DRAFT - DO NOT CITE OR QUOTE
oEPA
NCEA-S-3010
www.epa.gov/iris
TOXICOLOGIC AL REVIEW
OF
ETHYLENE GLYCOL
MONOBUTYL ETHER (EGBE)
NOTICE (CAS No. 111-76-2)
In Support of Summary Information on the
Integrated Risk Information System (IRIS)
APRIL 2008
This document is an External Review draft. This information is distributed solely for
the purpose of pre-dissemination peer review under applicable information quality guidelines. It
has not been formally disseminated by EPA. It does not represent and should not be construed to
represent any Agency determination or policy. It is being circulated for review of its technical
accuracy and science policy implications.
U.S. Environmental Protection Agency
Washington, DC
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Disclaimer
This document is a preliminary draft for review purposes only. 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. Mention of trade
names or commercial products does not constitute endorsement or recommendation for use.
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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 of Acronyms ix
Foreword xi
Authors, Contributors, and Reviewers xii
Chemical Managers/Authors xii
Internal EPA 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 12
4.1. Studies in Humans: Epidemiology, Case Reports, Clinical Controls 12
4.2. Subchronic and Chronic Studies and Cancer Bioassays in Animals: Oral and Inhalation 15
4.2.1. Subchronic Studies 15
4.2.1.1. Oral 15
4.2.1.2. Inhalation 20
4.2.2. Chronic Studies and Cancer Bioassays 25
4.2.2.1. Inhalation 25
4.3. Reproductive and Developmental Studies: Oral and Inhalation 29
4.4. Other Studies 34
4.4.1. Acute and Short-Term Exposure Studies 34
4.4.2. Dermal Exposure Studies 38
4.4.3. Ocular Exposure Studies 39
4.4.4. Genotoxicity 39
4.4.5. Immunotoxicity 42
4.4.6. Other In Vitro Studies 44
4.5. Synthesis and Evaluation of Major Noncancer Effects and Mode of Action: Oral and Inhalation 46
4.6. Evaluation of Carcinogenicity 51
4.6.1. Summary of Overall Weight of Evidence 51
4.6.2. Synthesis of Human, Animal, and Other Supporting Evidence 51
4.6.3. Mode of Action Information 53
4.6.3.1. Hypothesized Mode of Action for Liver Tumor Development in Male Mice 53
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4.6.3.1.1. Temporal association and species specificity 56
4.6.3.1.2. Dose-response relationships 56
4.6.3.1.3. Biological plausibility and coherence of the database 57
4.6.3.1.4. Relevance of the hypothesized MOA to humans 58
4.6.3.1.5. Other possible MOAs for liver tumor development in male mice 59
4.6.3.2. Hypothesized Mode of Action for Forestomach Tumor Development in Female Mice 62
4.6.3.2.1. Temporal association 64
4.6.3.2.2. Dose-response relationships 64
4.6.3.2.3. Biological plausibility and coherence of the database 65
4.6.3.2.4. Relevance of the hypothesized MOA to humans 66
4.6.3.2.5. Other possible MOAs for forestomach tumor development in female mice 66
4.6.3.3. Conclusions About the Hypothesized Modes of Action 67
4.7. Susceptible Populations 68
4.7.1. Possible Childhood Susceptibility 70
4.7.2. Possible Gender Differences 71
5. Dose-Response Assessments 73
5.1. Inhalation Reference Concentration (RfC) 73
5.1.1. Choice of Principal Study and Critical Effect, with Rationale and Justification 73
5.1.2. Methods of Analysis, Including Models (PBPK, BMD, etc.) 77
5.1.2.1. Derivation of the POD Using PBPK Modeling and the NOAEL/LOAEL Method 77
5.1.2.2. Derivation of the POD Using PBPK Modeling and BMD Modeling Methods 80
5.1.2.2.1. BMD approach applied to hematological data 80
5.1.2.2.2. BMD approach applied to hemosiderin staining data 82
5.1.2.3. Selection of the POD 85
5.1.3. RfC Derivation, Including Application of Uncertainty Factors (UFs) 86
5.1.4. RfC Comparison Information 87
5.1.5. Previous Inhalation Assessment 89
5.2. Oral Reference Dose (RfD) 89
5.2.1. Choice of Principal Study and Critical Effect, with Rationale and Justification 89
5.2.2. Methods of Analysis, Including Models (PBPK, BMD, etc.) 92
5.2.2.1. Derivation of POD Using PBPK Model and NOAEL/LOAEL Method 92
5.2.2.2. Derivation of POD Using PBPK and BMD Methods 94
5.2.2.2.1. BMD approach applied to hematological data 94
5.2.2.2.2. BMD approach applied to hemosiderin endpoint 96
5.2.2.3. Route-to-Route Extrapolation from Inhalation Data 96
5.2.2.4. Selection of the POD 96
5.2.3. RfD Derivation, Including Application of Uncertainty Factors (UFs) 97
5.2.4. Previous Oral Assessment 98
5.3. Uncertainties in the Derivation of the Inhalation Reference Concentration (RfC) and Oral
Reference Dose (RfD) 99
5.3.1. Choice of endpoint 99
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5.3.2. Choice of dose metric 99
5.3.3. Use of BMC approach 100
5.3.4. Choice of model for BMCL derivations 100
5.3.5. Choice of animal to human extrapolation method 100
5.3.6. Route-to-route extrapolation 100
5.3.7. Statistical uncertainty at the POD 101
5.3.8. Choice ofbioassay 101
5.3.9. Choice of species/gender 101
5.3.10. Human relevance of noncancer responses observed in mice 101
5.3.11. Human population variability 102
5.4. Cancer Assessment 102
5.4.1. Uncertainties in Cancer Risk Assessment 104
5.1.1.1. Choice of low-dose extrapolation method 106
5.4.1.2. Human relevance of cancer responses observed in mice 106
5.5. Potential Impact of Select Uncertainties on the RfC 107
6. Major Conclusions in the Characterization of Hazard and Dose Response 109
6.1. Human Hazard Potential 109
6.2. Dose Response Ill
Appendix A
Appendix B
Appendix C
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List of Tables
1 Table 4-1. Hematology and hemosiderin data from the 13-week drinking-water exposure to
2 EGBE in F344 rats 18
3 Table 4-2. Incidence* and severity of selected histopathological lesions from the 13-week
4 drinking water exposure to EGBE in F344 rats 19
5 Table 4-3. Hematology and hemosiderin data from a 14-week inhalation study of EGBE in
6 F344 rats 23
7 Table 4-4. Hematology and hemosiderin data from a 14-week inhalation study of EGBE in
8 B6C3F1 mice 24
9 Table 4-5. Selected Female and Male Rat and Mouse Non-neoplastic effects from the 2-year
10 Chronic EGBE Inhalation Study 26
11 Table 4-6. Comparison of Female and Male Rat and Mouse Hct (Manual) Values from 3-
12 and 12-month Inhalation Exposures to EGBE 27
13 Table 4-7. Summary of genotoxicity studies on EGBE, BAL, and BAA 41
14 Table 4-8. Incidence of liver hemangiosarcomas and hepatocellular carcinomas in studies of
15 NTP chemicals that caused increased hemosiderin in Kupffer cells in male mice 57
16 Table 5-1. Results of candidate studies 74
17 Table 5-2. Female and male rat and mouse liver hemosiderin staining incidence and RBC
18 from subchronic and chronic EGBE inhalation studies 75
19 Table 5-3. Summary of PBPK models 79
20 Table 5-4. Model estimates of BAA blood levels in female rats following inhalation
21 exposures 81
22 Table 5-5. Comparison of BMC/BMCL values for female rat RBC count data from a 14-
23 week subchronic inhalation study, using modeled blood Cmax (3 months) of the
24 EGBE metabolite BAA as a common dose metric 81
25 Table 5-6. AUC BAA blood concentrations measured at 12 months in both sexesgenders of
26 B6C3F1 mice and F344 rats 82
27 Table 5-7. Comparison of BMC/BMCL values for male and female rat liver hemosiderin
28 staining data from inhalation chronic study using measured blood AUC (12
29 months) of the EGBE metabolite BAA as a common dose metric 84
30 Table 5-8. Comparison of BMC/BMCL values for male and female mouse liver
31 hemosiderin staining data from inhalation chronic study using measured blood
32 AUC (12 months) of the EGBE metabolite BAA as a common dose metric 85
33 Table 5-9. Summary of the application of UFs for RfC derivations using NOAEL/LOAEL
34 and BMC modeling approaches for male rat liver hemosiderin staining 86
35 Table 5-10. Subchronic 91-day drinking water studies in rats and mice 90
36 Table 5-11. Modeled estimates of BAA in human blood exposed to EGBE in water 93
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1 Table 5-12. Model estimates of BAA blood levels in female rats following oral exposures 94
2 Table 5-13. Comparison of female rat RBC count and MCV BMD/BMDL values from an
3 oral subchronic study using modeled blood Cmax (3 months) of the EGBE
4 metabolite BAA as a common dose metric 95
5 Table 5-14. Summary of the application of UFs for RfD derivations using NOAEL/LOAEL
6 and BMD modeling approaches 97
7 Table 5-15. Summary of uncertainty in the EGBE noncancer and cancer risk assessments 105
8 Table 5-16. Illustrative potency estimates for tumors in mice, using a linear analysis
9 approach 106
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List of Figures
1 Figure 3-1. Proposed metabolic scheme of EGBE in rats and humans 6
2 Figure 4-3. Simulated concentrations of EGBE, BAL, and BAA in liver tissues of female
3 mice exposed via inhalation for 6 hours to 250 ppm EGBE 61
4 Figure 5-5. Points of departure for selected endpoints with corresponding applied UFs and
5 derived RfC 88
6 Figure 5-7. Potential impact of select uncertainties on the RfC for EGBE 108
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List of Acronyms
ADH alcohol dehydrogenase
AIC Akaike Information Criterion
ALD aldehyde dehydrogenase
AUC area under the curve
BAA 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
CHR contact hypersensitivity response
CI confidence interval
Cls clearance rate
con-A concanavalin-A
Cmax peak concentration
CYP450 cytochrome P450
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
EPA U.S. Environmental Protection Agency
GD gestation day
GFR glomerular filtration rate
G6PD glucose-6-phosphate dehydrogenase
GSH glutathione
Hb hemoglobin
Hct hematocrit
HEC human equivalent concentration
HED human equivalent dose
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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
MOA
mode of action
NK
natural killer
NOAEL
no-ob served-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 V2
half-life
TNFa
tumor necrosis factor-alpha
TNP-LPS
trinitrophenyl-lipopolysaccharide
UF
uncertainty factor
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).
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Authors, Contributors, and Reviewers
Chemical Managers/Authors
Angela Howard, Ph.D.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Research Triangle Park, NC
Jeffrey Gift, Ph.D.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Research Triangle Park, NC
Internal EPA Reviewers
IIa Cote, Ph.D., DABT
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Research Triangle Park, NC
Jennifer Jinot, Ph.D.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Jane Caldwell, Ph.D.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
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
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
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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 (MO A). The RfD, expressed in units of mg/kg-day, is defined as an estimate
(with uncertainty spanning perhaps an order of magnitude) of 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 effects
peripheral to the respiratory system (extrarespiratory or systemic effects). Reference values are
generally derived for chronic exposure (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 values are derived based on the 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 an upper bound on the
estimate of risk per mg/kg-day of oral exposure. Similarly, an 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
(1983). 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
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Assessment (U.S. EPA, 1991), Interim 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 August 2007.
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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, and 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 (National Toxicology Program [NTP], 2000).
Some relevant physical and chemical properties of EGBE are:
CASRN: 111-76-2
Empirical formula: C4H9-O-CH2CH2-OH
Molecular weight: 118.2
Vapor pressure: 0.88 mm Hg at 25°C (about 1200 ppm)
Water solubility: Miscible
LogKow: 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, it is anticipated that
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
temperatures and humidities (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
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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 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; European Centre for Ecotoxicology and Toxicology of Chemicals, 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 butoxyacetaldehyde
(BAL) and its corresponding carboxylic acid, BAA, and ethylene glycol (EG).
