamics or sensitivity (Corley et
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al., 2005a). Based on these findings, the proposed 10-fold intraspecies UF is clearly excessive.
For further information, see the response for charge question A5.
Response:
The current evidence indicates that the human RBC response to EGBE exposure is
considerably less than that of rodents. However, this information comes from a relatively small
number of studies. The current UFA value in the Toxicological Review is 1 and considers the
limited number of studies available as well as the unknown effect of EGBE on the cellular events
leading to hemolysis in human populations. In addition, little is known of the long-term or
repeated exposure responses in humans to EGBE.
The UF for human variation has been assigned a value of 10. Some studies have shown
that in vitro exposure of RBCs from proposed sensitive populations, such as the elderly and those
with iron handling diseases, produces similar effects as in the general population. However, the
database is small and represents only acute or short-term exposures. The effects of longer or
repeated exposures to the general population as well as the proposed sensitive populations
remain to be determined. Therefore, the use of the default value for UFH is justified.
Comment:
There are several problems with BMD modeling. While benchmark modeling is the most
scientifically appropriate approach for determining the POD using the available noncancer data
for EGBE, there are several problems in the implementation of the procedure. The additional
documentation of the modeling results provided in Appendix B of the draft has several
deficiencies. The output provided in Appendix B for the hemosiderin modeling does not use the
AUC doses so is not an example of the output used to derive the RFC. Specifically, the
multistage model output labeled "BMD Method for RFC: Hemosiderin deposition in male rats
versus AUC BAA, 2-year inhalation study (NTP, 2000)" in Appendix B does not use the doses
indicated in Table 5-6 as the AUC doses. The BMD and BMDL in this output are not the values
reported in Table 5-7. The same is also true of the log-logistic output labeled "BMD Method for
RfC: Hemosiderin deposition in female rats versus AUC BAA, 2-year inhalation study, (NTP,
2000)" in Appendix B. The BMD and BMDLs reported in these outputs are not given anywhere
else in the document. In addition, in Table 5-7, the female rat multistage (1-stage) output has the
BMDL in the BMD column and an incorrect number in the BMDL column (source unknown).
Response:
The reviewer is correct that the outputs of the BMD modeling do not match the cited
source studies. Older BMD outputs were inadvertently placed in the document appendix while
the correct values incorporated into the text. The BMD calculations have been revised and the
corrected outputs placed in the appendix.
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Additional External Peer Review Panel Comments - Second Review in Response to
Revisions As Indicated Above
General Charge Questions
Comment:
One reviewer questioned the replacement of the term "adverse" with the terms
"significant" and "biologically significant" throughout the text.
Response:
The term "adverse" was replaced to provide a more meaningful description. Many of the
instances where "biologically significant" was used in the document has been replaced with the
words "statistically significant". Occasionally, the term "biologically significant" remains to
reflect the author's interpretation of the biological significance of statistically significant
endpoints.
Chemical-Specific Charge Questions
(A) Inhalation reference concentration (RfC) for EGBE
A2.
Comment:
One reviewer commented that hyaline membrane degeneration could be regarded as a
precursor lesion and, while not suggesting that this is a suitable critical endpoint in isolation, it
might be a useful endpoint for quantitative comparison with other endpoints.
Response:
The potential use of hyaline membrane degeneration as endpoints was considered and
discussed in chapter 5 of the Toxicological Review. Based on the available literature, U.S. EPA
maintains the position that olfactory hyaline membrane degeneration is not suitable as a critical
endpoint. Based on the available literature, U.S. EPA concludes that hyaline membrane
degeneration represents an adaptive response and not an adverse effect. Any quantitative
comparison between hyaline membrane degeneration and hemosiderin deposition would
therefore not be appropriate for comparative purposes. Additionally, it would be unreasonable
given that there is no PBPK model to derive an HEC based on hyaline membrane degeneration.
Comment:
One reviewer felt the discussion on hemolytic anemia in Section 4.5 still required
revisions and provided alternative text. Specifically, the discussion on osmotic fragility and it's
relationship to EGBE-induced hemolytic anemia required clarification.
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Response:
The alternative text suggested by the reviewer has been incorporated into the discussion
on hemolytic anemia in Section 4.5 and the Toxicological Review was reviewed for other
occurrences of the phrase "osmotic fragility". The alternative text clearly discusses the available
information and limitation of what is known about the mechanism of EGBE-induced hemolytic
anemia.
A3.
Comment:
One reviewer questioned the selection of male rats as the most sensitive gender when the
data demonstrate that females are more sensitive to the hematological effects of EGBE and
display an increased incidence of hemosiderin deposition.
Response:
We agree that female rats are more sensitive to the hemolytic effects of EGBE.
However, hemolysis is one step in a complex process leading to the critical endpoint,,
hemosiderin accumulation. While the data from the subchronic study show females as more
sensitive, it is the 2-year chronic study that is the focus of our analysis. In that study, male rats
were shown to be more sensitive to hemosiderin accumulation than females (NTP, 2000). In
addition, the cumulative blood concentration of the active metabolite BAA was chosen as the
internal dose metric, as opposed to the exposure concentration of EBGE. The fit statistics and
BMC information derived from the dichotomous models available in the BMD software as
applied to the male and female rat hemosiderin staining data versus AUC BAA are shown in
Table 5-5. All models were fit using restrictions and option settings suggested in the U.S. EPA
BMD technical guidance document (U.S. EPA, 2000).b). The best model fit to these data, as
determined by visual inspection, examination of low dose model fit (i.e., scaled residual for the
dose group closest to the BMD)), and comparison of overall fit (i.e., AIC values) was obtained
using a multistage model (1st degree) for the male response data and a Log-Logistic model for
the female response data. The male rat BMCio was 196 |j,mol-h^mol-hour/L and the BMCLio
was determined to be 133 |j,mol-h[j,mol-hour/L, using the 95% lower confidence limit of the
dose-response curve expressed in terms of the AUC for BAA in blood. The BMCio and BMCLio
values for the female rat were determined to be 425 and 244 |j,mol-h[j,mol-hour/L, respectively.
Assuming continuous exposure (24 hours/day), the Corley et al. (1997) PBPK model was used to
back-calculate concentrations for the HECs of 3.4 ppm (16 mg/m3) from the male rat data and
4.9 ppm (24 mg/m3) from the female rat data (Section 5.1.2.1). If the data for males and females
are combined for the BMC analysis, the result is <10% of the original BMCLio for males alone.
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This indicates that there is a inconsistency with at least one of the data sets in the female study.
Based on all the information, the use of male rats in the assessment is warranted.
A5.
Comment:
One reviewer continued to question the use of a UFA of 1 when in their view the
scientific data support a UFA <1.
Response:
The basis for selection of a UFA of 1 is presented in the EPA response to the comment on charge
question A5. A new rationale has not been provided by the reviewer to alter the U.S. EPA's
selection of this value. The current UFa value in the Toxicological Review is 1. The UFa
accounts for a reduction from 3 for the toxicokinetic differences between animals and humans
through the use of a PBPK model for extrapolation of doses. Likewise, the toxicodynamic
portion was reduced from 3 to 1. The use of a 1 for the toxicodynamic portion of the UFA
represents the lowest reduction described by the current guidance (U.S. EPA, 1994b). The
toxicological effect in question being the deposition of hemosiderin is a key event of the MOA
for the development of liver hepatomas in mice. To implement a further reduction in the UFA
(i.e., fractional) would logically indicate that a preponderance of data are available to fully
describe this key event (toxicodynamics) in both animals and humans to the extent of describing
why humans are less sensitive to the hemolytic effects leading to hemosiderin deposition. This is
not the case, although in vitro data (Udden and Patton, 2005; Udden, 2002, 2000; Ghanayem,
1989) do suggest humans are less sensitive than rodents to the hemolytic effects of EGBE.
Likewise, the few human studies (Haufroid et al., 1997; Carpenter et al., 1956) indicate the same
finding. However, these studies (Udden and Patton, 2005; Udden, 2002, 2000; Haufroid et al.,
1997; Ghanayem, 1989; Carpenter et al., 1956) do not characterize hemosiderin deposition, the
key event for the MOA. The current UFa value in the Toxicological Review is 1 and considers
the limited number of studies available as well as the unknown effect of EGBE on the cellular
events leading to hemolysis in human populations. In addition, little is known of the long-term
or repeated exposure responses in humans to EGBE. Lastly, part of the definition of a UF is that
it is represented by a number that is generally an order of magnitude (i.e., 10). Implementation
of UFs is described by dividing the POD by a factor of 1, 3, or 10 for each of the defined UFs to
calculate a human health toxicity value. The use of a fractional UF would represent a deviation
from the current guidance and is beyond the scope of this Toxicological Review.
A6.
Comment:
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One reviewer commented that our interpretation of his original comment on the UFd in
charge question A6 was inaccurate. The reviewer stated that he did not think the value of the
UFd should be higher but that it was inconsistent with the Agency's confidence statement.
