EPA 635/R-03/009
www.epa.gov/iris
r/EPA
TOXICOLOGICAL REVIEW
OF
METHYL ETHYL KETONE
(CAS No. 78-93-3)
In Support of Summary Information on the
Integrated Risk Information System (IRIS)
September 2003
U.S. Environmental Protection Agency
Washington, DC
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DISCLAIMER
This document has been reviewed in accordance with U.S. Environmental Protection
Agency policy and approved for publication. Mention of trade names or commercial products
does not constitute endorsement or recommendation for use. Note: This document may undergo
revisions in the future. The most up-to-date version will be made available electronically via the
IRIS Home Page at http://www.epa.gov/iris.
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CONTENTS —TOXICOLOGICAL REVIEW OF
METHYL ETHYL KETONE (CAS No. 78-93-3)
LIST OF TABLES vi
LIST OF FIGURES vii
FOREWORD viii
AUTHORS, CONTRIBUTORS, AND REVIEWERS ix
1. INTRODUCTION 1
2. CHEMICAL AND PHYSICAL INFORMATION RELEVANT TO ASSESSMENTS 3
3. TOXICOKINETICS RELEVANT TO ASSESSMENTS 5
3.1. ABSORPTION 5
3.1.1. Oral Exposure 5
3.1.2. Inhalation Exposure 5
3.1.3. Dermal Exposure 6
3.2. DISTRIBUTION 7
3.3. METABOLISM 8
3.4. ELIMINATION AND EXCRETION 11
3.5. PHYSIOLOGICALLY-BASED PHARMACOKINETIC (PBPK) MODELS 12
4. HAZARD IDENTIFICATION 17
4.1. STUDIES IN HUMANS - EPIDEMIOLOGY, CASE REPORTS, CLINICAL
CONTROLS 17
4.1.1. Oral Exposure 17
4.1.2. Inhalation Exposure 17
4.1.2.1. Acute Exposure 17
4.1.2.2. Case Studies of Long-term Human Exposure to MEK 19
4.1.2.3. Occupational Studies of MEK Exposure 20
4.2. PRECHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN
ANIMALS-ORAL AND INHALATION 27
4.2.1. Oral Exposure 27
4.2.2. Inhalation Exposure 28
4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES 34
4.3.1. Studies in Humans 34
4.3.2. Studies in Animals 35
4.3.2.1. Oral Exposure 35
4.3.2.2. Inhalation Exposure 40
4.4. OTHER STUDIES 45
4.4.1. Acute Toxicity Data 45
4.4.1.1. Oral Exposure 45
4.4.1.2. Inhalation Exposure 47
4.4.2. Genotoxicity 47
4.4.3. Carcinogenicity 48
4.4.4. MEK Potentiation of Peripheral Neuropathy from Chemicals Metabolized to
Gamma-Diketones 49
iii
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4.5. SYNTHESIS AND EVALUATION OF MAJOR NONCANCER EFFECTS 52
4.5.1. Oral Exposure 52
4.5.2. Inhalation Exposure 54
4.5.3. Mode of Action Information 55
4.6. WEIGHT-OF-EVIDENCE EVALUATION AND CANCER
CHARACTERIZATION 55
4.7. SUSCEPTIBLE POPULATIONS AND LIFESTAGES 56
4.7.1. Possible Childhood Susceptibility 56
4.7.2. Possible Gender Differences 57
4.7.3. Other 57
5. DOSE RESPONSE ASSESSMENTS 59
5.1. ORAL REFERENCE DOSE (RfD) 59
5.1.1. Choice of Principal Study and Critical Effect 59
5.1.2. Methods of Analysis 60
5.1.2.1. Benchmark Dose Modeling 61
5.1.2.2. Route-to-route Extrapolation 69
5.1.3. RfD Derivation - Including Application of Uncertainty Factors 70
5.1.4. Previous Oral Assessment 72
5.2. INHALATION REFERENCE CONCENTRATION (RfC) 72
5.2.1. Choice of Principal Study and Critical Effect 72
5.2.2. Methods of Analysis 73
5.2.2.1. Benchmark Dose Modeling 74
5.2.2.2. Adjustment to a Human Equivalent Exposure Concentration 79
5.2.2.3. PBPK Modeling 81
5.2.3. RfC Derivation — Including Application of Uncertainty Factors 82
5.2.4. Previous Inhalation Assessment 84
5.3. CANCER ASSESSMENT 84
5.3.1. Oral Slope Factor 84
5.3.2. Inhalation Unit Risk 84
6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF
HAZARD AND DOSE RESPONSE 85
6.1. HUMAN HAZARD POTENTIAL 85
6.2. DOSE RESPONSE 87
6.2.1. Noncancer/Oral 87
6.2.2. Noncancer/Inhalation 87
6.2.3. Cancer/Oral and Inhalation 89
7. REFERENCES 90
APPENDIX A: SUMMARY OF EXTERNAL PEER REVIEW AND
PUBLIC COMMENTS AND DISPOSITION
APPENDIX B: BENCHMARK DOSE MODELING RESULTS AND OUTPUT