The two main oxidative pathways of EGBE metabolism observed in rats are via alcohol
dehydrogenase (ADH) and O-dealkylation by a cytochrome P450 (CYP450) 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, 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.
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CH3CH2CH2CH2OH
(Butanol)
CO,
^ earboligase oxidase dehydrogenase
hoch2 CH2OH
(Ethylene Glycol)
CH3 CH2 CH2 CH2 OCH2 CH2 O-Giuc
(EGBE - Giueuronicfe)
{Rats Only?)
(Rats Only?)
deatkylase
CHs C H ;• C H 2 CHj OCH2 CH, O-SOs H
(EGBE - Sulfate)
(Rats Only?)
CH3CH2CH2 CH2 OCH2 CH? OH
(EGBE)
(Rats and Human)
1
alcohol dehydrogenase
CH3 CH2 CH2 CH2 OCH2 CHO
(BAD
CHj CHj CHj CH;.OCH2 CCh Giu
(BAA - Glutamine)
(Human Only)
aldehyde dehydrogenase
CH.,CH2CH2CH2 0CH2C02-Gty
(BAA - Glycine)
(Human Only)
CH,CHjCH* CH2OCH2COa H
(BAA)
I
CO2
dealkyl carboligase
Figure 3-1. Proposed metabolic scheme of EGBE in rats and humans.
Source: Adapted from Medinsky et al. (1990) and Corley et al. (1997).
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).
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Percutaneous absorption of EGBE in rats is rapid and produces measured blood levels of
BAA sufficient to produce hemolysis (Bartnik et al., 1987). Metabolism, disposition, and
pharmacokinetic studies in male F344 rats conducted by Corley et al. (1994) produced hemolytic
blood concentrations of BAA (0.5 mM) following a single oral dose of 126 mg/kg of 14C-labeled
EGBE. Using their physiologically based pharmacokinetic (PBPK) model, they predicted that
such hemolytic blood concentrations would also be produced in rats following a single 6-hour
EGBE inhalation exposure greater than 200 ppm. A report that evaluated the NTP (2000)
inhalation bioassay suggests that BAA blood concentrations in rats exceeded 0.5 mM
(approximately 67 |ig 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
(ALD), respectively (Ghanayem et al., 1987b). Male F344 rats, 9 to 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.
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 ALD, 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 B6C3F1 mice (n = 30) and rats (n = 10; gender
and species not specified) and centrifuged at 41,000 x g3 with the supernatants used to examine
the metabolism of EGBE by ADH and ALD. The stomachs were separated into fore and
glandular sections and used to measure the metabolism of EGBE to BAL and BAL to BAA by
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ADH and ALD, respectively. A marked species difference in ALD 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 up to one order of magnitude greater in mice compared to rats. Based
upon the Km and Vmax values reported, while the mouse ALD 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 ALD 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 ALD 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 ALD 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.
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 months 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
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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 BAA.
The increased Cmax, AUC, and tu 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.
For humans, the elimination kinetics of EGBE and BAA appear to be independent of the
route of exposure. The ty2 in humans for the elimination of EGBE and BAA averaged 0.66 and
3.27 hours, respectively. 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 hour.
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). Post-shift 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 from exposure from five healthy, male research subjects exposed to 20 ppm EGBE via
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 |iM 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 hour, 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 Yd averaged 15 L (range 6.5-25 L) based on whole
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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 to 16 hours post-shift. Laitinen et al. (1998) reported
similar post-shift 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.
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 B6C3F1 mice (n = 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 post-dosing, but at levels that were
roughly 50-fold lower than in the forestomach. BAA concentrations in all organs were
substantial in the 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 post-dose, 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.
Dill et al. (1998) reported on toxicokinetic findings collected from exposures carried out
in male and female F344 rats and male and female B6C3F1 mice as part of the 2-year EGBE
inhalation toxicity and carcinogenicity study conducted by the NTP (2000). Blood samples were
collected from some of these animals post- exposure (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 and
31.2 ppm (rats only) and 62.5, 125, or 250 ppm (mice only) by whole-body inhalation and
assayed for EGBE. Post-exposure 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. Post-
exposure 16-hour urine samples were collected after 2 weeks and 3,6, 12, and 18 months of
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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. Post-
exposure 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 was less than
10 minutes, and ty2 for mice was less than 5 minutes. 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 post-exposure, ty2 in male rats was 9.4 minutes, and at 18 months
post-exposure 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 appears to be dependent on various factors, including species, gender, age,
time of exposure, and exposure concentration.
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 concentrations of 7 |iM EGBE, 0.5 |iM
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 ALD, and the subsequent metabolism
of BAL to BAA via ADH 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 |iM, respectively, following
oral 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 post-dosing, 19 and 33 |iM,
respectively, following oral gavage exposure to EGBE at 600 mg/kg (Deisinger and Boatman,
2004).
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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
thrombocytopenia. 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.
Gualideri 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
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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
discomfort within 3 hours post-ingestion. Approximately 10 hours post-admission, 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 post-ingestion. The patient was transferred to a
tertiary care hospital where hemodialysis was initiated at approximately 24 hours post-ingestion.
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, a 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 manifested. 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 post-ingestion. No evidence of hemolysis or renal abnormalities
was detected.
A 50-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 post-exposure, 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 tenth 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%>).
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Metabolic acidosis was manifest, and a single dose (15 mg/kg) of the ALD inhibitor 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 one 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 to 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: 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
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determinations of BAA in post-shift urine samples were used to assess exposure to low levels of
EGBE. No differences were observed for RBC counts, Hb, MCV, MCH, Hp, reticulocyte count,
and 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 adverse 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 significantly increased to a mean value of 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 oral gavage route of exposure have been
conducted. Krasavage (1986) conducted a toxicity study using groups of 10 COBS CD (SD)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
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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,
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 1430 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 B6C3F1 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 targets. 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 B6C3F1 mice, where
groups of 10 animals/gender/species received EGBE in drinking water at doses of 0, 750, 1500,
3000, 4500, and 6000 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-
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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 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-L 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 clear 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 and reduced Hb concentration, reduced platelets, and increased
bone marrow cellularity at > 367 mg/kg-day. 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.
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 lesions, 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; as the doses increased with the severity in the two highest dose groups
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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
(69 mg/kg-day)
1,500 ppm
(129 mg/kg-day)
3,000 ppm
(281 mg/kg-day)
4,500 ppm
(367 mg/kg-day)
6,000 ppm
(452 mg/kg-day)
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 ± 0.6** (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 ± 0.2" (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 (106/|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 ±0.12** (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 C106/j.iL)
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 (103/|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 ±0.03**
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 ±0.4*" (104)
52.3 ±0.4 (100)
60.5 ±0.4*** (110)
54.4 ±0.3*** (105)
62.4 ±0.6*** (114)
56.7 ±0.5*** (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 ±0.2**
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, p £ 0.01.
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Table 4-2. Incidence* and severity of selected histopathological lesions from the 13-week drinking water exposure to EGBE in F344
rats.
750 ppm
1500 ppm
3000 ppm
4500 ppm
6000 ppm
Control
(69 mg/kg-d)
(129 mg/kg-d)
(281 mg/kg-d)
(367 mg/kg-d)
(452 mg/kg-d)
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
Relative kidney weight (right) (mg organ wt/g body wt)
Females (only)
6.33 ±0.10
7.69 ±0.14*"
8.06 ±0.29*"
7.47 ±0.19***
7.55 ±0.18***
8.21± 0.26***
Necropsy body weight (g) §
Females (only)
31.1 ± 0.7
31.8 ± 0.8
30.9 ± 1.5
28.0 ±0.7**
28.4 ±0.5**
27.8 ±0.9***
* 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.
** p < 0.05
*** p < 0.01
NR = Statistics not reported
§ mean ± standard error
Source: NTP (1993)
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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 in both 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.
Female mice showed statistically significant reductions in body weight gain starting at
3000 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 1500 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 significantly increase at 750 and 1500 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 increases in tumors at any site.
4.2.1.2. Inhalation
Wistar-derived rats (23 animals/group, gender not specified) were exposed to 0, 135, or
320 ppm EGBE for 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 the 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 for 7 hours/day, 5 days/week for 12 weeks. Necropsies were performed
5 weeks post-exposure; 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;
they were reversed 5 weeks after the end of exposure.
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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 or higher. 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. No NOAELs were reported. The authors reported 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 radioiodine 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.
A 90-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 B6C3F1
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) for 6 hours/day, 5 days/week for 14 weeks.
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Hematologic and hemosiderin staining results are presented in Table 4-3 and Table 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 clear 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 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 in excess of
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.
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
extramedullary 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
greater than 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 in mice was 62.5 ppm, based on
histopathological changes in the forestomach.
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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 (10'/|iL)
Males
Females
9.05 ±0.08
8.48 ±0.05
8.71 ±0.14** (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 (10'/|iL)
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 ±0.04**
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 ±0.03**
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 ±0.2*** (105)
58.7 ±0.2*** (107)
53.8 ±0.3*** (107)
61.6 ±0.2*** (112)
58.5 ± 0.3***(117)
66.8 ±0.9*** (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, p < 0.01.
Source: NTP (1993).
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Table 4-4. Hematology and hemosiderin data from a 14-week inhalation study of EGBE in B6C3F1 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 ±0.1" (98)
15.9 ±0.1 (101)
15.4 ±0.1" (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 ±0.06" (98)
9.77 ±0.1 (101)
9.51 ±0.06" (98)
9.47 ± 0.06** (98)
9.18 ±0.05*** (94)
8.90 ± 0.07*** (92)
8.57 ± 0.06*** (88)
7.21 ±0.23*** (74)
7.35 ±0.07*** (76)
Reticulocytes (10'/|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 ±0.03**
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
Values listed are mean ± standard error (percent of control).
"Statistically significant difference,/) < 0.05.
"Statistically significant difference, p < 0.01.
Source: NTP (1993)
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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 B6C3F1 mice. The researchers 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). 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 (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 for 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. Non-neoplastic 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).
Non-neoplastic, statistically significant effects in mice included forestomach ulcers and
epithelial hyperplasia, hematopoietic cell proliferation and hemosiderin pigmentation in the
spleen, Kupffer cell pigmentation in the livers, and bone marrow hyperplasia (males only).
Hyaline degeneration of the olfactory epithelium (females only) was increased relative to
chamber controls, but was not statistically significant. As in the rats, the Kupffer cell
pigmentation was considered a secondary effect of the hemolytic activity of EGBE. Bone
marrow hyperplasia, hematopoietic cell proliferation and hemosiderin pigmentation in the 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.
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Table 4-5. Selected Female and Male Rat and Mouse Non-neoplastic 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
34/50*
36/50*
42/50*
47/50*
NT
NT
Hyaline degeneration of the olfactory epithelium
Male
Female
13/48
13/50
21/49*
18/48
23/49*
28/50*
40/50*
40/49*
NT
NT
Mouse
Kupffer cell pigmentation, hemosiderin in the liver
Male
Female
0/50
0/50
NT
NT
0/50
5/50*
8/49**
25/49**
30/49**
44/50**
Hematopoietic cell proliferation in the spleen
Male
Female
12/50
24/50
NT
NT
11/50
29/50
26/48**
32/49
42/50**
35/50*
Hemosiderin in the spleen
Male
Female
0/50
39/50
NT
NT
6/50*
44/50
45/48**
46/49**
44/49**
48/50**
Forestomach ulcers
Male
Female
1/50
1/50
NT
NT
2/50
7/50*
9/49**
13/49**
3/48
22/50**
Forestomach epithelial hyperplasia
Male
Female
1/50
6/50
NT
NT
7/50*
27/50**
16/49**
42/49**
21/48**
44/50**
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
9/49*
5/50*
'Statistically significant difference, p < 0.05.
"Statistically significant difference,/) <0.01.
NT = not tested
Source: NTP (2000)
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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 each male and 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-
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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, anda 12 months in the 125 ppm female mice
and the 250 pm male and female 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 (decreased 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.