Response:
The inaccuracies have been corrected in the reviewer responses for charge question A6.
The assigning of a value of 1 for the UFD reflects that the database contains a composite of the
basic studies needed for an overall toxicological assessment. However, U.S. EPA's conclusion
regarding the overall confidence for the derivation of a human health toxicity value (i.e., RfD,
RfC) considers the quality, strength, and adequacy of the principal study as well as the database
The selection of a confidence level of medium - high is appropriate and also considers the
potential for effects in humans from repeat, long-term exposures has not been investigated.
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APPENDIX B. CORLEY ET AL. (2005a, 1997,1994) PBPK MODELS
Corley et al. (1994) developed PBPK models for rats and humans with the primary
objective of describing the concentration of BAA in the target tissue (blood) of rats and humans
for use in risk assessment (Figure B-l). The models incorporate allometrically scalable
physiological and biochemical parameters (e.g., blood flows, tissue volumes, and metabolic
capacity) in place of the standard values for a 70-kg human. These parameters normalize
standard values to the actual body weights of the subjects in several human kinetic studies. The
physiology of humans under exercise conditions initially used by Johanson (1986) to describe
the kinetics of EGBE following inhalation exposures while exercising was maintained in the
model. The rat was included to expand the database for model validation and to assist in
interspecies comparisons of target tissue doses (BAA in blood).
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Model for 2-Butoxyethanol
Inhalation
Model for 2-Butoxyacetic Acid
IV Infusion
Exhalation
Venous B ao4-
Ven
Fat
terial Blood
Ar
erial Blood
Bio
Skin
Dermal
Vapo
/Liquid
Muscle
Metabolism to
Butoxyacetic Acid
Drinking Water
or Gavage
Other Metabolites
Urine
Liver
Liver
Kidney
Lungs and Arterial
Blood
Lungs and Arterial
Blood
Gastrointestinal
Tract
Gastrointestinal
Tract
Rapidly Perfused
Organs
Rapidly Perfused
Organs
Slowly Perfused
Organs
Slowly Perfused
Organs
Muscle
Skin
Fat
Urine
(Butoxyacetic Acid)
The formation of BAA from EGBE was assumed to occur only in the liver and
was simulated in a second model linked via the formation of BAA.
Source: Corley et al. (1994).
Figure B-l. PBPK model of Corley et al. (1994).
The Corley et al. (1994) model included additional routes of exposure such as gavage,
drinking water, i.v. infusion, and dermal (liquids and vapor) to facilitate comparisons to several
published toxicokinetic studies utilizing these routes of administration. The formation of BAA
was assumed to occur only in the liver, using the rat liver perfusion data of Johanson et al. (1986)
scaled to the human. A second model was linked to the EGBE model specifically to track the
disposition of BAA following its formation in the liver. The kidney was added to the BAA
model because it is the organ of elimination for BAA. All other metabolic routes for EGBE
(formation of EG and glucuronide conjugate) were combined, as they were used only to account
for the total disposition of EGBE in the rat metabolism studies and not for cross-species
extrapolations.
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Contrary to observations in rats, Corley et al. (1997) found no evidence of metabolites in
urine that would indicate that humans form conjugates of EGBE or EG. Thus, these pathways,
which were lumped together in the model of Corley et al. (1994) to simulate rat kinetic data,
were eliminated for human simulations. The human blood:air partition coefficient of 7,965, from
Johanson and Dynesius (1988), was also used in the Corley et al. (1994) model. In addition, the
partition coefficients for both EGBE and BAA were measured in human blood, rat blood, and rat
tissues by using a modification of the Jepson et al. (1994) technique for ultrafiltration. Human
tissue:blood partition coefficients were assumed to be equal to those of the rat. The skin:air
partition coefficient, used to calculate the dermal uptake of vapors, was assumed to be the same
as the blood:air partition coefficient. With the exception of the lung:blood partition coefficient
for EGBE (11.3), the tissue:blood partition coefficients ranged from 0.64 to 4.33 for EGBE and
from 0.77 to 1.58 for BAA. Protein binding of BAA in blood and saturable elimination of BAA
by the kidneys were necessary components to describe the BAA kinetic data in rats and humans
as discussed above. Since no direct measurements of protein binding were available, these
parameters were arbitrarily set to the molar equivalent values reported for phenolsulfonphthalein
as described by Russel et al. (1987). Constants for the saturable elimination of BAA by the
kidneys were then estimated by optimization from the data of Ghanayem et al. (1990), where rats
were administered EGBE i.v. and the concentrations of BAA in blood were determined,
following three different dose levels. These parameters were then held constant (protein
binding) or scaled by (body weight)0'74 x (renal elimination) for all simulations.
Corley et al. (1997) published the results from a human dermal exposure study and
updated their 1994 human model. This study tested the hypothesis put forward by Johanson and
Boman (1991) that dermal absorption can be more significant than inhalation absorptions by
humans exposed to EGBE vapors. The results from this study verified the predictions from the
Corley et al. (1994) model that blood samples taken via finger-prick sampling methods are
confounded by locally high concentrations of EGBE on the surface or within the skin
compartment and do not represent system blood concentrations. This study, and PBPK
simulations have more recently been verified by Franks et al. (2006) and Jones et al. (2003).
Lee et al. (1998) published an upgrade to the rat PBPK model of Corley et al. (1994) and
included female rats, male and female rats, and the effects of aging to simulate the results from
the kinetic studies conducted as part of the NTP 2-year bioassay (Dill et al., 1998). In their
model, Lee et al. (1998) adjusted the rates of metabolism of EGBE to BAA in female rats,
plasma protein binding of BAA, and the rates of renal active transport of BAA into urine to
describe the kinetics of EGBE and BAA following short-term and long-term exposures. This
model was used in the 1999 EGBE Toxicological Review to calculate the blood BAA Cmax values
in female rats, which were used as the basis for RfC determination.
Corley et al. (2005a) continued the work of Lee et al. (1998) by experimentally
measuring the blood and tissue partition coefficients for BAA, plasma protein binding of BAA,
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liver metabolism of EGBE, and renal clearance of BAA in young and old, male and female
F344 rats and B6C3Fi mice to replace assumptions used in the rat and mouse PBPK model of
Lee et al. (1998). This revised PBPK model (Figure B-2) was used as the basis for calculating
the Cmax for BAA in the blood of female rats in the subchronic NTP inhalation study as the basis
for the current RfC. The HEC was then calculated from BMD analysis of internal dose-response
for hemolysis using the Corley et al. (1997) human PBPK model. A summary of the female rat
and human model parameters are shown in Table B-l.
Butoxyethanol Butoxyacetic Acid
Slowly Pert
Organs
Fat
Skin
Exposed Skin
Dermal
Salvaiy Gland
Liver
i r
Gl Tract
BAA
A i
} r
Liver
Gl Tract
Kidney
Fat
Rapidly Pert
Organs
Stowty Pert
Organs
Skin
Gavage
BE-Gluc EG
Urine
BAA-Conj & C02
The model was based upon the PBPK model of Lee et al. (1998) with the exception that
(a) the spleen was combined with the rapidly perfused tissue group; (b) the i.p. injection
route was added to the EGBE submodel; (c) salivary glands were added; (d) the skin
compartment was separated into exposed and unexposed skin; and (e) metabolism of
BAA to conjugates or CO2 were included. Changes to the Lee et al. (1998) model are
highlighted in gray.
Figure B-2. PBPK model of Corley et al. (2005a).