IV
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Output B-l: Reduced Pup Body Weight in Wistar Rats, Fl A Generation at
Postnatal Day 4 (Cox et al., 1975)
Output B-2: Reduced Pup Body Weight in Wistar Rats, Fl A Generation at
Postnatal Day 21 (Cox et al., 1975)
Output B-3: Reduced Fetal Weight in Wistar Rats, FIB Generation
(Coxetal., 1975)
Output B-4: Reduced Pup Body Weight in Wistar Rats, F2 Generation at
Postnatal Day 4 (Cox et al., 1975)
Output B-5: Reduced Pup Body Weight in Wistar Rats, F2 Generation at
Postnatal Day 21 (Cox et al., 1975)
Output B-6: Increased Incidence of Extra Ribs in Sprague-Dawley Rats
(Deacon et al., 1981)
Output B-7: Reduced Fetal Weight in CD-I Mice (Schwetz et al., 1991/
Mastetal., 1989)
Output B-8:, Increased Incidence of Misaligned Sternebrae in CD-I Mice
(Schwetz et al., 1991/Mast et al., 1989)
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LIST OF TABLES
Table 1. Kinetic parameters used for PBPK models for MEK kinetics in humans and rats ... 16
Table 2. Mean Fl A litter body weight on days 4 and 21 in rats exposed to 2-butanol in drinking
water 37
Table 3. Incidence of skeletal variations in FIB fetuses 38
Table 4. Maternal and fetal effects in 2-butanol-exposed rats 44
Table 5. Summary of key repeat-exposure reproductive and developmental toxicity studies in
animals exposed to MEK or 2-butanol 62
Table 6. Mean litter pup body weight in F1A generation Wistar rats exposed to 2-butanol in
drinking water in a two-generation reproductive and developmental toxicity study 65
Table 7. Litter mean fetal weight in FIB generation Wistar rats exposed to 2-butanol in drinking
water in a two-generation reproductive and developmental toxicity study 66
Table 8. Benchmark dose modeling results using litter mean body weight data for
FIB fetuses 66
Table 9. Litter mean pup body weight in F2 generation Wistar rats exposed to 2-butanol in
drinking water in a two-generation reproductive and developmental toxicity study 67
Table 10. Benchmark dose modeling results using litter mean body weight data for F2 pups on
postnatal days 4 and 21 68
Table 11. Benchmark doses for developmental effects in rats from various generations and
potential points of departure for the RfD 69
Table 12. Incidence of extra ribs (litters with an affected fetus) in Sprague-Dawley rats exposed
to MEK 7 hours/day on gestation days 6-15 74
Table 13. Benchmark concentration modeling results using litter incidence data for Sprague-
Dawley rat fetuses with extra ribs exposed to MEK during gestation days 6-15 75
Table 14. Litter mean fetal weight (both sexes combined) in CD-I mice exposed to MEK 7
hours/day on gestation days 6-15 76
Table 15. Benchmark concentration modeling results using litter mean body weight data .... 76
Table 16. Total number of fetuses (combined for both sexes) with misaligned
sternebrae per exposure group in CD-I mice exposed to MEK 7 hours/day
on gestation days 6-15 77
VI
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Table 17. Benchmark concentration modeling results using individual litter data
for mouse fetuses with misaligned sternebrae exposed to MEK during gestation
days 6-15 (without litter size as covariates) 78
Table 18. Benchmark concentrations for developmental effects in mice and rats
and potential points of departure for the RfC 79
LIST OF FIGURES
Figure 1. Proposed pathways for methyl ethyl ketone metabolism 9
Figure 2. Comparison of fetal body weight changes in animals exposed to MEK or 2-butanol
during gestation 46
vn
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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 methyl
ethyl ketone. It is not intended to be a comprehensive treatise on the chemical or toxicological
nature of methyl ethyl ketone.