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 rats*
3 months
12 months
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 rats*
3 months
12 months
44.9 ±0.2
47.8 ±0.4
46.9 ±0.5 (104)
44.8 ±0.4 (100)
45.9 ±0.8** (96)
42.9 ±0.5** (95)
42.9 ± 1.2*** (90)
-
Female mice*
3 months
12 months
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 mice*
3 months
12 months
47.5 ±0.3
47.9 ±0.4
-
47.3 ±0.5 (100)
48.7 ± 1.9** (102)
46.0 ± 0.4** (97)
46.4 ± 1.0(97)
43.7 ±0.2*** (92)
42.1 ±0.4*** (88)
'These 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).
- data were not available.
"Statistically significant difference,/) < 0.05.
"Statistically significant difference, p < 0.01.
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)
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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 the
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 hematocrit values in male and female rats
and mice at 3 months and 12 months. Statistically significant (p < 0.05) decreases in automated
and 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 the
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), 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%) does 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). 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 <
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0.05) in relation to the control chamber group (16/50, 8/50, 7/49, 8/49). However, in light of the
high survival rate of the exposed female mice relative to the control animals (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 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.
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 and 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 adverse 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 (>1000 mg/kg). Maternal toxicity, related to the hematologic effects
of EGBE, and relatively minor developmental effects, have been reported in developmental
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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(SD)BR adult
male rats treated by gavage with 222, 443, or 885 mg/kg-day undiluted EGBE 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- or 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 adverse effect on fertility was observed in the 700 mg/kg-day
dose group.
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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 (Fi) pups. There were insufficient numbers of offspring to assess the two highest dose
groups, and no adverse 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 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 adverse 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 B6C3F1 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 1364 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
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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 B6C3F1 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 B6C3F1 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).
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
1,000 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
>1500 mg/kg-day). Based on clinical signs in the prenatal study, the LOAEL for maternal effects
was 350 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 650 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 adverse reproductive or 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 are 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
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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 aNOAEL 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 adverse effects were noted
thereafter, not in offspring. The LOAEL was 200 ppm for slight maternal toxicity; a NOAEL was
identified at 150 ppm. The NOAEL for developmental toxicity was 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 days 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 (N not
specified) via dermal administration during GDs 6-15, four times per day at 1800 and
5400 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, gross malformations or variations,.
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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) 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; 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 also were 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 by older rats to 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
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to normal by posttreatment day 22, except for liver and spleen weights that were at 1,000 mg/kg-
day, somewhat increased levels (~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
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 clearly
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.
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Corley et al. (1999) conducted a series of studies in B6C3F1 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
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 B6C3F1 mice (16/gender/group) were
exposed by oral gavage to 0, 400, 800, or 1200 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
the lower-dose groups, as well as in both genders. 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
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consumption via grooming were measured. An average of 205-69 |ig of EGBE was detected on
the fur of the mice exposed whole-body, while an average of 170-52 |ig 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 B6C3F1 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 B6C3F1 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 B6C3F1 mice (n = 12) were given a single i.v. injection of 10 mg/kg 2-
butoxy[l-14C]ethanol (850 |iCi/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.
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 male and five females.
After excessive mortality (3/5 males, 5/5 females) was observed at this dose, reduced doses of
500 and 1000 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] = 1020-1961 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.
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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, 1993a,
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) of 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 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
(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.
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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 post-challenge. 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 is 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 |iL, 99% pure) on eye irritation. The undiluted chemical was dropped onto the
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 post-instillation. 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 (amol/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
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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.
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 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
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 sister chromatid exchanges (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 ALD 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.
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Table 4-7. Summary of genotoxicity studies on EGBE, BAL, and BAA.
Type of test, test species
Dose*
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)
Zeigeretal. (1992)
(work done for NTP)
Reverse mutation, S. typhimurium his-TA
98, TA 100, TA 102
115 (.Linol/plate
(14.0 mg/plate)
Negative (w/and w/o metabolic
activation)
Hoflack et al. (1995)
Reverse mutation, S. typhimurium his-
TA97a
38 (.Linol/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 (MNs) 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 et al. (1996)"
Potentiation of clastogenicity induced by
methyl methanesulfonate
8.5 mM
Positive (w/o metabolic
activation)
Elias et al. (1996)"
Chromosomal aberrations, V79 cells and
human lymphocytes
Not available
Negative (w/o metabolic
activation)
Elias et al. (1996)"
Gene mutation, Chinese hamster ovary
cells
38.1 mM*"
(4.5 mg/mL)
Negative (w/o metabolic
activation)
Chiewchanwit and Au
(1995)
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 |-Lmol/platc
(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)**
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)**
Aneuploidy, V79 cells
0.38 mM
Weakly positive (w/o
metabolic activation)
Elias et al. (1996)**
MN assay, V79 cells
10 mM
Positive (w/o metabolic
activation
Elias et al. (1996)**
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-day, mice
450 mg/kg-day, rats
Negative
Negative
NTP (1996)
DNA adducts FVB/N mice
Sprague-Dawley rats
120 mg/kg-day; mice
and rats
No changes in DNA
methylation
Keith et al. (1996)
Doses are either the lowest effective dose or the highest ineffective dose.
"All in vitro assays were performed without the addition of an exogenous metabolic activation system.
"The authors found that this dose was cytotoxic.
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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 (amol/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
(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,
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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 |ig/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-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.
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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 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
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, 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.
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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-
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 pre-hemolytic
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
(1995, 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 post-splenectomy and 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 2.0 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
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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 that were tested (0.5 |iM and 1.0
|iM), The next experiment examined the ability of EGBE, BAA, ferrous sulfate, and hemolyzed
RBCs to stimulate tumor necrosis factor-alpha (TNFa) 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.
4.5. Synthesis and Evaluation of Major Noncancer Effects and Mode
of Action: Oral and Inhalation
Intravascular hemolysis 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. Hemolysis can be induced in
vivo following administration of EGBE or in vitro following addition of BAA to either whole
blood or washed erythrocytes. In vitro tests have shown that BAA produces a concentration- and
time-dependent swelling of rat erythrocytes, and changes in the normal erythrocyte morphology
from the typical discocyte form to a spherocyte form prior to lysis. This response appears to be
mediated by the erythrocyte membrane and results in an increase in osmotic fragility and a loss
of deformability of the erythrocyte, thereby leading to hemolysis. Older erythrocytes are
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apparently more sensitive to the hemolytic effects of BAA than are younger cells or newly
formed reticulocytes.
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
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. 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
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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
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 clear dose-response
relationship as well as to increase in severity in the chronic rat and mouse NTP studies; it shows
a statistically significantly 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, converselywhich
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 been 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 400-1500 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
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explain the slight increase in incidence and severity of the anemic response found in female, as
compared to male, mice. 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.
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 with 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 by 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. Apparently, there is 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,
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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 are
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.
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 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 expected environmental concentrations,
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 at levels at or below the RfC and RfD values established in this assessment. Carpenter
et al. (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 3000 to
3500 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. Non-neoplastic effects
in rats included hyaline degeneration of the olfactory epithelium and Kupffer cell pigmentation.
Non-neoplastic 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 cytogenetic 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 MOAfor 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.
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 B6C3F1 mice. Rats (50/gender/group) were exposed to concentrations of 0, 31, 62.5, and
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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 pm (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 for 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
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
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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
pheochromocytoma of the adrenal medulla. The researchers reported some evidence of
carcinogenic activity in male B6C3F1 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 B6C3F1 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 non-neoplastic 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 non-
neoplastic 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 Mode of Action 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
high dose group of these two types of tumors were only slightly higher than the upper end of the
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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 contribute to the transformation of male mouse endothelial
cells to hemangiosarcomas and hepatocytes to hepatocellular carcinomas. 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 and hemolysis and an increase in Hb levels.
(3) Excess Hb from damaged RBCs is taken up by phagocytic (Kupffer) cells of the
spleen and 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
(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.
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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(7) ROS stimulates hepatocyte and endothelial cell proliferation.
(8) ROS promotes initiation of hepatocyte and endothelial cells.
(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 clearly 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 (Siesky et al., 2002; Park 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 B6C3F1 mice from NTP bioassays and concluded that a significantly
increased risk of inducing hepatic hemangiosarcomas in male B6C3F1 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
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 supports to the proposed steps of initiation and promotion of
neoplasms by ROS (Klaunig et al., 1998).
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4.6.3.1.1. 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.
4.6.3.1.2. Dose-response relationships
Six chemicals have been determined by the NTP to cause hemosiderin buildup in the
livers of mice. As shown in Table 4-8, male mice exposed to chemicals that caused a significant
hemosiderin buildup in Kupffer cells 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, but
liver tumors were not increased in rats at any dose. However, the high dose used in the rat study
was only half that of the high dose used in the mouse study, leaving the possibility that similar
responses could have been observed in rats if higher doses of EGBE were tolerated by this
species.
Of these six chemicals, liver hemangiosarcomas were observed with the only four
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 (U.S. EPA,
2005c; Gift, 2005). Two of the four 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
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
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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
SC*
Hemangiosarcoma
Hepatocarcinoma
Type*"
2-Butoxyethanol
(EGBE) (TR-484)
0/50, 0/50, 8/49°,
30/49°
yes
0/50, 1/50, 2/49, 4/49§
10/50, 11/50, 16/49,
21/49°
I
p-Chloroaniline
hydrochloride (TR-351)
0/50, 0/49, 0/50, 50/50±
yes
2/50, 2/49, 1/50, 6/50
3/50, 7/49, 11/50®,
17/50°
G
p- Nitroaniline
(TR-418)
1/50, 1/50, 8/50§,
50/50°
yes
0/50, 1/50, 2/50, 4/50
10/50, 12/50, 13/50, 6/50
G
Pentachloroanisole
(TR-414)
1/50, 50/50°, 50/50**'°
yes
2/50, 8/50, 10/50®
9/50, 16/50, 12/50
G
C.I. Pigment Red 3
(TR-407)±
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
* Chemicals that caused hemosiderin accumulation in Kupffer cells following subchronic (SC) exposure are
identified with a "yes" in this column.
** The authors of this study could not identify the source of the Kupffer cell pigmentation. They speculate that it may
consist of porphyrins known to be produced from exposure to chlorinated hydrocarbons, but the possibility that it
may have consisted of hemosiderin could not be discounted.
*** I = inhalation, G = gavage, F = feed
§p<0.05
°p<0.01
Statistics not reported
4.6.3.1.3. 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
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 Kamendulis et al. (1999) and Siesky et al. (2002), 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 are 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.
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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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 ALD 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 BAAt/2 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.
4.6.3.1.4. 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
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 (400-1500 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
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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, as opposed to 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, the highest tested in
the study, 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 would not be at
increased risk of tumor development through this MO A.
4.6.3.1.5. 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
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 A) to include the metabolism of EGBE to BAL
via ALD and the subsequent metabolism of BAL to BAA via ADH in both the liver and
forestomach. As shown in Figure 4-2, 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
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1 [2000] study) would result in peak Cmax concentrations of 7 |iM EGBE, 0.5 |iM BAL, and 3,250
2 |iM BAA in liver tissue of male mice at the end of a 6-hour exposure period.
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4.6.3.2. Hypothesized Mode of Action 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/BAA in 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
(5) Continued injury by BAA and degeneration leading to high cell proliferation and
turnover
(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 (Poet et al., 2002; Green et al., 2002), nose-only inhalation (Poet et al.,
2002), i.v. exposure (Poet et al., 2002; Green 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 (Poet
et al., 2002; Green 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
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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 B6C3F1 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 ALD 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 ALD activity in 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 clearly 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
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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.
4.6.3.2.1. 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 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.
4.6.3.2.2. Dose-response relationships
The incidence 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
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increased incidence of the forestomach neoplasms occurred in groups with ulceration and
hyperplasia, suggesting a dose-dependent relationship between the non-neoplastic and the
neoplastic lesions.
4.6.3.2.3. 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., oral 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 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).