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Table B-l. Selected parameters used in the PBPK model for EGBE developed
by Corley et al. (2005a, 1997)
Parameter
Human
Young female rat
Old female rat
Weights
Body weight (kg)
70
0.20
0.34
Liver (%BW)
3.14
3.22
2.52
Kidney (EGBE model)
(BAA model) (% BW)
N/A
0.44
0.69
0.69
0.694
0.694
Rapidly perfused (EGBE model)
(BAA model) (% BW)
3.71
3.27
4.39
4.39
4.39
4.39
Slowly perfused (% BW)
9.4
24.6
14.3
Flows
Alveolar ventilation (L/hr)
347.9
5.47
8.10
Cardiac output (COP) (L/hr)
347.9
5.47
8.10
Liver (% COP)
25.0
18.3
18.3
Kidney (% COP)
25.0
14.1
14.1
Rapidly perfused (EGBE model) (% COP)
(BAA model)
50.0
25.0
23.3
23.3
23.3
23.3
Slowly perfused (% COP)
2
2
2
Partition coefficients
Blood/air
7,965
7,965
7,965
Liver/blood
(BAA model)
1.46
1.30
1.48
0.66
1.48
0.66
Kidney/blood (EGBE model)
(BAA model)
1.83
1.07
1.83
0.87
1.83
0.87
Rapidly perfused/blood
(BAA model)
1.46
1.30
1.47
0.66
1.47
0.66
Slowly perfused/blood
(BAA model)
0.64
1.31
0.65
0.54
0.65
0.54
Metabolic constants
EGBE to BAA
VmaxlC (mg/hr/kg BW)
375
213
189
Kml (mg/L)
26.9
20.1
20.1
EGBE to others (human only)
Vmax2C (mg/hr/kg BW)
5
Km2 (mg/L)
0.5
EGBE to EG (rat only)
Vmax2C (mg/hr/kg BW)
3.3
3.3
Km2 (mg/L)
2.7
2.7
EGBE to glucuronide conjugate (rat only)
Vmax3C (mg/hr/kg BW)
30
30
Km3 (mg/L)
55.7
55.7
BAA to C02 (rat only)
Vmax4C (mg/hr/kg BW)
2.6
2.6
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Table B-l. Selected parameters used in the PBPK model for EGBE developed
by Corley et al. (2005a, 1997)
Parameter
Human
Young female rat
Old female rat
Km4 (mg/L)
31.8
31.8
Plasma protein binding of BAA
P (binding sites; mg/L)
164
n/a
n/a
Kd (dissociation constant; mg/L)
46
n/a
n/a
Bind (unitless fraction)
n/a
0.298
0.433
Renal elimination of BAA
VmaxEC (mg/hr/kg BW)
20.4
4.0
10.8
KmEC (mg/L)
21.9
40.0
40.0
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APPENDIX C. RFD AND RFC DERIVATION OPTIONS
C.l. RfC DERIVATIONS
C.l.l. RfC Derivations Based on Hematologic Data
The lowest NOAEL/LOAEL observed in any subchronic or chronic study of EGBE is the
31 ppm LOAEL for hematologic effects observed in the NTP study (2000) in rats. In order to
estimate a corresponding human equivalent exposure, an internal dose metric associated with this
exposure level is estimated and a PBPK model is used to estimate the human exposure level that
would result in that internal dose.
Initially, it was important to determine what estimate of internal dose (i.e., dose metric)
could serve as the most appropriate metric for the effects under consideration. PBPK models of
Lee et al. (1998) and Corley et al. (1997, 1994) are capable of calculating several measures of
dose for both EGBE and BAA, including:
CmaxC The peak concentration of EGBE or BAA in the blood during the exposure
period;
AUCC Represents the cumulative product of concentration and time for EGBE and BAA
in the blood.
Dill et al. (1998) published measured AUC, but not Cmax, blood concentrations of EGBE
and its principal metabolite BAA at various exposure durations in both genders of B6C3Fi mice
and F344 rats exposed to the same concentrations used in the NTP (2000) chronic studies. Cmax
values would need to be derived from a PBPK model. Two pieces of information were used to
select Cmax for BAA in the blood as the more appropriate dose metric for the main hemolytic
endpoint associated with this LOAEL. First, as discussed in Section 4.5, there is convincing
evidence to indicate that an oxidative metabolite, BAA, is the causative agent for EGBE-induced
hemolysis (Ghanayem et al., 1990, 1987b; Carpenter et al., 1956). With this in mind, dose
metrics for BAA in blood appear to be more appropriate than those for EGBE in blood, since
they are more closely linked mechanistically to the toxic response. Second, EGBE-induced
hemolysis appears to be dependent upon the dose rate. Ghanayem et al. (1987c) found that
gavage doses to F344 male rats of 125 mg/kg EGBE resulted in hemolytic effects including
reduced RBC count, Hb, and Hct, as well as kidney pathology (Hb casts and intracytoplasmic
Hb). However, it should be noted that hemolytic effects were not reported at a similar acute
drinking water dose of 140 mg/kg (Medinsky et al., 1990). While a drop in RBC count and Hb
(9 and 7%, respectively) was noted in F344 male rats after 1 week of drinking water exposure to
129 mg/kg-day EGBE, dose-related kidney pathology was not observed in these rats, even after
13 weeks of drinking water exposure up to 452 mg/kg-day EGBE (NTP, 1993). Consistent with
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the hypothesis that exposure concentration plays a larger role than exposure duration for EGBE-
induced hemolytic effects, hematological endpoints indicative of hemolysis do not progress with
increased inhalation duration (Section 5.1.1). Corley et al. (1994) also suggested that Cmax is a
more appropriate dose metric for the hemolytic effects of EGBE than AUC.
The PBPK models developed for EGBE are briefly summarized in Table C-l. Shyr et al.
(1993) and Johanson (1986) do not address BAA distribution, and are only parameterized for
humans and rats, respectively. In the 1999 EGBE Toxicological Review, the model described by
Lee et al. (1998) was determined to be the most appropriate model for the estimation of rat and
mouse internal doses following inhalation exposure. Since the 1999 Toxicological Review,
Corley et al. (2005a) published a revision to the Lee et al. (1998) model for rats and mice where
several assumptions used by Lee et al. (1998) were replaced with measured values (e.g., protein
binding, partition coefficients, metabolism rate constants for multiple pathways, and renal
clearance) as a function of species, gender, and age. That model is used here to estimate the Cmax
of BAA in blood following inhalation exposure to female rats, the more sensitive gender. For
transparency, the results from both PBPK models are presented in Table C-2 below. However,
only the values derived from the Corley et al. (2005a) model were utilized in the derivation of
the LOAEL and BMD in Sections C. 1.1.1 and C. 1.1.2. The human PBPK model of Corley et al.
(1997, 1994) was then used to obtain estimates of human inhalation exposure concentrations
associated with the female rat BAA blood concentrations.10 Established U. S. EPA (2006c)
methods and procedures were used to review, select, and apply these chosen PBPK models.11
Table C-l. Summary of PBPK models
Model
Species
Routes of exposure
Comments
Johanson (1986)
Human
Inhalation
BAA not addressed
Shyretal. (1993)
Rat
Inhalation, oral, dermal
BAA excretion
Corley et al. (1997,
1994)
Rat and human
Inhalation, oral,
dermal, i.v.
BAA distribution and excretion; male rats
only
Lee et al. (1998)
Rat and mouse
Inhalation
BAA distribution and excretion; males and
females
Corley et al. (2005a)
Rat and mouse
Inhalation, oral,
dermal, i.p., i.v.
Age-dependent BAA distribution,
metabolism, and excretion, males and females
Franks et al. (2006)
Human
Inhalation and dermal
Extended Corley et al. (1997) model to
include bladder compartment for human
biomonitoring studies
10The basic components of the Corley model are summarized in Appendix B.
nEPA notes that the review of the PBPK models was conducted prior to their use in the 1999 EGBE toxicological
review.
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Table C-2. Model estimates of BAA blood levels in female rats following
inhalation exposures
Exposure
concentration (ppm)
Female rat body weight (g)
BAA in arterial blood Cmax in
female rats (juIVl)
Lee et al. (1998)
BAA in arterial blood Cmax
in female rats (juIVl)
Corley et al. (2005a)
31
216
285
167
61.5
211
603
408
125
214
1,243
1,091
250
210
1,959
2,752
500
201
4,227
6,483
Lee et al. (1998) model results used in the 1999 EGBE Toxicological Review are included for comparison to the
updated model of Corley et al. (2005a).
C.l.1.1. NOAEL/LOAEL Method Applied to Hematologic Data
A five-step procedure was used to calculate the LOAEL HEC:
Step 1: Calculate the internal dose surrogate (Cmax for BAA in blood) corresponding to
female rat LOAEL (Corley et al., 2005a) by using the actual experimental exposure
regimen (6 hours/day, 5 days/week) in model simulations.
Female rat LOAEL = 31 ppm
Cmax BAA = 167 [iM
Step 2: Verify that steady state was achieved (e.g., no change in BAA Cmax as a result of
prolonging the exposure regimen).
There were no changes in the Cmax of BAA in blood during any 24-hour simulation
period, using a 6 hour/day, 5 day/week exposure regimen at the female rat LOAEL,
indicating that steady state was achieved.
Step 3: Simulate the internal dose surrogate by calculating the Cmax for BAA in blood for
humans continuously exposed (24 hours/day, 7 days/week) to varying concentrations
of EGBE (Corley et al., 1997, 1994).
Concentration of EGBE
Cmax BAA in blood
in air (ppm)
(uM)
1
2.6
5
13.0
10
26.1
20
52.9
50
137.1
100
295.0
200a
732.3
"Steady-state concentrations are not achieved within 5 days of continuous exposure to EGBE
concentrations >200 ppm.
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Step 4: Calculate the LOAELHec of EGBE for continuous human exposure in air that
resulted in the same internal dose (Cmax of BAA) in blood calculated for the animal
study in step 1.
Female rat Cmax BAA= 167 |iM
HEC continuous exposure = 56 ppm (calculated by regression of internal dose
versus the concentration of EGBE in air from step 3).
Step 5: Convert the EGBE exposure units from ppm to mg/m3
LOAELhec (mg/m3) = conversion factor x LOAELHec (ppm)
= 4.84 (mg/m3) ^ (ppm) x 56 ppm
= 271 mg/m3
C.l.1.2. BMC Method Applied to Hematologic Data
For the purposes of deriving an RfC for EGBE from hemolytic endpoints, both MCV and
RBC count response data were evaluated in female rats from the 14-week subchronic NTP
(2000) study (see Section 5.1.1). BMCs derived for these same hemolytic endpoints in male rats
of this study were approximately twofold higher than for female rats (data not shown).
The current BMD technical guidelines (U.S. EPA, 2000b) suggest the use of 1 SD from
the control mean for the BMR level for continuous data in the absence of additional information,
such as a minimal level of change in the endpoint that is generally considered to be biologically
significant. Because the hemolytic endpoints are continuous measurements that have a relatively
small historical variance in rats, and because low-dose responses for these endpoints were
generally near or within 5% of the control mean, the BMC Los was considered to be a more
appropriate POD for derivation of the RfC (U.S. EPA, 2000b, 1995). The steepest concentration-
response curves (and the lowest BMC Los estimate) were obtained for decreased RBC count in
female rats, and a 5% change was found to be statistically significant. Higher levels of response
(e.g., >10% reductions) are in the exposure range where other more severe responses related to
anemia occur, such as MCV increases and increased reticulocyte counts. Lower levels of
response, for instance, 1 SD from the control mean, an approximate 2% reduction for these data,
are relatively distant from the observable data and other responses related to anemia. Cmax for
BAA in arterial blood of rats was determined by using the PBPK model of Corley et al. (2005a).
Dermal exposure to EGBE vapor was not considered in the predicted blood levels because the
estimated relative contribution of the skin to the total uptake of unclothed humans exposed to
25 ppm EGBE for 8 hours ranged from only 4.6 to 27.5%, depending on temperature, humidity,
and exercise level (Corley et al., 1997). Thus, dermal uptake is predicted to contribute <10%,
even if 50% of an individual's skin is exposed. The results of this modeling effort are
summarized in Table C-2.
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All BMD analyses were performed using models in U.S. EPABMDS, version 1.4.1c
(U.S. EPA, 2000b). The fit statistics and BMC information derived from the continuous models
available in the BMD software as applied to the female rat RBC count data versus Cmax BAA are
shown in Table C-3. The best model fit to these data, from visual inspection and comparison of
AIC values, was obtained using the Hill model. The BMCLos was determined to be 133 [xM
using the 95% lower confidence limit of the dose-response curve expressed in terms of the Cmax
for BAA in blood. The graphic and textual output from the Hill model run is displayed after
Table C-3. BMCLSd values are provided for comparative purposes. The Corley et al. (1997)
PBPK model was used to back-calculate a HEC of 46.5 ppm (225 mg/m3) from the 133 [xM
BMCL05, assuming continuous exposure (24 hours/day).
Table C-3. Comparison of BMC/BMCL values for female rat RBC count
data from a 14-week subchronic inhalation studya, using modeled blood Cmax
(3 months) of the EGBE metabolite BAA as a common dose metric
Model
BMC 05 (jtM)
BMCL0S
(jiM)
bmcsd
(jiM)
bmclsd
(jiM)
/j-Value
AICa
Scaled
residualb
Polynomial (2°)
495.468
460.317
357.566
301.089
<0.0001
-66.057235
-1.51
Power0
119.313
82.782
29.4231
16.5251
0.022
-106.196640
1.29
Hillcd
189.394
133.005
72.9703
38.8183
0.621
-111.611016
-0.287
aTo obtain adequate model fits, the high dose group data was dropped from the analysis.
bAIC = -2L + 2P, where L is the log-likelihood at the maximum likelihood estimates for the parameters and P is the
number of model degrees of freedom (usually the number of parameters estimated).
c/2 residual (measure of how model predicted response deviates from the actual data) for the dose group closest to
the BMC scaled by an estimate of its SD. Provides a comparative measure of model fit near the BMC. Residuals
that exceed 2.0 in absolute value should cause questioning of the model fit in this region.
dPower and Hill models were run with power terms unrestricted to obtain adequate fit; estimates of the power terms
were 0.56 and 0.97 for the Power and Hill model, respectively.
"Model choice based on adequate p-valuc (>0.1), visual inspection, low AIC, and low (absolute) scaled residual.
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Hill Model with 0.95 Confidence Level
a)
CO
c
o
Q.
CO
a)
a:
c
ro
a)
0 500 1000 1500 2000 2500
dose
1 16:26 02/27 2009
Hill Model. (Version: 2.14; Date: 06/26/2008)
Input Data File: C:\Usepa\BMDS21Beta\Data\EGBE\femalerat\2Hilrbchil.(d)
Gnuplot Plotting File: C:\Usepa\BMDS21Beta\Data\EGBE\femalerat\2Hilrbchil.plt
FriFeb 27 16:26:00 2009
BMDS Model Run
The form of the response function is:
Y[dose] = intercept + v*doseAn/(kAn + doseAn)
Dependent variable = rbc
Independent variable = Cmax
rho is set to 0
Power parameter is not restricted
A constant variance model is fit
Total number of dose groups = 5
Total number of records with missing values = 0
Hill
8.5
7.5
6.5
BMDL BMD
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Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
alpha = 0.0357308
rho = 0 Specified
intercept = 8.48
y = -2.41
n = 1.34368
k = 775.437
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -rho
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
alpha intercept v n k
alpha 1 9.2e-008 -5.9e-008 -8.5e-008 2.5e-008
intercept 9.2e-oos 1 -o.36 -0.47 0.23
V -5.9e-008 -0.36 1 0.95 -0.99
n -8.5e-008 -0.47 0.95 1 -0.92
k 2.5e-008 0.23 -0.99 -0.92 1
Parameter Estimates
95.0% Wald Confidence Interval
Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. Limit
alpha 0.0323151 0.00646302 0.0196478 0.0449824
intercept 8.47617 0.0566598 8.36512 8.58722
V -3.86075 0.809864 -5.44805 -2.27344
n 0.971125 0.157912 0.661623 1.28063
k 1634.54 734.333 195.274 3073.81
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Table of Data and Estimated Values of Interest
Dose N Obs Mean Est Mean Obs Std Dev Est Std Dev Scaled Res.
0 10 8.48 8.48 0.158 0.18 0.0674
167 10 8.08 8.1 0.221 o.is -0.287
408 10 7.7 7.68 0.253 0.18 0.353
1091 10 6.91 6.92 0.158 0.18 -0.175
2752 10 6.07 6.07 0.126 0.18 0.0415
Model Descriptions for likelihoods calculated
Model A1: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)A2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} — SigmaA2
Model A3 uses any fixed variance parameters that
were specified by the user
Model R: Yi = Mu + e(i)
Var{e(i)} = SigmaA2
Likelihoods of Interest
Model Log(likelihood) # Part
A1 60.927568
A2 64.091535
A3 60.927568 6
fitted 60.805508
R -18.641036 2
's AIC
6 -109.855137
10 -108.183070
-109.855137
5 -111.611016
41.282072
Explanation of Tests
Test 1: Do responses and/or variances differ among Dose levels?
(A2 vs. R)
Test 2: Are Variances Homogeneous? (A1 vs A2)
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Test 3: Are variances adequately modeled? (A2 vs. A3)
Test 4: Does the Model for the Mean Fit? (A3 vs. fitted)
(Note: When rho=0 the results of Test 3 and Test 2 will be the same.)
Tests of Interest
Test -2*log(Likelihood Ratio) Test df p-value
Test 1 165.465 8 <.0001
Test 2 6.32793 4 0.176
Test 3 6.32793 4 0.176
Test 4 0.244121 1 0.6212
The p-value for Test 1 is less than .05. There appears to be a
difference between response and/or variances among the dose levels
It seems appropriate to model the data
The p-value for Test 2 is greater than . 1. A homogeneous variance
model appears to be appropriate here
The p-value for Test 3 is greater than .1. The modeled variance appears
to be appropriate here
The p-value for Test 4 is greater than . 1. The model chosen seems
to adequately describe the data
Benchmark Dose Computation
Specified effect = 0.05
Risk Type = Relative risk
Confidence level = 0.95
BMD = 189.394
BMDL = 133.005
1
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C.1.2. BMC Method Applied to Hemosiderin Data
For the purposes of deriving an RfC for EGBE, hemosiderin staining data were evaluated
in male and female rats from the 2-year chronic study by NTP (2000). The current BMD
technical guidelines (U.S. EPA, 2000b) suggest the use of 10% extra risk as a BMR level for
quantal data as this is at or near the limit of sensitivity in most cancer bioassays and in some
noncancer bioassays as well. Because the hemosiderin staining endpoint was observed in control
animals and a 10% increase in incidence was within the observable range of the data, 10% extra
risk was considered an appropriate BMR and a BMCLio an appropriate POD for derivation of
the RfC (U.S. EPA, 2000b, 1995). All BMD assessments in this review were performed using
U.S. EPA BMDS version 1.4.1c.