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4.6.3.2.4. Relevance of the hypothesized MO A to humans
While this proposed MOAis 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 margin of exposure 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.
4.6.3.2.5. 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
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 MO A 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 no clear 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 0.2-1 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
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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 bind 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 B6C3F1 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
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 MOAs 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
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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 Corely 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 MOAs 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 expected environmental concentrations, when examining it on its physical-chemical
properties, toxicokinetic and dynamic factors, and MO A 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 CO2 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 B AA-glutamine and BAA-glycine, that 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
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
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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 as 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 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 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 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.
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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, age 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 25-1500 mg EGBE exposures). Two children who had taken greater than 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 adverse 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 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
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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 by older rats to
metabolize the toxic metabolite BAA to C02 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 to 40) and hospitalized children (n = 25 to 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 |im3; 2%) and children (92.8-95.2 |im3; 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 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. In addition, females, but not males, displayed significantly increased urea
nitrogen creatine.
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,
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180, or 360 mg/kg (6 hours/day, nine applications) produced hemoglobinuria in males at
360 mg/kg and 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 maximum blood
concentrations (Cmax) of BAA were 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,
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 benchmark dose (BMDL), 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 subchronically 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 a well-conducted
one, 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 mode of action information published since the 1999 EGBE
toxicological review is included in this document, and this information supports the hemosiderin
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deposition endpoint as an important, key event in the proposed MO A. A comparison of the
NOAELs and LOAELs for the candidate studies are summarized in Table 5-1.
Table 5-1. Results of candidate studies
Species
(strain)
No./dose
Effect levels (ppm)
Reference
Gender
group
Duration/effect
NOAEL
LOAEL
NTP (2000)
Rat (F344)
M
9-10
50
14 week, hematologic
2 year, hematologic,
hemosiderin
31
62.5
31
F
9-10
50
14 week, hematologic
2 year, hematologic,
hemosiderin
-
31
31
NTP (2000)
Mouse
(B6C3F1)
M
9-10
2 year, histopathology of
the forestomach
62.5
50
2 year, hematologic,
hemosiderin
62.5
125
F
9-10
50
2 year, histopathology of
the forestomach
2 year, hematologic,
hemosiderin
-
62.5
62.5
Tyl et al.
(1984)
Rat (F344)
F
30-31
GD 6-15, hematologic
50 ppm
100 ppm
Nelson et al.
(1984)
Rat
(Sprague-
Dawley)
F
15
GD 7-15, hematologic
150 ppm
200 ppm
Dodd et al.
(1983)
Rat (F344)
M, F
10
13 week, hematologic
25 ppm
77 ppm
The primary effects of EGBE exposure, hematological effects, 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
benchmark dose (BMD) analysis, the proper benchmark response (BMR) level for the BMD
derivation was unclear. In addition, while these hematologic effects were observed in both the
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1 subchronic and chronic studies and persisted with exposure duration, they did not progress in
2 severity in the subchronic-to-chronic study (see Table 4-3, Table 4-6, and Table 5-2). Further,
3 better model fits were obtained from the BMD analysis of the subchronic study, which used two
4 more exposure concentrations than the chronic study. For these reasons, the hematologic
5 responses from the 14-week subchronic study were chosen for use in the BMD analyses of this
6 endpoint (see Section 5.1.2.2.1). Selection of the most appropriate hematologic endpoints for use
7 in the BMD analysis also required consideration of EGBE's MOA for hemolysis.
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 week
2 year
0/10
15/50
0/10
19/50
10/10§
36/50°
10/10*§
47/50°
9/9®
NT
5/5§
NT
RBC count*
14 week
1 year
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 week
2 year
0/10
23/50
0/10
30/50
0/10
34/50°
7/10®
42/50°
10/10§
NT
10/10§
NT
RBC count"
14 week
1 year
9.05±0.08
8.88±0.08
8.71±0.14n (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 week
2 year
0/10
0/50
0/10
NT
0/10
5/50°
0/10
25/49§
10/10§
44/50§
6/6§
NT
RBC count"
14 week
1 year
9.72±0.05
9.32±0.09
9.55±0.06d (98)
NT
9.51±0.06°
(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 week
2 year
0/10
0/50
0/10
NT
0/10
0/50
0/10
8/49§
0/10
30/49§
6/6§
NT
RBC count"
14 week
1 year
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±0.06° (98)
9.36±0.32° (98)
8.90±0.07§ (92)
8.33±0.10§ (87)
7.21±0.23§ (74)
NT
*Mean ± standard error; percent of control in parentheses.
#RBC count as 106/|iL.
nStatistically significant difference, p < 0.05.
§ Statistically significant difference, p < 0.01; NT = not tested.
Source: NTP (2000).
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The suggested MOA of EGBE hemolysis is based on data indicating that BAA, an
oxidative metabolite of EGBE, the first hypothesized event in the MOA, is likely to be the
causative agent in hemolysis (Ghanayem et al., 1990, 1987b; Carpenter et al., 1956). The second
event in the MOA is erythrocyte swelling, and cell lysis, which is believed to be preceded by an
increase in the osmotic fragility and a loss of deformability of the erythrocyte (Udden, 1995b,
1994; Udden and Patton, 1994; Ghanayem, 1989). This results in decreased values for RBC
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 clear 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) was apparently 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
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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. 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. (1988) to obtain NOAEL and BMCL estimates.
5.1.2.1. Derivation of the POD Using PBPK Modeling and the NOAEL/LOAEL Method
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. (1994, 1997) 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 Area under the curve (AUC), 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 B6C3F1 mice
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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 (Carpenter et al., 1956; Ghanayem et al., 1987b, 1990). 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). Finally, as
is discussed in Section 5.1.1, hematological endpoints indicative of hemolysis do not progress
with increased inhalation duration. 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 5-3. Johanson
(1986) and Shyr et al. (1993) do not address BAA distribution, and are only parameterized for
humans and rats, respectively. 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. That
model is used here to estimate the Cmax of BAA in blood following inhalation exposure to female
rats, the more sensitive gender.6 The human PBPK model of Corley et al. (1994, 1997) was then
used to obtain estimates of human inhalation exposure concentrations associated with the female
rat BAA blood concentrations.7 Established EPA (2006c) methods and procedures were used to
review, select and apply these chosen PBPK models. 8
6 The Lee et al. (1998) model was chosen in this case because it is an extension of the Corley et al. (1994, 1997)
model that includes added parameters for female rats.
7 The basic components of the Corley model are summarized in Appendix A.
8 EPA 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 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
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 (Lee et al., 1998) by using the actual experimental exposure
regimen (6 hours/day, 5 days/week) in model simulations.
Female rat LOAEL = 31 ppm
Cmax BAA= 41 nM
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 days/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
in air (ppm)
Cmax BAA in blood
(MM)
1
2.6
5
13.0
10
26.1
20
52.9
50
137.1
100
295.0
Step 4: Calculate the LOAELrec 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 =41 |iM
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HEC continuous exposure =18 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
LOAELrec (mg/m3) = conversion factor x LOAELrec (ppm)
= 4.84 (mg/m3) ^ (ppm) x 18 ppm
= 88 mg/m3
5.1.2.2. Derivation of the POD Using PBPK Modeling and BMD Modeling Methods
It is recognized that the NOAEL/LOAEL designations listed in Table 5-1 do not
necessarily indicate the slope of the concentration-response curve, an important factor in
benchmark concentration (BMC) analysis, used to assess inhalation studies in the same manner
as BMDs are used to assess oral studies (U.S. EPA, 2000, 1995b). For this reason, BMC analyses
were performed on selected hemolytic endpoints from the subchronic study and on the
hemosiderin staining endpoint from the chronic study in male and female rats (NTP, 2000).
5.1.2.2.1. BMD approach applied to hematological 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). The current BMD technical guidelines (U.S. EPA, 2000)
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 BMCL05 was considered to be an appropriate POD for derivation of the RfC (U.S. EPA,
2000, 1995b). 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 (see Appendix B for the 1 SD data). Cmax for BAA in
arterial blood of rats was determined by using the PBPK model of Lee et al. (1998). 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
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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 less than 10%,
even if 50% of an individual's skin is exposed. The results of this modeling effort are
summarized in Table 5-4.
Table 5-4. 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 (jliIVI)
31
216
40.4
61.5
211
85.9
125
214
189.5
250
210
451.3
500
201
1143.3
Source: Lee et al. (1998).
All BMD analyses were performed using models in EPABMD software (BMDS),
version 1.4.1 (U.S. EPA, 2000). Graphical figures and text output files for selected BMC
analyses are provided in Appendix B. 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 5-5. All models were fit using restrictions and option
settings suggested in the EPABMD technical guidance document (U.S. EPA, 2000). The best
model fit to these data, from visual inspection and comparison of Akaike Information Criterion
(AIC) values, was obtained using the Hill model. The BMCLos was determined to be 37.2 |iM,
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. (1997) PBPK model was used to back-calculate an HEC of
17 ppm (81.4 mg/m3), assuming continuous exposure (24 hours/day).
Table 5-5. 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.
Model
BMC 05 OiM)
BMCLos (jliIVI)
p Value
AIC*
Scaled residual"
2nd degree polynomial
60.8414
54.5902
<0.0001
-90.947514
0.15
Power
74.5253
68.2041
<0.0001
-52.341720
0.218
Hill"*
42.2297
37.1792
0.1669
-109.272293
0.168
*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).
**Chi-square 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 model fit in this region.
***Model choice based on adequate p value (>0.1), visual inspection, low AIC and low (absolute) scaled residual. .
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5.1.2.2.2. 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, 2000) 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, 2000, 1995b). All BMD assessments in this review were performed using
EPABMDS version 1.4.1. Graphical figures and text output files for selected BMC analyses are
provided in Appendix B.
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 5-6 reports AUC BAA blood concentrations measured at 12 months9 published
by Dill et al. (1998) in both genders of B6C3F1 mice and F344 rats exposed to the same
concentrations used in the NTP (2000) chronic studies of these test animals.
Table 5-6. AUC BAA blood concentrations measured at 12 months in both sexesgenders of
B6C3F1 mice and F344 rats.
Exposure Concentration (ppm)
Gender
AUCbaa (jimol-h/L)*
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
1128.9
50.9
125
Male
9
2225.6
71.1
Female
12
3461.8
154.8
Mice
62.5
Male
10
1206.6
205.6
Female
12
1863.6
112.4
125
Male
9
2819.8
685.1
Female
6
5451.6
508.9
250
Male
10
17951.5
1770.4
Female
11
18297.1
609.7
Authors reported AUC values in terms of |ag-min/g, which were converted to units consistent with the PBPK model
of |-imol-h/L by dividing by 60 min/h and 132.16 g/mol and multiplying by 1060 g/L.
Source: Dill et al. (1998).
9 Dill 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 5-7. All models were fit using restrictions and option settings suggested
in the EPABMD technical guidance document (U.S. EPA, 2000). 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 BMC 10 was 196 |j,mol-hour/L and the
BMCL10 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-h/L, respectively. Assuming
continuous exposure (24 h/day), the Corley et al. (1997) PBPK model was used to back-calculate
human equivalent concentrations of 3.4 ppm (16 mg/m3) from the male rat data and 4.9 ppm (24
mg/m3) from the female rat data,
Likewise, the fit statistics and BMC information for male and female mouse hemosiderin
staining data versus AUC BAA are shown in Table 5-8. All models were fit using restrictions and
option settings suggested in the EPABMD technical guidance document (EPA, 2000). 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 3077 |j,mol-h/L and the BMCLio was determined to be 2448 |j,mol-
h/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 mouse were determined
to be 1735 and 1322 |j,mol-h/L, respectively. Assuming continuous exposure (24 h/day), the
Corley et al. (1997) PBPK model was used to back-calculate human equivalent concentrations of
36 ppm (174 mg/m3) from the male mouse data and 20 ppm (97mg/m3) from the female mouse
data.