The AUC was selected as the appropriate dose metric due to the nature of the endpoint,
hemosiderin deposition. This endpoint increased in severity with increased duration (subchronic
to chronic) and is believed to be the result of the cumulative exposure to EGBE as opposed to a
peak event. Table C-4 reports AUC BAA blood concentrations measured at 12 months12
published by Dill et al. (1998) in both genders of B6C3Fi mice and F344 rats exposed to the
same concentrations used in the NTP (2000) chronic studies of these test animals.
Table C-4. AUC BAA blood concentrations measured at 12 months in both
genders of B6C3Fi mice and F344 rats
Exposure concentration (ppm)
Gender
AUCbaa (jimol-hr/L)a
n
Mean
SE
Rats
31.2
Male
7
358.3
16.6
Female
5
638.8
18.7
62.5
Male
6
973.0
86.2
Female
9
1,128.9
50.9
125
Male
9
2,225.6
71.1
Female
12
3,461.8
154.8
Mice
62.5
Male
10
1,206.6
205.6
Female
12
1,863.6
112.4
125
Male
9
2,819.8
685.1
Female
6
5,451.6
508.9
250
Male
10
17,951.5
1,770.4
Female
11
18,297.1
609.7
aAuthors reported AUC values in terms of (ig-min/g, which were converted to units consistent with the PBPK model
of |imol-hr/L by dividing by 60 min/h and 132.16 g/mol and multiplying by 1,060 g/L.
Source: Dill et al. (1998).
12Dill et al. (1998) also reported 18 month data, but due to the smaller number of animals and higher variability in
this data the 12 month data were used for the purposes of this analysis.
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The fit statistics and BMC information derived from the dichotomous models available in
the BMD software as applied to the male and female rat hemosiderin staining data versus AUC
BAA are shown in Table C-5. All models were fit using restrictions and option settings
suggested in the U.S. EPA BMD technical guidance document (U.S. EPA, 2000b). The best
model fit to these data, as determined by visual inspection, examination of low dose model fit
(i.e., scaled residual for the dose group closest to the BMD), and comparison of overall fit (i.e.,
AIC values) was obtained using a multistage model (1st degree) for the male response data and a
Log-Logistic model for the female response data. The male rat BMCio was 196 [j,mol-hour/L
and the BMCLio was determined to be 133 [j,mol-hour/L, using the 95% lower confidence limit
of the dose-response curve expressed in terms of the AUC for BAA in blood. The BMCio and
BMCLio values for the female rat were determined to be 425 and 244 [j,mol-hour/L, respectively.
The graphic and textual output from the model runs that resulted in these male and female rat
BMC and BMCLio estimates are displayed below, after Table C-6. .Assuming continuous
exposure (24 hour/day), the Corley et al. (1997) PBPK model was used to back-calculate HECs
of 3.4 ppm (16 mg/m3) from the male rat data and 4.9 ppm (24 mg/m3) from the female rat data.
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Table C-5. Comparison of BMC/BMCL values for male and female rat liver
hemosiderin staining data from inhalation chronic study using measured
blood AUC (12 months) of the EGBE metabolite BAA as a common dose
metric
Model
BMCio (jimol-hr/L)
BMCLio (nmol-hr/L)
/7-Value
AICa
Scaled residualb
Male rats
Multistage-lst degree0
196.252
133.141
0.8680
247.234
0.441
Gamma0
196.253
133.141
0.8680
247.234
0.441
Logistic
259.296
192.773
0.7692
247.476
0.526
Log-Logistic
166.376
69.3279
0.5623
249.283
0.313
Probit
271.525
205.882
0.7450
247.54
0.517
Log-Probit
368.336
241.992
0.6309
247.876
0.765
Weibull0
196.253
133.141
0.8680
247.234
0.441
Female rats
Multistage-lst degree
122.166
214.555
0.0698
218.868
-1.945
Gamma
316.635
134.02
0.0554
219.229
-1.238
Logistic
273.693
221.689
0.0993
218.188
-1.294
Log-Logisticc
424.527
243.69
0.1533
217.526
-0.896
Probit
291.017
241.206
0.0683
218.985
-1.260
Log-Probit
427.728
248.683
0.1238
217.884
-0.965
Weibull
266.515
130.801
0.0454
219.58
-1.377
AIC = -2L + 2P, where L is the log-likelihood at the maximum likelihood estimates for the parameters and P is the
number of model degrees of freedom (usually the number of parameters estimated)
b%2 residual (measure of how model predicted response deviates from the actual data) for the dose group closest to
the BMC scaled by an estimate of its SD. Provides a comparative measure of model fit near the BMC. Residuals
that exceed 2.0 in absolute value should cause one to question model fit in this region.
°Model choice based on adequate p-valuc (>0.1), visual inspection, low AIC, and low (absolute) scaled residual.
The Multistage (1st degree) is referred to as the chosen model for male rats, though equivalent fit was obtained by
the restricted Gamma and Weibull models.
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Table C-6. Comparison of BMC/BMCL values for male and female mouse
liver hemosiderin staining data from inhalation chronic study using
measured blood AUC (12 months) of the EGBE metabolite BAA as a
common dose metric
Model
BMC10
(nmol-hr/L)
BMCL10
(nmol-hr/L)
/7-Value
AICa
Scaled residualb
Male mice
Multistage-1st
degree
2,100.07
1,613.9
0.3067
117.571
-1.766
Gamma
2,725.35
1,702.27
0.1452
118.559
1.358
Logistic
6,605.45
5,333.72
0.0022
127.326
2.789
Log-logistic
2,616.51
1,628.48
0.1882
118.02
1.193
Probit
5,917.06
4,825.09
0.0031
126.405
2.734
Log-probitc
3,076.8
2,448.3
0.1290
116.614
1.946
Weibull
2,689.76
1,687.09
0.1445
118.712
-1.448
Female mice
Multistage-1st
degree
946.491
769.879
0.3680
142.669
-1.583
Gamma
1,402.92
818.367
0.3420
143.288
-0.817
Logistic
2,897.15
2,341.03
0.0002
162.338
-0.942
Log-logistic
1,705.75
1,121.43
0.8223
141.501
-0.343
Probit
2,860.03
2,364.52
0.0002
161.681
-0.829
Log-probitc
1,734.53
1,322.06
0.8237
141.498
-0.315
Weibull
1,282.82
804.234
0.2958
143.631
-0.988
aAIC = -2L + 2P, where L is the log-likelihood at the maximum likelihood estimates for the parameters and P is the
number of model degrees of freedom (usually the number of parameters estimated).
b%2 residual (measure of how model predicted response deviates from the actual data) for the dose group closest to
the BMC scaled by an estimate of its SD. Provides a comparative measure of model fit near the BMC. Residuals
that exceed 2.0 in absolute value should cause one to question model fit in this region.
°Model choice based on adequate p-valuc (>0.1), visual inspection, low AIC, and low (absolute) scaled residual.
The Log-probit model provides a slightly better fit than other models for both genders.
Likewise, the fit statistics and BMC information for male and female mouse hemosiderin
staining data versus AUC BAA are shown in Table C-6. All models were fit using restrictions
and option settings suggested in the U.S. EPABMD technical guidance document (U.S. EPA,
2000b). The best model fit to these data, as determined by visual inspection, examination of low
dose model fit (i.e., scaled residual for the dose group closest to the BMD), and comparison of
overall fit (i.e., AIC values) was obtained using a log-probit model for both the male and female
response data. The male mouse BMCio was 3,077 [j,mol-hour/L and the BMCLio was
determined to be 2,448 [xmol-hour/L using the 95% lower confidence limit of the dose-response
curve expressed in terms of the AUC for BAA in blood. The BMCio and BMCLio values for the
female mouse were determined to be 1,735 and 1,322 [j,mol-hour/L, respectively. Assuming
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continuous exposure (24 hour/day), the Corley et al. (1997) PBPK model was used to back-
calculate HECs of 36 ppm (174 mg/m3) from the male mouse data and 20 ppm (97 mg/m3) from
the female mouse data.
Multistage Model with 0.95 Confidence Level
Multistage
0.9
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BMDL BMD
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Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: U:\BMDS\DATA\EGBE\RAT-HEMOSID-12MTH-DILL-
AUC\MALE-MULT 1. (d)
Gnuplot Plotting File: U:\BMDS\DATA\EGBE\RAT-HEMOSID-12MTH-DILL-
AUC\M ALE-MULT 1 pit
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BMD Method for RfC: Hemosiderin deposition in male rats versus AUC BAA, 2 year
inhalation study (NTP, 2000)
The form of the probability function is:
P[response] = background + (l-background)*[l-EXP(
-betal*doseAl)]
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The parameter betas are restricted to be positive
Dependent variable = Hemo M
Independent variable = AUC-M-Dill
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
Background = 0.481866
Beta(l) = 0.000526977
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.64
Beta(l) -0.64 1
Parameter Estimates
95.0% Wald Confidence Interval
Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. Limit
Background 0.477747 * * *
Beta(l) 0.000536865 * * *
* - Indicates that this value is not calculated.
Analysis of Deviance Table
Model Log(likelihood) #Param's Deviance Testd.f. P-value
Full model -121.475 4
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Fitted model -121.617 2 0.28443 2 0.8674
Reduced model -130.097 1 17.2453 3 0.0006292
AIC: 247.234
Goodness of Fit
Scaled
Dose Est.Prob. Expected Observed Size Residual
0.0000
0.4777
23.887
23
50
-0.251
358.3000
0.5691
28.457
30
50
0.441
973.0000
0.6902
34.512
34
50
-0.157
2225.6000
0.8419
42.094
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-0.037
ChiA2 = 0.28 d.f. = 2 P-value = 0.8680
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 196.252
BMDL = 133.141
BMDU = 342.382
Taken together, (133.141, 342.382) is a 90 % two-sided confidence
interval for the BMD
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Log-Logistic Model with 0.95 Confidence Level
1
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Logistic Model. (Version: 2.10; Date: 09/23/2007)
Input Data File: U:\BMDS\DATA\EGBE\RAT-HEMOSID-12MTH-DILL-
AUC\FEMALE-LOGLOG.(d)
Gnuplot Plotting File: U:\BMDS\DATA\EGBE\RAT-HEMOSID-12MTH-DILL-
AUCVFEMALE-LOGLOG.plt
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BMD Method for RfC: Hemosiderin deposition in female rats versus AUC BAA, 2 year
inhalation study (NTP, 2000)
The form of the probability function is:
P[response] = background+(l-background)/[l+EXP(-intercept-slope*Log(dose))]
Dependent variable = Hemo F
Independent variable = AUC-F-Dill
Slope parameter is restricted as slope >= 1
Total number of observations = 4
Total number of records with missing values = 0
Log-Logistic
BMDL
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Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
User has chosen the log transformed model
Default Initial Parameter Values
background = 0.3
intercept = -17.7354
slope = 2.49241
Asymptotic Correlation Matrix of Parameter Estimates
background intercept slope
background 1 -0.24 0.19
intercept -0.24 1 -1
slope 0.19 -1 1
Parameter Estimates
Variable
background
intercept
slope
95.0% Wald Confidence Interval
Estimate
0.280971
-16.4459
2.35478
Std. Err.
*
*
Lower Conf. Limit Upper Conf. Limit
* *
- Indicates that this value is not calculated.
Analysis of Deviance Table
Model Log(likelihood) #Param's Deviance Testd.f. P-value
Full model -104.742 4
Fitted model -105.763 3 2.04091 1 0.1531
Reduced model -135.725 1 61.9658 3 <.0001
AIC: 217.526
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Goodness of Fit
Scaled
Dose Est.Prob. Expected Observed Size Residual
0.299
-0.896
0.903
-0.575
ChiA2 = 2.04 d.f. = 1 P-value = 0.1533
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 424.527
BMDL = 243.69
0.0000 0.2810
638.8000 0.4430
1128.9000 0.6595
3461.8000 0.9566
14.049 15 50
22.149 19 50
32.974 36 50
47.828 47 50
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C.1.3. BMD Method Applied to Forestomach Lesions in Female Mice
Cmax of blood BAA is considered a reasonable measure of internal dose for the
forestomach lesions reported from EGBE exposure. BAA is believed to be the toxic moiety
responsible for the forestomach effects observed following EGBE exposure, with concentration
(rather than AUC) appearing to be determinative in the development of these effects as well
(Corley et al., 2005b). Other information supportive of the Cmax as an appropriate metric include
the findings of NTP (1993), where no signs of forestomach irritation were observed in mice at
very high dose levels of 1,400 mg/kg-day in 2-week and 13-week drinking-water studies
conducted by NTP (1993). It is likely that such oral nonbolus dosing of EGBE does not result in
high enough local concentrations of EGBE and BAA (Poet et al., 2003) to produce irritation.
Consistent with this observation are results with other forestomach carcinogens that are not
mutagenic, demonstrating that forestomach effects are dependent not only on the dose but also
on the chemical concentration in the dosing solution (Ghanayem et al., 1985).
The blood level of BAA is a valid surrogate for the local (i.e., forestomach) BAA
concentration as blood levels of BAA follow the severity of the irritant and irritant-associated
hyperplastic responses in the forestomach. Further, this relationship between blood BAA levels
and irritant response holds true for routes of EGBE administration other than oral, (e.g.,
inhalation, i.p., and i.v.) (Green et al., 2002; Corley et al., 1999). The basis for this route-
independent response may be related to the Green et al. (2002) results from whole-body
autoradiography of mice exposed via inhalation, showing appreciable amounts of EGBE-
associated radioactivity being present in salivary glands and ducts.
Plausible evidence exists for considering the incidence of forestomach tumors in female
mice, following chronic inhalation exposure to EGBE, to occur via a nonlinear, nongenotoxic
MOA. EGBE appears to be one of a group of compounds that are not mutagenic but can
indirectly cause forestomach tumors through the sustained cytotoxicity and cell regeneration
brought about by irritant and irritant-associated hyperplastic effects and breakdown of the
forestomach's gastric mucosal barrier. This sequence of events is considered obligate for the
formation of the observed neoplasms. Strategies intended to control or omit any of these key
mode-of-action events, including the initial hyperplastic event, would interrupt the process and
prevent formation of neoplasms. While this MOA may be of qualitative relevance to humans,
the exposure concentrations that would be necessary to cause these hyperplastic effects and
resultant tumors in humans, if attainable, are likely to be much higher than the concentrations
necessary to cause forestomach effects in mice, primarily because humans lack a comparable
organ for storage and long-term retention of EGBE.
Another line of reasoning that may be used to address issues relating to the occurrence of
these irritant/hyperplastic lesions (and potential for progression) in humans is to order the dose-
response among those effects noted for EGBE, most prominently the hematologic effects that
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underlie the hepatic tumors and that are the basis of the oral RfD and inhalation RfC. A BMD
analysis of the irritant/hyperplastic response observed in the NTP (2000) follows.
The endpoint used in this analysis was epithelial hyperplasia of the female mouse
forestomach, since it was the most sensitive forestomach effect observed in the NTP (2000)
study. Consistent with the 1999 IRIS assessment, four steps were employed to estimate human
equivalent oral and inhalation benchmark exposures from this endpoint: (1) estimate a BMDLio
value using modeled "end-of-the-week" internal dose (Cmax BAA in blood) levels; (2) verify that
steady state was achieved (e.g., no change in BAA Cmax as a result of prolonging the exposure
regimen); (3) simulate the internal dose surrogate (Cmax BAA in blood) for humans (continuous
air exposure; drinking water assumption was that a 70-kg human consumes an average of 2 L of
water during a 12-hour awake cycle); and (4) calculate the HED/concentration that resulted in
the same internal dose (Cmax BAA) simulated for the animal in step 1.
Step 1: Estimation of BMDLio (Cmax) dose.
Cmax for BAA in arterial blood was determined using the PBPK model of Lee et al.
(1998). The model results and incidence data for the endpoint of concern are summarized in
Table C-7.
Table C-7. PBPK model estimates of BAA Cmax blood levels and incidence
of forestomach epithelial hyperplasia in female mice
Air concentration (ppm)
Cmax BAA (nM)
Incidence of forestomach hyperplasia
0
0
0/50
62.5
529
6/50
125
1,200
27/50
250
2,620
42/49
BMD and BMDLio estimates were derived using the available models in version 1.3.2 of
the U.S. EPABMDS. The estimates for each model, along with statistical goodness-of-fit
information, are provided in Table C-8.
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Table C-8. BMDS model estimates of Cmax BMDio and BMDLio values for
forestomach epithelial hyperplasia in female mice
BMDS model
BMD (jiM)
BMDL (jiM)
AICa
(lowest = best fit)
/7-Value
(>0.1 = adequate fit)
Gamma
420.56
266.87
151.16
0.5287
Logistic
544.757
444.896
162.191
0.0067
Log-logisticb
462.513
329.04
150.153
0.8717
Multistage (1st degree)
177.442
145.713
156.244
0.0648
Multistage (2nd degree)
338.483
202.437
152.681
0.0976
Multistage (3rd degree)
338.485
197.436
152.681
0.2535
Probit
525.521
430.612
161.304
0.0086
Log-probit
470.876
344.412
150.163
0.8673
Weibull
376.085
238.952
151.855
0.3807
aAIC = -2L + 2P, where L is the log-likelihood at the maximum likelihood estimates for the parameters and P is the
number of model degrees of freedom (usually the number of parameters estimated).
bModel choice based on adequate p-valuc (>0.1), visual inspection, low AIC, and low (absolute) scaled residual.
The Log-logistic model provides a slightly better fit than other models.