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Table 5-7. 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
BMC10 (jimol-h/L)
BMCL10 (jimol-h/L)
p value
AIC*
Scaled Residual"
Male Rats
Multistage-1st degree***
196.252
133.141
0.8680
247.234
0.441
Gamma***
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
Weibull***
196.253
133.141
0.8680
247.234
0.441
Female Rats
Multistage-1st 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-Logistic"
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 = Akaike Information Criteria = -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)
** Chi-square 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 standard deviation. 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 value (> 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 5-8. 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
(jimol-h/L)
BMCL10
(jimol-h/L)
p value
AIC*
Scaled Residual"
Male Mice
Multistage-1st
2100.07
1613.9
0.3067
117.571
-1.766
degree
Gamma
2725.35
1702.27
0.1452
118.559
1.358
Logistic
6605.45
5333.72
0.0022
127.326
2.789
Log-Logistic
2616.51
1628.48
0.1882
118.02
1.193
Probit
5917.06
4825.09
0.0031
126.405
2.734
Log-Probit "
3076.8
2448.3
0.1290
116.614
1.946
Weibull
2689.76
1687.09
0.1445
118.712
-1.448
Female Mice
Multistage-1st
946.491
769.879
0.3680
142.669
-1.583
degree
Gamma
1402.92
818.367
0.3420
143.288
-0.817
Logistic
2897.15
2341.03
0.0002
162.338
-0.942
Log-Logistic
1705.75
1121.43
0.8223
141.501
-0.343
Probit
2860.03
2364.52
0.0002
161.681
-0.829
Log-Probit "
1734.53
1322.06
0.8237
141.498
-0.315
Weibull
1282.82
804.234
0.2958
143.631
-0.988
* AIC = Akaike Information Criteria = -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)
" Chi-square 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 standard deviation. 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 value (> 0.1), visual inspection, low AIC and low (absolute) scaled residual.
The Log-Probit model provides a slightly better fitthan other models for both genders.
5.1.2.3. Selection of the POD
Consideration of the available data has led to the selection of the two-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 mode of action, EPA concluded this is a
precursor to an adverse effect and is appropriate for use in deriving the RfC. ABMCL10 of 133
|j,mol-h/L for hemosiderin staining in liver of male rats chronically exposed to EGBE (NTP,
2000) was used as the point of departure to calculate the RfC. A human PBPK model (Corley et
al., 1997) was used to back-calculate to a human equivalent concentration of 16 mg/m3 (3.4
ppm) for the BMCLHEC. Atotal uncertainty factor (UF) of 10 was applied to the BMCLHEC: 10
for consideration of intraspecies variation (UFH: human variability) to obtain an RfC of 1.6
mg/m3. The rationale for the application of these UFs is provided in Section 5.1.3. The final
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calculation of an RfC derived from the NOAEL/LOAEL and BMC modeling approaches are
presented in Table 5-9.
Table 5-9. Summary of the application of UFs for RfC derivations using NOAEL/LOAEL
and BMC modeling approaches for male rat liver hemosiderin staining.
Approach*
Factor"
LOAEL
BMCL10
UFh
10
10
ufa
1
1
UFS
1
1
ufl
3
1
ufd
1
1
UF (total)
30
10
POD/UFs = RfC (mg/m3)
88/30 = 2.9
16/10= 1.6
PODs—LOAELhec = 88 mg/m3 based on hematological effects in male and female rats; BMCLiohec = 16 mg/m3
based on hemosiderin staining in male rats.
"The rationale for the selection of these UFs is discussed in Section 5.1.3. UFh—intrahuman variability, UFA—
interspecies variation, UFS—subchronic to chronic extrapolation, UFL—LOAEL to NOAEL extrapolation, UFD—
database insufficiencies.
Source: NTP (2000).
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 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 value of 10 was selected to account for variation in sensitivity within the human
population (UFH). 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. AUF of 10 was used to account for the uncertainty associated with the variability of the
human response to the effects of EGBE. 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
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normal adults. Laboratory animal studies suggest that older animals are more sensitive than
neonates, and that females are more sensitive than males (see Section 4.7). However, actual
human responses to EGBE have not been observed under a broad enough range of exposure
conditions, such as repeat/long-term exposures, and potentially sensitive subjects, such as
individuals predisposed to hemolytic anemia or infants, to warrant the reduction of the UFH
below the default value of 10. 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.
AUF of 1 was used to account for interspecies variation (UFA) for toxicodynamic and
toxicokinetic differences between animals and humans. Traditionally, these components 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. In vivo (Carpenter et al.,
1956) and in vitro (Udden, 2002; Udden and Patton, 1994; Ghanayem and Sullivan, 1993) results
indicate that, toxicodynamically, humans may be less sensitive than rats to the hematological
effects of EGBE.
AUF to account for extrapolation from subchronic to chronic exposure (UFS) was not
needed because the RfC was derived from a chronic study.
AUF for LOAEL-to-NOAEL extrapolation was not used because the current approach is
to address this factor as one of the considerations in selecting a BMR for BMD modeling. In this
case, a BMR of a 10% change in hemosiderin staining was selected under an assumption that it
represents a biologically significant change.
A value of 1 was used for the database UF (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 limited studies available
looking at humans following short-term inhalation exposure
The combined PBPK and BMC modeling method using hemosiderin as an endpoint was
used to derive the RfC. In addition, MO A information was used to inform the choice of the
critical effect. The total UF is 10. Thus the RfC, is 16 mg/m3 -M0 = 1.6 mg/m3.
5.1.4. RfC Comparison Information
For comparison purposes, Figure 5-3 presents the POD, applied UFs, and derived
reference values, including the RfC, for the effect endpoints discussed. BMC modeling was done
using EPABMDS version 1.4.1 (U.S. EPA, 2000), and results are provided in Appendices B and
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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 Appendices C
and D.
1000
100
10
\
I
NTP. 2000, 2-yr NTP. 2000; 3 NTP, 2000, 3- NTR 2000; 2 yr NTP, 2000; 2-yr
male and female month fomate rat month mile tm female study, male rat study,
rat study; study. HBC. sttfdy; RBC hemosiderin hemosiderin
hematological BMCL(HEC)05 BMCl(HEC)05 slaving, staining,
effects; IOAEL 3MCL(HEC)10 BMCt{HEC)10
NTP, 2000; 2-yr NTP, 2000; 2-yr
female mouse mate mouse
study,
hemosiderin
staining,
BMCL(HEC)10
study,
hemosiderin
staining,
BMCl(HEC)10
NTP, 2000, 2 yr
female mouse
shidy
forestomach
hyperplasia.
8MCL
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BMCLrec of 81.4 mg/m3) is provided for comparison purposes. 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 |iM-h/L using the AUC for BAA in arterial blood at
12 months and was converted to a BMCLrec of 16 mg/m3 using the Corley et al (1994; 1997)
human PBPK model. The BMCLio for hemosiderin staining in female rats was 244 |iM-h/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 (1994; 1997) 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 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 5-10.
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Table 5-10. Subchronic 91-day drinking water studies in rats and mice
Reference
Species
(strain)
Gender
Animals/dose
Effect
Effect
(mg/k
NOAEL
levels
g-day)
LOAEL
NTP (1993)
Rat
(F344)
M
10
Hepatocellular changes
-
54.9*
F
10
Hematological
-
58.6*
NTP (1993)
Mouse
(B6C3F1)
M
10
Body weight
223
553**
F
10
Body weight
370
676**
Doses 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.
"The 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—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
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observed in EGBE-exposed animals is the most appropriate basis for an RfC/RfD (see Section
5.1.1).
ABMD analysis has also been performed on the hemosiderin pigmentation endpoint
observed in the NTP (2000) chronic EGBE inhalation study (Section 5.1.2.2.2), and PBPK
models have been applied to extrapolate this BMD to a human equivalent oral exposure (Section
5.2.2.2.2). 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 non-bolus 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., 1999). 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
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(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.
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.
Of the available PBPK models (Table 5-3), the Corley et al. (1997, 1994) 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, covers both
oral and inhalation routes of exposure, and addresses both the distribution and excretion of the
toxic metabolite, BAA, following oral EGBE exposure. This model is summarized in Appendix
A. As in the case of the RfC (see Section 5.1.2.1), Cmax is considered a more appropriate dose
metric than AUC for the hematological effects. The PBPK model of Corley et al. (1997, 1994)
was used to obtain estimates of human Cmax concentrations from the female rat drinking water
study data.
The four steps involved in using the Corley et al. (1997, 1994) PBPK model as modified
by Corley et al. (1997) to calculate the HED corresponding to the LOAEL identified in the
animal study (LOAELred) 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.
5.2.2.1. Derivation of POD Using PBPK Model and NOAEL/LOAEL Method
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= 103 |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.
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Step 3: Calculate the Cmax for BAA in blood for humans continuously exposed to varying
concentrations of EGBE.
Table 5-11 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 5-11. 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-day)
Cmax BAA in blood
(MM)
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 =103 |iM
LOAELhed continuous exposure = 7.6 mg/kg-day (calculated by
regression of the internal dose versus the dose of EGBE from step 3)
The LOAELhed calculated using the PBPK model is likely a high estimate of the HED,
since the model is based on male rat kinetic data, and female rats have been observed to have
slightly higher concentrations of BAA in blood than male rats at similar exposure levels. In other
words, use of male rat kinetic data results in estimates of the BAA concentrations in human
blood associated with an effect (LOAELhed) that are lower than if female rat kinetic data had
been used. In addition, 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.
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5.2.2.2. Derivation of POD Using PBPK and BMD Methods
5.2.2.2.1. BMD approach applied to hematological 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, 2000, 1995b), 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.
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, as was determined using the PBPK model of Corley et al. (1994) as modified by
Corley et al. (1997) (see Section 5.2.2.1). The results of this modeling effort are summarized in
Table 5-12.
Table 5-12. Model estimates of BAA blood levels in female rats following oral exposures.
EGBE concentration
in water (ppm)
Water EGBE intake
(L/day)
Female body
weight (g)
BAA in blood
Dose (mg/kg-day)
Cmax (jiM)
750
0.0147
188
59
103
1500
0.0155
185
125
253
3000
0.0125
180
208
495
4500
0.0101
164
277
738
6000
0.0101
150
404
1355
Source: Corley etal. (1997, 1994).
A BMD analysis was performed using EPABMDS version 1.4.1. As can be seen from the
results in Table 5-13, RBC count was the more sensitive of the two hematological endpoints
assessed. All models were fit using restrictions and option settings suggested in the EPA BMD
technical guidance document (U.S. EPA, 2000) except for the choice of BMR.
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Table 5-13. 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 QiM)
BMDLos (jliIVI)
p Value
AIC*
Scaled residual"
RBC count
1st degree polynomial
393.105
325.835
<0.0001
-26.323
-2.35
Power
393.105
325.835
<0.0001
-26.323
-2.35
Hill"*
62.3999
36.2595
0.7038
-48.884
0.204
MCV
1st degree polynomial
229.179
200.063
<0.0001
141.85
3.12
Power
229.179
200.063
<0.0001
141.85
3.12
Hill*"
126.844
104.244
0.1757
120.96
-0.315
* 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 (number of parameters estimated).
"Chi-square 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 value (>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, 2000)
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 BMDLqs was considered to be a more
appropriate POD for derivation of the RfD (U.S. EPA, 2000, 1995b). 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 (1 SD data have been provided in
Appendix B).
The best model fit to the RBC count and MCV data (from visual inspection and
comparison of AIC values and scaled residuals near the BMD) was obtained using a Hill model
(see Table 5-13). 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 in Appendix B.
The BMDo5 was 62 |iM and the BMDL05 was determined to be 36 |iM 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. (1997, 1994) PBPK model was used to back-calculate an HED
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(BMDLred) of 3 mg/kg-day, assuming that rats and humans receive their entire dose of EGBE
from drinking water over a 12-hour period each day.
5.2.2.2.2. BMD approach applied to hemosiderin endpoint.
Due to the limited oral database, EPA believes 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.