Step 2: Verification of steady state.
As can be seen from Table C-9, Cmax levels are relatively constant through 6 months then
increase at and beyond 12 months, presumably due to clearance problems in aging animals.
However, the earlier steady-state levels are appropriate for use in this assessment, because that is
the more conservative approach, and because similar effects were observed during the
subchronic portion of the NTP (2000) study at the same dose levels, indicating that the higher
internal doses at and beyond 12 months were not required for the effects to appear.
Table C-9. Female mouse Cmax values for various time points of the NTP
(2000) study estimated by the Lee et al. (1998) model
Mos on study
62.5 ppm
125 ppm
250 ppm
Male
Female
Male
Female
Male
Female
1
403
529
921
1,200
2,080
2,620
3
402
527
925
1,202
2,120
2,652
6
399
523
914
1,184
2,071
2,582
12
484
639
1,079
1,414
2,349
2,951
16
643
849
1,443
1,839
2,798
3,501
18
756
995
1,625
2,102
3,067
3,803
Step 3: Simulation of internal human doses.
The tables below summarize the results of model simulations of the internal dose
surrogate (Cmax BAA in blood) for a 70-kg human who consumes an average of 2 L of drinking
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water during a 12-hour awake cycle (Table C-10) or is continuously exposed to air
concentrations (Table C-ll) of EGBE.
Table C-10. Estimated Cmax for BAA in blood for humans continuously
exposed to varying drinking water concentrations of EGBE
EGBE concentration in water (ppm)
Calculated dose of EGBE from drinking
water (mg/kg-d)
Cmax BAA
in blood (jliIVI)
24
0.7
9
48
1.4
18
94
2.7
36
188
5.4
73
375
10.7
147
750
21.4
299
Sources: Corley etal. (1997, 1994).
Table C-ll. Estimated Cmax for BAA in blood for humans continuously
exposed to varying concentrations of EGBE
Concentration of EGBE in air (ppm)
Cmax BAA in blood (juIVI)
1
2.6
5
13.0
10
26.1
20
52.9
50
137.1
100
295.0
200
733.7
Sources: Corley etal. (1997, 1994).
Step 4: Calculate the HED/concentration
The Corley et al. (1997, 1994) PBPK model was used to back-calculate a human
equivalent oral dose of 23.6 mg/kg-day from the Cmax BMDLio of 329 [jM estimated in step 1,
assuming that mice and humans receive their entire dose of EGBE from drinking water over a
12-hour period each day. The Corley et al. (1997) PBPK model was used to back-calculate a
human equivalent air concentration of 551 mg/m3 (113 ppm) from the Cmax BMDLio of 320 [xM
estimated in step 1, assuming continuous exposure (24 hours/day).
The PODs calculated above are significantly higher than the PODs of 1.4 mg/kg-day and
16 mg/m3 used to derive the RfD and RfC, respectively. Thus, these results indicate that the RfD
and RfC values for EGBE, which were based on hemosiderin accumulation due to hemolytic
effects in rats, should be adequate for the prevention of gastrointestinal hyperplastic effects as
well.
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BMDL
Log-logistic Model with 0.95 Confidence Level
Concentration (mM)
500 1000 1500 2000 2500
Log-logistic
Logistic Model $Revision: 2.1 $ $Date: 2000/02/26 03:38:20 $
Input Data File: F:\BMDS\DATA\EGBE\F_MOUSE_HYP_LOG-LOGIST.(d)
Gnuplot Plotting File: F:\BMDS\DATA\EGBE\F_MOUSE_HYP_LOG-LOGIST.plt
Fri Jul 11 19:53:31 2003
Log-Logistic model run for female mice with forestomach epithelial hyperplasia following
inhalation exposure (NTP, 2000) versus internal dose metric (BAA Cmax, jiM).
The form of the probability function is:
P[response] = background+(l-background)/[l+EXP(-intercept-slope*Log(dose))]
Dependent variable = F Hyperplasia
Independent variable = Dose
Slope parameter is restricted as slope >= 1
Total number of observations = 4
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
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Parameter Convergence has been set to: le-008
User has chosen the log transformed model
Default Initial Parameter Values
background = 0
intercept = -16.768
slope = 2.36735
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -background
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
intercept slope
intercept 1-1
slope -11
Parameter Estimates
Variable Estimate Std. Err.
background 0 NA
intercept -16.7132 2.64108
slope 2.36545 0.372243
NA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
has no standard error.
Analysis of Deviance Table
Model Log(likelihood) Deviance Test DF P-value
Full model -72.9391
Fitted model -73.0765 0.274637 2 0.8717
Reduced model -131.841 117.804 3 <0001
AIC: 150.153
Goodness of Fit
Scaled
Dose Est. Prob. Expected Observed Size Residual
0.0000 0.0000 0.000 0 50 0
529.0000 0.1324 6.622 6 50 -0.2596
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1200.0000 0.5145 25.725 27 50 0.3608
2620.0000 0.8705 42.653 42 49 -0.2777
Chi-square = 0.27 DF = 2 P-value = 0.8717
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 462.513
BMDL = 329.04
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C.2. RfD DERIVATIONS
C.2.1. RfD Derivations Based on Hematologic Data
PODs for the RfD derivation in terms of the HEDs have been calculated via the
application of PBPK modeling to NOAEL and BMDL estimates.
Of the available PBPK models (Table C-l), the Corley et al. (2005a, 1997) model is
considered the most complete and appropriate for use in the derivation of the oral RfD because it
has been experimentally validated for the most sensitive species (rats) and humans, it covers both
oral and inhalation routes of exposure, and it addresses both the distribution and excretion of the
toxic metabolite, BAA, following oral EGBE exposure. This model is summarized in
Appendix B. As in the case of the RfC (see Section C. 1.1), Cmax is considered a more
appropriate dose metric than AUC for the hematological effects. The PBPK model of Corley et
al. (2005a, 1997) was used to obtain estimates of human Cmax values from the female rat drinking
water study data.
The four steps involved in using the Corley et al. (2005a, 1997) PBPK model to calculate
the HED corresponding to the LOAEL identified in the animal study (LOAELHed) were to:
(1) calculate the internal dose surrogate (Cmax BAA in blood) corresponding to the female rat
LOAEL, assuming that the drinking water was consumed only during a 12-hour awake cycle on
a 7 day/week schedule in model simulations; (2) verify that steady state was achieved (e.g., no
change in BAA Cmax as a result of prolonging the exposure regimen); (3) simulate the internal
dose surrogate (Cmax BAA in blood) for humans consuming EGBE in drinking water, assuming
that a 70-kg human consumes an average of 2 L of water during a 12-hour awake cycle; and
(4) calculate the HED (mg/kg-day) for the amount of EGBE consumed in 2 L of water that
resulted in the same internal dose (Cmax BAA) simulated for the animal in step 1 as shown below.
C.2.1.1. NOAEL/LOAEL Method and PBPK Model Applied to Hematologic Data
Step 1: Calculate the Cmax for BAA in blood corresponding to female rat LOAEL.
Female rat LOAEL = 59 mg/kg-day (calculated for final week of 13-week study to
correspond with the final hematological determination) Cmax BAA = 255 |iM,
Step 2: Verify steady state.
There were no changes in the Cmax of BAA in blood during any 24-hour simulation
period using a 12 hour/day, 7 day/week drinking water exposure regimen at the female rat
LOAEL, indicating that steady state was achieved.
Step 3: Calculate the Cmax for BAA in blood for humans continuously exposed to varying
concentrations of EGBE.
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Table C-12 shows modeled estimates of BAA in blood of humans exposed continuously
to varying concentrations of EGBE in water (Corley et al., 1997, 1994). Drinking water volume
is 2 L consumed over 12 hours in a day.
Table C-12. Modeled estimates of BAA in human blood exposed to EGBE in
water.
EGBE concentration (ppm) in
water
Calculated dose of EGBE from drinking water
(mg/kg-d)
Cmax BAA in blood
(jiM)
24
0.7
9
48
1.4
18
94
2.7
36
188
5.4
73
375
10.7
147
750
21.4
299
Step 4: Calculate the LOAEL HED for a 70-kg human consuming EGBE in 2 L of
drinking water/day that results in the same internal dose of EGBE (Cmax of BAA in blood)
calculated for the animal study in step 1.
Female rat Cmax for BAA in blood at LOAEL = 255 [xM
LOAELhed continuous exposure =18 mg/kg-day (calculated by regression of the internal
dose versus the dose of EGBE from step 3).
The internal dose surrogate, Cmax for BAA in blood, is highly dependent on the rate of
water ingestion. Since drinking-water exposures are highly complex and variable, a simplifying
assumption was used in all simulations that the entire dose of EGBE in drinking water was
consumed over a 12-hour period each day corresponding to the awake cycle for both rats and
humans. This assumption resulted in higher Cmax blood concentrations of BAA in both rats and
humans than would have been calculated using the original Corley et al. (1994) structure that
assumed that drinking-water uptake occurred over a 24 hour/day dosing period.
C.2.1.2. BMD Method and PBPK Model Applied to Hematologic Data
Although a lower LOAEL was reported in male rats, this value gives no indication of the
relative slopes of the dose-response curves for males and females. Because this is an important
factor for BMD analyses (U.S. EPA, 2000b, 1995), a comparison of the MCV and RBC count
results for both male and female rats was performed and demonstrated that female rats are more
sensitive to the effects of EGBE than are males. Therefore, dose-response information on these
hematological effects in female rats was selected as the basis for the oral RfD BMD analyses
discussed below.
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As was discussed in Section 5.2.1, MCV and RBC count are continuous response
measurements of precursor events associated with EGBE exposure and are considered the most
appropriate hematologic endpoints for use in a BMD analysis. Cmax is considered the more
appropriate dose metric for use in evaluating the chosen hemolytic endpoints. Cmax for BAA in
arterial blood was determined using the PBPK model of Corley et al. (2005a, 1997). The results
of this modeling effort are summarized in Table C-13.
Table C-13. Model estimates of BAA blood levels in female rats following
oral exposures
EGBE concentration in
water (ppm)
Water EGBE intake
(L/d)
Female body
weight (g)
BAA in blood
Dose (mg/kg-d)
Cmax (jlM)
750
0.0147
188
59
255
1,500
0.0155
185
125
914
3,000
0.0125
180
208
2,370
4,500
0.0101
164
277
3,23 r
6,000
0.0101
150
404
5,464a
aSteady-state not reached within 5 days.
Source: Corley et al. (2005a).
A BMD analysis was performed using U.S. EPABMDS version 1.4.1. As can be seen
from the results in Table C-14, RBC count was the more sensitive of the two hematological
endpoints assessed. All models were fit using restrictions and option settings suggested in the
U.S. EPABMD technical guidance document (U.S. EPA, 2000b) except for the choice of BMR
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Table C-14. Comparison of female rat RBC count and MCV BMD/BMDL
values from an oral subchronic study using modeled blood Cmax (3 months)
of the EGBE metabolite BAA as a common dose metric
Model
BMD0S
(jiM)
BMDL0S
(jiM)
BMDsd
(jiM)
BMDLsd
(jiM)
/7-Value
AICa
Scaled residualb
RBC count
Polynomial (1°)
1,623.31
1,350.41
1,937.36
1,532.88
<0.0001
-27.563915
-0.85
Power
1,623.31
1,350.41
1,937.36
1,532.88
<0.0001
-27.563915
-0.85
Hillc
180.667
93.9053
166.258
84.5854
0.4084
-47.398016
-0.228
MCV
Polynomial (1°)
1,036.38
934.455
578.088
442.513
<0.0001
146.041856
3.37
Power
1,036.38
934.455
578.088
442.513
<0.0001
146.041856
3.37
Hill
475.072
347.356
156.553
100.308
0.001166
135.552419
0.947
aAIC = -2L + 2P, where L is the log-likelihood at the maximum likelihood estimates for the parameters and P is the
number of model degrees of freedom (number of parameters estimated).
b%2 residual (measure of how model predicted response deviates from the actual data) for the dose group closest to
the BMD scaled by an estimate of its SD provides a comparative measure of model fit near the BMD. Residuals
that exceed 2.0 in absolute value should cause questioning of model fit in this region.
°Model choice based on adequate p-valuc (>0.1), visual inspection, low AIC, and low (absolute) scaled residual.
Source: NTP(1993).
For continuous response data, the current BMD technical guidelines (U.S. EPA, 2000b)
suggest the use of 1 SD for the BMR level in the absence of additional information, such as a
minimal level of change in the endpoint that is generally considered to be biologically
significant. Because the chosen hemolytic endpoints are continuous measurements that have a
relatively small historical variance in rats, and because low-dose responses for these endpoints
were generally near or within 5% of the control mean, the BMDL05 was considered to be a more
appropriate POD for derivation of the RfD (U.S. EPA, 2000b, 1995). Higher levels of response
(e.g., >10% reduction) are in the exposure range where other more severe responses related to
anemia occur (e.g., RBC decreases, increased reticulocyte counts). Lower levels of response
(e.g., 1 SD, approximately a 3% reduction for these data) are relatively distant from the
observable data and other responses related to anemia.
Adequate model fit could not be obtained for the MCV data. The best model fit to the
RBC count data (from visual inspection and comparison of AIC values and scaled residuals near
the BMD) was obtained using a Hill model (see Table C-14). A graphical plot and textual
description of the results of the Hill model assessment of RBC count responses in female rats
(NTP, 2000) versus corresponding PBPK estimates of Cmax for BAA in female rat blood are
provided below.
The BMDo5 was 181 [xM and the BMDL0s was determined to be 94 [xM using the 95%
lower confidence limit of the dose-response curve expressed in terms of the Cmax for BAA in
blood. The Corley et al. (2005a, 1997) PBPK model was used to back-calculate a HED
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(BMDLred) of 6.8 mg/kg-day, assuming that rats and humans receive their entire dose of EGBE
from drinking water over a 12-hour period each day.
8
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Hill Model. (Version: 2.14; Date: 06/26/2008)
Input Data File: C:\Usepa\BMDS21Beta\Data\EGBE\femaleratoral\lHilrbchil.(d)
Gnuplot Plotting File: C:\Usepa\BMDS21Beta\Data\EGBE\femaleratoral\lHilrbchil.plt
Tue Mar 03 15:29:39 2009
BMD Method for RfD: RBC Response in Orally Exposed Female Rats (NTP, 1993)
The form of the response function is:
Y[dose] = intercept + v*doseAn/(kAn + doseAn)
Dependent variable = RBC
Independent variable = Cmax
rho is set to 0
Power parameter restricted to be greater than 1
ElMDL
BMD
Hill Model with 0.95 Confidence Level
0 1000 2000 3000 4000 5000
dose
2009
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A constant variance model is fit
Total number of dose groups = 6
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
alpha = 0.154717
rho = 0 Specified
intercept = 8.15
v= -1.57
n = 0.982521
k = 551.55
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -rho -n
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
alpha intercept v k
alpha 1 2.2e-010 1.9e-009 6.6e-009
intercept 2.2e-010 1 -0.55 -0.54
v 1.9e-009 -0.55 1 -0.29
k 6.6e-009 -0.54 -0.29 1
Parameter Estimates
95.0% Wald Confidence Interval
Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. Limit
alpha
intercept
v
n
k
0.146124
8.13659
-1.57421
1
518.419
0.0266785
0.120911
0.156637
NA
227.998
0.0938351
7.89961
-1.88122
71.5517
0.198413
8.37357
-1.26721
965.286
NA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
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has no standard error.
Table of Data and Estimated Values of Interest
Dose N Obs Mean Est Mean Obs Std Dev Est Std Dev Scaled Res.
0
10
8.15
8.14
0.28
0.382
0.111
255
10
7.59
7.62
0.48
0.382
-0.228
914
10
7.09
7.13
0.43
0.382
-0.348
2370
10
7
6.84
0.37
0.382
1.28
3231
10
6.8
6.78
0.36
0.382
0.165
5464
10
6.58
6.7
0.41
0.382
-0.983
Model Descriptions for likelihoods calculated
Model A1: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)A2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A3 uses any fixed variance parameters that
were specified by the user
Model R: Yi = Mu + e(i)
Var{e(i)} = SigmaA2
Likelihoods of Interest
Model
Log(likelihood)
# Param's AIC
A1
29.145609
7
-44.291218
A2
30.744881
12
-37.489762
A3
29.145609
7
-44.291218
fitted
27.699008
4
-47.398016
R
-3.574142
2
11.148285
Explanation of Tests
Test 1: Do responses and/or variances differ among Dose levels?
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(A2 vs. R)
Test 2: Are Variances Homogeneous? (A1 vs A2)
Test 3: Are variances adequately modeled? (A2 vs. A3)
Test 4: Does the Model for the Mean Fit? (A3 vs. fitted)
(Note: When rho=0 the results of Test 3 and Test 2 will be the same.)
The p-value for Test 1 is less than .05. There appears to be a
difference between response and/or variances among the dose levels
It seems appropriate to model the data
The p-value for Test 2 is greater than . 1. A homogeneous variance
model appears to be appropriate here
The p-value for Test 3 is greater than . 1. The modeled variance appears
to be appropriate here
The p-value for Test 4 is greater than . 1. The model chosen seems
to adequately describe the data
Benchmark Dose Computation
Specified effect = 0.05
Risk Type = Relative risk
Confidence level = 0.95
Tests of Interest
Test -2*log(Likelihood Ratio) Test df p-value
Test 1
Test 2
Test 3
Test 4
68.638 10 <0001
3.19854 5 0.6694
3.19854 5 0.6694
2.8932 3 0.4084
BMD
180.667
BMDL = 93.9053
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