5.2.2.3. 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 species-to-species extrapolation used in the development of the RfC, the dose
metric used for species-to-species (rat to human) and route-to-route (inhalation to oral) is the
AUC of BAA at 12 months in arterial blood. This dose metric was used for dose-response
modeling of chronic inhalation data (Section 5.1.2.2.2) to derive the POD of 133 |j,mol-h/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 (BMDLred) of
1.4 mg/kg-day. As for the HED estimations in Sections 5.2.2.1 and 5.2.2.2.1, a simplifying
assumption was used that the entire dose of drinking water EGBE was consumed over a 12-hour
period each day.
5.2.2.4. 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 mode of action information (see Section
4.6.3.1) supports the hemosiderin deposition endpoint as an important key event in the proposed
MO A. The BMCL for the RfC (AUC of 133 |iM-H/L BAA in arterial blood at 12 months) is
converted using the Corley et al. (1994, 1997) model to an oral human equivalent dose
(BMDLhed) of 1.4 mg/kg-day. This extrapolated oral value is consistent with and slightly lower
than the LOAELred of 7.6 mg/kg-day and the BMDLhed of 3 mg/kg-day estimated from the
subchronic oral (NTP, 1993) study. The final calculation of an RfD derived from the
NOAEL/LOAEL and BMC modeling approaches is presented in Table 5-14.
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Table 5-14. Summary of the application of UFs for RfD derivations using NOAEL/LOAEL
and BMD modeling approaches.
Approach*
Factor"
LOAEL
(NTP, 1993)
BMDLio male rat liver hemosiderin staining (NTP, 2000)
UFh
10
10
ufa
1
1
UFS
1
1
ufl
3
1
ufd
1
1
UF (total)
30
10
POD/UFs = RfD (mg/kg-day)
7.6/30 = 0.3
1.4/10 = 0.14
The rationale for the selection of these UFs is discussed in Section 5.1.3 below.
"PODs—LOAELhed = 7.6 mg/kg-day based on hematological effects from the oral study; the BMDLi0Hed =1.4
mg/kg-day, and the approach is based on a route-to-route extrapolation from inhalation data (see Section 5.2.2.3).
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 value of 10 was selected to account for variation in sensitivity within the human
population (UFH). 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. AUF of 10 was used to account for the uncertainty
associated with the variability of the human response to the effects of EGBE. 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, and laboratory animal (rats, calves, and mice)
studies suggest that older animals are more sensitive than neonates, and that females are more
sensitive than males (see Section 4.7). However, actual human responses to EGBE have not been
observed in a broad enough range of exposure conditions (e.g., repeat/long-term exposures) and
potentially sensitive subjects (e.g., individuals predisposed to hemolytic anemia, infants) to
warrant the reduction of the UFH below the default value of 10. 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.
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AUF of 1 was used to account for interspecies variation (UFA) for toxicodynamic and
toxicokinetic differences between animals and humans. Traditionally, these components 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: toxicokinetic and toxicodynamic. In this assessment, the toxicokinetic uncertainty is
addressed by the determination of an HEC using PBPK modeling. There is in vivo (Carpenter
et al., 1956) and in vitro (Udden, 2002; Udden and Patton, 1994; Ghanayem and Sullivan, 1993)
information indicating that, toxicodynamically, humans may be less sensitive than rats to the
hematological effects of EGBE. Thus, a value of 1 was used to account for toxicodynamic
differences between rats and humans.
AUF to account for extrapolating from subchronic to chronic exposure (UFS) was not
needed because the RfC was derived from a chronic study.
AUF for LOAEL-to-NOAEL extrapolation was not used because the current approach is
to address this factor as one of the considerations in selecting a BMR for BMD modeling. In this
case, a BMR of a 10% change in hemosiderin staining was selected under an assumption that it
represents a biologically significant change.
A value of 1 was used for 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. A comparison of NOAELs and BMDLs from the subchronic oral
and chronic inhalation studies using a common internal dose metric (AUC BAA) shows that
effects occur at approximately the same internal dose following inhalation and oral exposures
adding confidence to the application of the chosen PBPK model (Corley et al., 1997, 1994) to
derive an oral RfD from the NTP (2000) chronic inhalation study. Thus, the total UF is 10. The
RfD is 1.4 mg/kg-day -M0 = 0.1 mg/kg-day.
5.2.4. 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, 2000c).
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5.3. Uncertainties in the Derivation of the Inhalation Reference
Concentration (RfC) and Oral Reference Dose (RfD)
The following is a more extensive discussion of uncertainties associated with the RfC and
RfD for EGBE beyond the quantitative discussion in Sections 5.1.2, 5.1.3, 5.2.2, and 5.2.3. A
summary of these uncertainties, along with uncertainties specific to the Section 5.4 cancer
analysis, is presented in Table 5-15.
5.3.1. Choice of endpoint
The impact of endpoint selection on the derivation of the RfC and RfD was discussed in
Sections 5.1.2.4 and 5.2.2.4. Comparison RfC values were also 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 1-18 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;
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 BMCLio(HEC) value would have been
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39 mg/m3, approximately 2.4-fold higher than the 16 mg/m3 BMCLi0(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 Tables 5-9 and 5-14, use of the BMC approach
on the incidence of chronic hemosiderin deposition resulted in RfC and RfD values about
threefold 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 |imol-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, 1994), resulting in a BMCLio(HEC) of 25 mg/m3. This default value would have been
twofold higher than the value derived using the PBPK model.
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 is discussed above, that the metric most closely associated with
the effects seen is the AUC measure in blood of BAA. 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
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(see Section 5.2.2.3) and shows that the values derived using the second procedure were
consistently lower than using the former (i.e., were more potent by a factor of up to fourfold).
Thus, estimates using this procedure for route-to-route extrapolation would uniformly
overestimate the toxicity value and would 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.
10 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.
10 Increased liver hemangiosarcomas were not observed in the NTP (2000) chronic study of rats or female mice.
Possible reasons for this species and gender specificity are discussed in Section 4.6 and 4.7.2, respectively.
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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 supports an MOA that is anticipated to be linear. The
linear approach is used as a default option if the MOA of carcinogenicity is not understood. The
nonlinear approach "can be used for cases with sufficient data to ascertain the mode of action
and to conclude that it is not linear at low doses..(U.S. EPA, 2005). Alternatively, the MOA
may theoretically have 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 MOA of carcinogenicity for hepatic hemangiosarcoma,
hepatocellular adenoma and carcinoma, and forestomach tumor formation in animals is
reasonably well understood. A reference concentration and reference dose approach has been
used for EGBE because "When adequate data on mode of action provide sufficient evidence to
support a nonlinear mode of action for the general population and/or any subpopulations of
concern, a different approach — a reference dose/reference concentration that assumes that
nonlinearity - is used." (U.S. EPA, 2005). 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
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associated with erythrocyte hemolysis, leading to oxidative damage and increased hepatocyte
and endothelial cell proliferation and initiation for the liver tumors, and associated 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 4.2 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 B6C3F1 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 B6C3F1 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 hemolyses 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 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
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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 there was
difficulty distinguishing pheochromocytomas from non-neoplastic 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 non-neoplastic adrenal medullary hyperplasia, this
tumor was not given significant weight in the qualitative or quantitative assessment of EGBE
cancer potential.
5.4.1. 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
5-15.
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Table 5-15. 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 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 t liver hemosiderin) are not
likely to occur in humans at the RfC or
RfD.
Choice of
endpoint
Use of forestomach
endpoint could t 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 t or j
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 t liver hemosiderin
and t liver tumors. AUC chosen because
hemosiderin f 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
t RfC up to threefold (see
Section 5.3).
Multistage (1st degree)
model chosen.
The best-fitting model was chosen based
on EPA (2000) BMD technical guidance.
Choice of animal
to human
extrapolation
method
Alternatives could t or j
RfC/RfD (e.g., default
would t 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 EPA
BMD guidance (U.S.
EPA, 2000).
Limited size of bioassay results in
sampling variability; lower bound is 95%
CI on administered exposure.
Choice of
bioassay
Alternatives could t or j
RfC/RfD.
NTP (2000) study.
Alternative bioassays were inadequate.
Choice of
species/ gender
RfC would be t 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
MOA, 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."
MOAs 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 t 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.
*t = increase; j = decrease.
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5.1.1.1. 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, 2005). The results of this linear analysis are presented in Table 5-16.
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^. 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.
Table 5-16. Illustrative potency estimates for tumors in mice, using a linear analysis
approach.
BMDLiohec
(mg/m3)*
Slope factor
0. 1/BMDLiohec
(risk/mg-m3)
Hepatocellular carcinoma (males)
208
4.8 x KT4
Hemangiosarcomas (males)
575
1.7 x 10"
Papilloma or carcinoma of the forestomach (females)
544
1.8 x 10"
*BMDL10Hec values were calculated using AUC as the dose metric.
5.4.1.2. 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
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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 5-4 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 5-15). 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 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 5-4 provides a graphic of the areas
of uncertainty, described in Table 5-15, for which there is quantitative information and impact on
the RfC can be estimated. The "Cancer Approach" value in Figure 5-4 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 indicates 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 5-16, 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 in
an effort to make apparent the limitations of the assessment and to aid and guide the risk assessor
in the ensuing steps of the risk assessment process.
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Areas of Uncertainty ¦ X-foId Change from Alternative Choices
18
18
2.4
~ ~ ~
2
~
1.4
~
56
1 Nonlinear —» Linear (10 ° risk)
2 Hemosiderin —> Forestomach
3 AU C —+ C mas
4 Low BMDL -.NOAEL
5 Low -.High BMDL
6 PBPK —» Default
7 BMDL — BMD
8 Rat—> Mouse
Cancer . , Dose ROD BMD HEC Benchmark
i -Jr Enopoint « . c c 7
Approach Metric3 Method Model5 Derivation6 POD7 bpecies
Figure 5-4. Potential impact of select uncertainties on the RfC for EGBE.
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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6. Major Conclusions in the 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 clearly 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 did not 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;
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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
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 adverse 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 (>1000 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 B6C3F1 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 B6C3F1 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
EPA guidelines (U.S. EPA, 2005), 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.
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6.2. Dose Response
The quantitative estimates of human risk from lifetime exposure to EGBE are based on
animal experiments, as no relevant human data exist.
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 RfC based on this endpoint is 1.6 mg/m3. This concentration is based on the
human equivalent BMCLio of 16 mg/m3, which was back-calculated from rat data using the
BMD and PBPK approach and the application of a tenfold UF.
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 (Lee et al., 1998; Corley et al., 1997, 1994) and actual
measurements of internal blood concentrations in test aniumals 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. Confidence is not high, because the potential for effects in humans
from repeat, long-term exposures has not been investigated.
In the derivation of the RfC, a tenfold UF was applied, which was intended to account for
intrahuman variability. A value of 10 was selected to account for variation in sensitivity within
the human population (UFH). 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. AUF of 10 was retained to account for the
uncertainty associated with the variability of the human response to the effects of EGBE. Human
responses to EGBE have not been observed in a broad enough range of exposure conditions (e.g.,
repeat/long-term exposures) and potentially sensitive subjects (e.g., individuals predisposed to
hemolytic anemia, infants) to warrant the reduction of the UFH below the default value of 10.
For derivation of the RfD, the BMD PBPK approach, along with the same internal dose
metric used in the derivation of the RfC, was used. As with the RfC, the RfD was based on
hepatic hemosiderin deposition data. The RfD based on this endpoint is 0.1 mg/kg-day. This
value was obtained by using a route-route extrapolation from the RfC, and dividing the estimated
human equivalent by a UF of 10. The HED was estimated using the AUC values for BAA in
blood as the dose metric, and calculating a BMDLio of 133 |iM-hour/L BAA in arterial blood at
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12 months. The PBPK model was then used to back-calculate an HED, assuming that rats and
humans received their entire dose of EGBE from drinking water over a 12-hour period each 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 and case reports and 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. Confidence is not
high, because the potential for effects in humans from repeat, long-term exposures has not been
investigated.
A value of 10 was selected to account for variation in sensitivity within the human
population (UFH). Potentially susceptible subpopulations include individuals with enhanced
metabolism or decreased excretion of BAA and individuals whose RBC membranes are more
susceptible to the BAA-induced hemolysis, the principal precursor event to the critical effect of
hemosiderin deposition in the liver. AUF of 10 was retained to account for the uncertainty
associated with the variability of the human response to the effects of EGBE. Human responses
to in vivo EGBE exposure have not been observed in a broad enough range of exposure
conditions (e.g., repeat or long-term exposures) and potentially sensitive subjects (e.g.,
individuals predisposed to hemolytic anemia, infants) to warrant the reduction of the UFH below
the default value of 10. For a more detailed discussion of the RfD UF, see Section 5.2.3.
Information regarding the reported liver and forestomach tumors observed in laboratory
animals exposed to EGBE indicates that the modes of action 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.
11 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
1' These analyses are consistent with the nonlinear assessment approach described in the 2005 cancer guidelines
(U.S. EPA, 2005).
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of these key events and would serve to prevent the occurrence of carcinogenic effects in humans.
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 Sections 5.1.3 and 5.2.3 for the RfC
and RfD, respectively. 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. Table 5-9 and Table 5-14 show 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 toxicological review are summarized in
Table 5-15 and in Sections 5.4.1 and 5.5.
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Appendix A
Corley et al. (1997,1994) PBPK Model
1 Corley et al. (1994) developed physiologically based pharmacokinetic (PBPK) models
2 for rats and humans with the primary objective of describing the concentration of 2-butoxyacetic
3 acid (BAA) in the target tissue (blood) of rats and humans for use in risk assessment (Figure
4 B-l). The models incorporate allometrically scalable physiological and biochemical parameters
5 (e.g., blood flows, tissue volumes, and metabolic capacity) in place of the standard values for a
6 70 kg human. These parameters normalize standard values to the actual body weights of the
7 subjects in several human kinetic studies. The physiology of humans under exercise conditions
8 was maintained in the model. The rat was included to expand the database for model validation
9 and to assist in interspecies comparisons of target tissue doses (BAA in blood).
Model for 2-Butoxyethanol
Inhalation
Model for 2-Butoxyacetic Acid
IV Infusion
Exhalation
Van
Ar arial Blood
Bio
od
Dermal
Vapc
Metabolism
Butoxyacstic Acid
Drinking Watar
or Gavage
Other Metabolitea
Urine
Liver
Liver
Kidney
Lungs and Arterial
Blood
Lungs and Arterial
Blood
Gaatrolnteatlnal
Tract
Gastrointestinal
Tract
Rapidly Perfused
Organs
Rapidly Perfused
Organs
8lowly Perfused
Organs
8lowly Perfused
Organs
Muscle
Skin
Fat
Skin
Muscle
Fat
(Butoxyacetlc Add)
Figure A-l. PBPK model of Corley et al. (1994). The formation of BAA from ethylene
glycol monobutyl ether 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).
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The Corley et al. (1994) model included additional routes of exposure such as oral
(gavage), drinking water, intravenous (i.v.) infusion, and dermal (liquids and vapor). 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 ethylene glycol
monobutyl ether (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 ethylene glycol [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.
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 0.64-4.33 for EGBE and
0.77-1.58 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. Significant increases in
the concentrations of EGBE were observed by Corley et al. (1997) in the first postexposure
blood samples. Because the subjects were able to move their arms freely after the exposure,
Corley et al. (1997) hypothesized that the local blood flow to the exposed arm increased for a
few minutes postexposure. By adjusting the blood flow to the skin by fourfold for 5 minutes
postexposure, the model is able to simulate this change in the concentration of EGBE in blood.
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Table A-l. Selected parameters used in the PBPK model for EGBE developed by
Corley et al. (1994)
Parameter
Human
Rat
Weights
Body weight (kg)
70
0.23
Liver
3.14%
2.53
Rapidly perfused
3.71%
5.1
(BAA model)
3.27%
4.39%
Slowly perfused
9.4
24.8
Flows (L/hour)
Alveolar ventilation
347.9
5.06
Cardiac output (COP)
347.9
5.06
Liver (% COP)
25.0
25.0
Rapidly perfused (% COP)
50.0
51.0
(BAA model)
25.0
26.0
Slowly perfused (% COP)
2
2
Partition coefficients
Blood/air
7,965
7,965
Liver/blood
1.46
1.46
(BAA model)
1.30
1.30
Rapidly perfused/blood
1.46
1.46
(BAA model)
1.30
1.30
Slowly perfused/blood
0.64
0.64
(BAA model)
1.31
1.31
Metabolic constants
EGBE to BAA
375
375
Kml (mg/L)
26.9
26.9
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Appendix B
Text Output from Benchmark Dose Software Runs Used in the
Derivation of RfC and RfD Values
Hill Model with 0.95 Confidence Level
0 200 400 600 800 1000 1200
Cmax
BMD
Hill Model, Version Number: 1.4.1
Input Data File: C:\BMDS\DATA\EGBE\REDBLOODCELL_F.(D)
Tue Mar 06 09:07:25 2007
BMD Method for RfC: Red blood cell count in female rats versus Cmax BAA, 14 week inhalation
study (NTP, 2000)
BMDS MODEL RUN
The form of the response function is:
Y[dose] = intercept + v*doseAn/(kAn + doseAn)
Dependent variable = MEAN
Independent variable = CmaxInternalD
Power parameter restricted to be greater than 1
The variance is to be modeled as Var(i) = exp(lalpha + rho * ln(mean(i)))
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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
lalpha = -3.17322
rho = 0
intercept = 8.48
v = -3.71
n = 1.07589
k = 278.325
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -n
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
lalpha rho intercept v k
lalpha 1 -1 0.11 0.2 -0.21
rho -1 1-0.11 -0.2 0.21
intercept0.ll -0.11 1 0.27 -0.61
v 0.2-0.2 0.27 1 -0.88
k -0.21 0.21 -0.61 -0.88 1
Parameter Estimates
95.0% Wald Confidence Interval
Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. Limit
lalpha -1.10102 2.15389 -5.32256 3.12053
rho-1.07311 1.09214-3.21366 1.06743
intercept 8.47695 0.0493243 8.38028 8.57363
v -5.12159 0.285841 -5.68183 -4.56135
n 1 NA
k 468.055 60.889 348.715 587.395
NA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
has no standard error.
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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 0.16 0.183 0.0526
40.4 10 8.08 8.07 0.22 0.188 0.168
85.9 10 7.7 7.68 0.25 0.193 0.282
189.5 10 6.91 7 0.16 0.203 -1.42
451.3 9 6.07 5.96 0.12 0.221 1.45
1143 5 4.77 4.84 0.33 0.247 -0.66
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)} = exp(lalpha + rho*ln(Mu(i)))
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 61.856962 7 -109.713924
A2 66.476628 12 -108.953256
A3 62.170236 8 -108.340472
fitted 59.636147 5 -109.272293
R-34.183663 2 72.367326
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)
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
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Test -2*log(Likelihood Ratio) Test df p-value
Test 1 201.321 10 <0001
Test 2 9.23933 5 0.09989
Test 3 8.61278 4 0.07154
Test 4 5.06818 3 0.1669
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 less than .1. A non-homogeneous variance
model appears to be appropriate
The p-value for Test 3 is less than . 1. You may want to consider a
different variance model
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 = 42.2297
BMDL = 37.1792
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Hill Model with 0.95 Confidence Level
BMD
Hill Model. (Version: 2.12; Date: 02/20/2007)
Input Data File: U:\BMDS\D AT A\EGBE\EGBE_M_14WK.(d)
Gnuplot Plotting File: U:\BMDS\DATA\EGBE\EGBE_M_14WK.plt
Thu Jul 05 16:39:03 2007
0 200 400 600 800 1000
dose
16:39 07/05 2007
BMD Method for RfC: Red blood cell count in male rats versus Cmax BAA, 14 week inhalation
study (NTP, 2000)
The form of the response function is:
Y[dose] = intercept + v*doseAn/(kAn + doseAn)
Dependent variable = MEAN
Independent variable = Cmax-90d
Power parameter restricted to be greater than 1
The variance is to be modeled as Var(i) = exp(lalpha + rho * ln(mean(i)))
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
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Default Initial Parameter Values
lalpha =-2.62607
rho = 0
intercept = 9.05
v = -3.08
n = 1.56979
k = 333.008
Asymptotic Correlation Matrix of Parameter Estimates
lalpha rho intercept v n k
lalpha 1 -1 0.0097 0.064 0.075 -0.09
rho -1 1 -0.014 -0.064 -0.075 0.091
intercept 0.0097 -0.014 1 -0.63 -0.64 0.25
v 0.064 -0.064 -0.63 1 0.96 -0.9
n 0.075 -0.075 -0.64 0.96 1 -0.82
k-0.09 0.091 0.25 -0.9-0.82 1
Parameter Estimates
95.0% Wald Confidence Interval
Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. Limit
lalpha -11.3409 2.55713 -16.3527 -6.32896
rho 4.21018 1.23612 1.78742 6.63294
intercept 9.02037 0.0937743 8.83657 9.20416
v -3.96474 0.470028 -4.88597 -3.0435
n 1.39015 0.229376 0.940583 1.83972
k 462.11 80.7255 303.891 620.329
Table of Data and Estimated Values of Interest
Dose N Obs Mean Est Mean Obs Std Dev Est Std Dev Scaled Res.
0 10 9.05 9.02 0.253 0.353 0.265
40.8 10 8.71 8.89 0.443 0.343 -1.65
86.6 10 8.91 8.67 0.19 0.325 2.35
190.7 10 8.01 8.12 0.253 0.283 -1.27
449.7 10 7.1 7.08 0.221 0.212 0.366
1096 10 5.97 5.97 0.158 0.148 -0.0614
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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)} = exp(lalpha + rho*ln(Mu(i)))
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 51.943035 7 -89.886070
A2 58.944827 12 -93.889653
A3 54.680821 8 -93.361643
fitted 49.559802 6 -87.119604
R-37.662661 2 79.325322
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)
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 1 193.215 10 <0001
Test 2 14.0036 5 0.01559
Test 3 8.52801 4 0.07404
Test 4 10.242 2 0.00597
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
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The p-value for Test 2 is less than .1. A non-homogeneous variance
model appears to be appropriate
The p-value for Test 3 is less than .1. You may want to consider a
different variance model
The p-value for Test 4 is less than . 1. You may want to try a different
model
Benchmark Dose Computation
Specified effect = 0.05
Risk Type = Relative risk
Confidence level = 0.95
BMD = 105.535
BMDL = 71.4622
Multistage Model with 0.95 Confidence Level
Multistage
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14:35 02/28 2007
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Multistage Model, Version Number: 1.4.1
Input Data File: C:\BMDS4ME\DATA\\EGBE\HEMOSIDERIN_M.(D)
Wed Feb 28 01:46:52 2007
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)]
The parameter betas are restricted to be positive
Dependent variable = HemosiderinM
Independent variable = MAUCDose3mo
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 = 0.464249
Beta(l) = 0.00104028
Asymptotic Correlation Matrix of Parameter Estimates
B ackground B eta( 1)
Background 1 -0.67
Beta(l) -0.67 1
Parameter Estimates
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95.0% Wald Confidence Interval
Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. Limit
Background 0.464473 0.0851405 0.297601 0.631346
Beta(l) 0.00103871 0.000331559 0.000388869 0.00168856
Analysis of Deviance Table
Model Log(likelihood) # Param's Deviance Test d.f. P-value
Full model -121.475 4
Fitted model -121.506 2 0.0626604 2 0.9692
Reduced model -130.097 1 17.2453 3 0.0006292
AIC: 247.012
Goodness of Fit
Scaled
Dose Est. Prob. Expected Observed Size Residual
0.0000 0.4645 23.224 23 50 -0.063
250.0000 0.5869 29.347 30 50 0.187
525.9000 0.6899 34.494 34 50 -0.151
1154.5000 0.8386 41.929 42 50 0.027
ChiA2 = 0.06 d.f. = 2 P-value = 0.9691
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 101.434
BMDL = 69.6269
BMDU = 174.647
Taken together, (69.6269, 174.647) is a 90% two-sided confidence
interval for the BMD
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Log-Logistic Model with 0.95 Confidence Level
Log-Logistic
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Log-Logistic Model, Version Number: 1.4.1
Input Data File: C:\BMDS4ME\DATA\\EGBE\HEMOSIDERIN_F.(D)
Wed Feb 28 02:03:32 2007
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 = HemosiderinF
Independent variable = FAUCDose3mo
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
Parameter Convergence has been set to: le-008
User has chosen the log transformed model
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Default Initial Parameter Values
background = 0.3
intercept = -17.8009
slope = 2.88261
Asymptotic Correlation Matrix of Parameter Estimates
background intercept slope
background 1 -0.39 0.35
intercept -0.39 1 -1
slope 0.35 -1 1
Parameter Estimates
95.0% Wald Confidence Interval
Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. Limit
background 0.291372 0.0594025 0.174945 0.407799
intercept -16.9704 4.29214 -25.3829 -8.55798
slope 2.77044 0.669882 1.45749 4.08338
Analysis of Deviance Table
Model Log(likelihood) # Param's Deviance Test d.f. P-value
Full model -104.742 4
Fitted model-104.839 3 0.192188 1 0.6611
Reduced model -135.725 1 61.9658 3 <.0001
AIC: 215.677
Goodness of Fit
Scaled
Dose Est. Prob. Expected Observed Size Residual
0.0000 0.2914 14.569 15 50 0.134
245.0000 0.3981 19.906 19 50 -0.262
515.5000 0.7039 35.193 36 50 0.250
1131.0000 0.9466 47.332 47 50 -0.209
ChiA2 = 0.19 d.f. = 1 P-value = 0.6606
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Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 206.942
BMDL = 120.82
Hill Model with 0.95 Confidence
8.5
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(/>
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BMDL. BMD
20 40 60 80
Cmax BAA (jjM)
100
120
140
Hill Model. (Version: 2.12; Date: 02/20/2007)
Input Data File: U:\BMDS\DATA\EGBE\EGBEORAL.(d)
Gnuplot Plotting File: U:\BMDS\DATA\EGBE\EGBEORAL.plt
Wed Mar 07 08:21:49 2007
BMD Method for RfD: RBC Response in Orally Exposed Female Rats (NTP, 1993)
The form of the response function is:
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Modeling Results Using 1SD as the BMR for Blood Effects
1 Table 5-5 and Table 5-13 present the modeling results for the potential blood effect endpoints
2 considered for deriving a reference concentration (RfC) and reference dose (RfD) by using a
3 benchmark response (BMR) of 5% relative change in the mean. The following tables present the
4 same effects using a BMR of 1 standard deviation (SD). A discussion for why this BMR was not
5 considered the best choice can be found in Section 5.1.2.2.1 for the RfC and in Section 5.2.2.2.1
6 for the RfD.
Table B-l. Comparison of BMC/BMCL values for female rat RBC count data from
inhalation subchronic study (14 week) using modeled blood Cmax (3 months) of the EGBE
metabolite BAA as a common dose metric
Model
BMC1SD QiM)
BMCL1SD OiM)
p Value
AIC*
Scaled residual"
2nd degree polynomial
31.5363
22.6222
<0.0001
-90.947514
0.15
Power
41.9568
33.8953
<0.0001
-52.341720
0.218
Hill"*
17.361
13.476
0.1669
-109.272293
0.168
*AIC = Akaike Information Criterion = -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).
**Chi-square 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 model fit in this region.
**Model choice based on adequate p value (>0.1), visual inspection, low AIC, and low (absolute) scaled residual.
Table B-2. Comparison of BMC/BMCL values for female rat RBC count and MCV from
an oral subchronic study using modeled blood Cmax (3 months) of the EGBE metabolite AA
as a common dose metric
Model
BMC1SD QiM)
BMCL1SD QiM)
p Value
AIC*
Scaled residual**
RBC count
1st degree polynomial
474.233
374.108
<0.0001
-26.323
-1.35
Power
474.233
374.108
<0.0001
-26.323
-1.35
Hill"*
56.5593
32.0476
0.7038
-48.884
0.204
MCV
1st degree polynomial
108.691
76.203
<0.0001
141.85
-0.223
Power
108.691
76.203
<0.0001
141.85
-0.223
Hill
40.516
failed
0.1768
120.94
-0.317
* 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 (number of parameters estimated).
"Chi-square 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 model fit in this region.
"Model choice based on adequate p value (>0.1), visual inspection, low AIC, and low (absolute) scaled residual.
Source: NTP (1993).
7 ====================================================================
8 Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
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Y[dose] = intercept + v*doseAn/(kAn + doseAn)
Dependent variable = MEAN
Independent variable = Cmax
rho is set to 0
Power parameter restricted to be greater than 1
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= 1.30181
k = 170.5
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 1.4e-009 1.2e-007 -7.3e-008
intercept 1.4e-009 1 -0.41 -0.51
v 1.2e-007 -0.41 1 -0.48
k -7.3e-008 -0.51 -0.48 1
Parameter Estimates
95.0% Wald Confidence Interval
Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. Limit
alpha 0.142549 0.0260259 0.0915397 0.193559
intercept 8.15021 0.117707 7.91951 8.38092
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v -1.75852 0.187534 -2.12608 -1.39096
n 1 NA
k 206.872 79.7468 50.5715 363.173
NA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
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 8.15 0.28 0.378 -0.0018
103 10 7.59 7.57 0.48 0.378 0.204
253 10 7.09 7.18 0.43 0.378 -0.777
495 10 7 6.91 0.37 0.378 0.754
738 10 6.8 6.78 0.36 0.378 0.195
1355 10 6.58 6.62 0.41 0.378 -0.374
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 28.441981 4 -48.883963
R-3.574142 2 11.148285
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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)
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 68.638 10 <0001
Test 2 3.19854 5 0.6694
Test 3 3.19854 5 0.6694
Test 4 1.40726 3 0.7038
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 = 62.3999
BMDL = 36.2595
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Input Data File: C:\BMDS\Data\EGBE_M_HC.(d)
Gnuplot Plotting File: C:\BMDS\Data\EGBE_M_HC.(d)
ThuFeb 21 10:55:46 2008
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (l-background)*[l-EXP(
-beta 1 * doseA 1 -beta2 * doseA2-beta3 * doseA3)]
The parameter betas are restricted to be positive
Dependent variable = Hep Car ADJ M
Independent variable = AUC Mouse M
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 4
Total number of specified parameters = 0
Degree of polynomial = 3
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 = 0.209552
Beta(l) = 0.000187331
Beta(2) = 0
Beta(3) = 0
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(2) -Beta(3)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
B ackground B eta( 1)
Background 1 -0.71
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Beta(l) -0.71 1
Parameter Estimates
95.0% Wald Confidence Interval
Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. Limit
Background 0.195776 * * *
Beta(l) 0.000198015 * * *
Beta(2) 0 * * *
Beta(3) 0 * * *
* - Indicates that this value is not calculated.
Warning: Likelihood for the fitted model larger than the Likelihood for the full model.
Error in computing chi-square; returning 2
Analysis of Deviance Table
Model Log(likelihood) # Param's Deviance Test d.f. P-value
Full model -119.553 4
Fitted model -118.389 2 -2.32685 2 2
Reduced model -124.691 1 10.2756 3 0.01636
AIC: 240.779
Goodness of Fit
Scaled
Dose Est. Prob. Expected Observed Size Residual
0.0000 0.1958 9.867 10 50 0.189
402.0000 0.2573 12.981 11 50 -0.493
925.0000 0.3304 16.209 19 49 0.866
2120.0000 0.4715 23.334 22 49 -0.240
ChiA2 = 1.09 d.f. = 2 P-value = 0.5809
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
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1 BMD = 532.083
2
3 BMDL = 333.797
4
5 BMDU= 1475.14
6
7 Taken together, (333.797, 1475.14) is a 90 % two-sided confidence
8 interval for the BMD
9
10 Multistage Cancer Slope Factor = 0.000299583
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Appendix C
Benchmark Dose Assessment of Forestomach Lesions in Female
Mice Using PBPK Models to Estimate Human Equivalent
Exposures
Several PBPK models have been developed for EGBE, all of which are capable of
estimating internal doses. These models are summarized briefly in Section 5 of this assessment.
Consistent with the assessment approach outlined in Section 5, the Lee et al. (1998) model was
used to estimate internal dose levels from inhalation exposures to the experimental animal, in
this case the female mouse.
Peak concentration in the blood during the exposure period (Cmax) of BAA was used in
the derivation of the RfC when using hematological endpoints because of (1) convincing
evidence that BAA is the causative agent for EGBE-induced hemolysis and (2) EGBE-induced
hemolysis appears to be highly dependent on the BAA concentration attained. Area under the
curve was used in the derivation of the RfC during use of the hemosiderin deposition as the
critical endpoint, because this endpoint increases in severity with increasing duration of exposure
and because it is representative of a continuous process and not a one-time event.
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 non-bolus 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 apparently 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, intraperitoneal, 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-
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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
mode of action. 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 mode of action 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
underlie the hepatic tumors and that are the basis of the oral RfD and inhalation RfC. A
benchmark dose (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) a BMDLio value was
estimated 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 human equivalent
dose/concentration that resulted in the same internal dose (Cmax BAA) simulated for the animal
in Step 1.
Step 1: Estimation ofBMDLio (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-l.
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Table C-l. PBPK model estimates of BAA Cmax blood levels and incidence of forestomach
epithelial hyperplasia in female mice
Air concentration (ppm)
Cmax BAA (|LllVI)
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 EPA benchmark dose software (BMDS). The estimates for each model, along
with statistical goodness-of-fit information, are provided in Table C-2.
Table C-2. BMDS model estimates of Cmax BMDi0 and BMDLio values for forestomach
epithelial hyperplasia in female mice
BMDS model
BMD (julVl)
BMDL (julVl)
AIC
(lowest = best fit)
p Value
(>0.1 = adequate fit)
Gamma
420.56
266.87
151.16
0.5287
Logistic
544.757
444.896
162.191
0.0067
Log-logistic
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
Step 2: Verification of steady state.
As can be seen from Table C-3, 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.
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Table C-3. Female mouse Cmax values for various time points of the NTP (2000) study
estimated by the Lee et al. (1998) model
Months 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
Log-logistic Model with 0.95 Confidence Level
1
Log-logistic
0.8
0.6
0.4
0.2
0
BMD
BMDL
0
500
1000
1500
2000
2500
Concentration (pM)
Figure C-l. BMD plot ef fraction of female mice with forestomach epithelial hyperplasia
following inhalation exposure (NTP, 2000) versus internal dose metric (BAA Cmax, jiM).
1 Step 3: Simulation of internal human doses.
2 The tables below summarize the results of model simulations of the internal dose
3 surrogate (Cmax BAA in blood) for a 70 kg human who consumes an average of 2 L
4 of drinking water during a 12-hour awake cycle (Table C-4) or is continuously
5 exposed to air concentrations (Table C-5) of EGBE.
6
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Table C-4. 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-day)
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
Source: Corley etal. (1997, 1994).
Table C-5. 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 (jliIVI)
1
2.6
5
13.0
10
26.1
20
52.9
50
137.1
100
295.0
200
733.7
Source: Corley etal. (1997, 1994).
Step 4: Calculate the human equivalent dose/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 320 |iM 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 human equivalent air concentration of 551 mg/m3
(113 ppm) from the Cmax BMDLio of 320 |iM estimated in step 1, assuming
continuous exposure (24 hours/day).
Conclusion
The PODs calculated above are significantly higher than the PODs of 1 mg/kg-day and 12
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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Log-Logistic Model Results for
Forestomach Epithelial Hyperplasia
in Female Mice (NTP, 2000)
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
BMDS MODEL RUN
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
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
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11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
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
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
April 2008
C-7
DRAFT-DO NOT CITE OR QUOTE
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