In Section 6, EPA has characterized its overall confidence in the quantitative and
qualitative aspects of hazard and dose response. Matters considered in this characterization
include knowledge gaps, uncertainties, quality of data, and scientific controversies. 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.
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.
Vlll
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AUTHORS, CONTRIBUTORS, AND REVIEWERS
CHEMICAL MANAGER
Susan Rieth
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
AUTHORS
Susan Rieth
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Karen Hogan
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Mark H. Follansbee, Ph.D.
Syracuse Research Corporation
Scarborough, ME
Peter McClure, Ph.D., DABT
Syracuse Research Corporation
North Syracuse, NY
Regina McCartney
Syracuse Research Corporation
Cincinnati, OH
Syracuse Research Corporation staff performed work under Contract No. 68-C-00-122,
Work Assignment 2-06.
REVIEWERS
This document and summary information on IRIS have received peer review both by
EPA scientists and by independent scientists external to EPA. Subsequent to external review
and incorporation of comments, this assessment has undergone an Agency-wide review process
whereby the IRIS Program Director has achieved a consensus approval among the Office of
Research and Development; Office of Air and Radiation; Office of Prevention, Pesticides, and
IX
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Toxic Substances; Office of Solid Waste and Emergency Response; Office of Water; Office of
Policy, Economics, and Innovation; Office of Children's Health Protection; Office of
Environmental Information; and the Regional Offices.
INTERNAL EPA REVIEWERS
Katherine Anitole, Ph.D.
Office of Pollution Prevention and Toxics
U.S. Environmental Protection Agency
Washington, DC
Daniel Axelrad, M.P.P.
Office of Policy, Economics, and Innovation
National Center for Environmental Economics
U.S. Environmental Protection Agency
Washington, DC
Philip Bushnell, Ph.D.
Office of Research and Development
National Health and Environmental Effect Research Laboratory
U.S. Environmental Protection Agency
Research Triangle Park, NC
Audrey Cummings, Ph.D.
Office of Research and Development
National Health and Environmental Effect Research Laboratory
U.S. Environmental Protection Agency
Research Triangle Park, NC
Robert Dewoskin, Ph.D.
Office of Research and Development
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Research Triangle Park, NC
Gary Foureman, Ph.D.
Office of Research and Development
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Research Triangle Park, NC
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David Herr, Ph.D.
Office of Research and Development
National Health and Environmental Effect Research Laboratory
U.S. Environmental Protection Agency
Research Triangle Park, NC
Tracey Woodruff, Ph.D., M.P.H.
Office of Policy, Economics, and Innovation
National Center for Environmental Economics
U.S. Environmental Protection Agency
San Francisco, CA
EXTERNAL PEER REVIEWERS
Bryan Hardin, Ph.D.
Global Tox
Assistant Surgeon General, U.S. Public Health Service (retired)
Dale Hattis, Ph.D.
Clark University
Worcester, MA
Arthur R. Gregory, Ph.D.
Techto Enterprises
Luray, VA
Rochelle W. Tyl, Ph.D.
Research Triangle Institute
Research Triangle Park, NC
Summaries of the external peer reviewers' comments, public comments and the
disposition of their recommendations are in Appendix A.
XI
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1. INTRODUCTION
This document presents background and justification for the hazard and dose-response
assessment summaries in EPA's Integrated Risk Information System (IRIS). IRIS Summaries
may include an oral reference dose (RfD), inhalation reference concentration (RfC) and a
carcinogenicity assessment.
The RfD and RfC provide quantitative information for noncancer dose-response
assessments. The RfD is based on the assumption that thresholds exist for certain toxic effects
such as cellular necrosis but may not exist for other toxic effects such as some carcinogenic
responses. It is expressed in units of mg/kg-day. In general, the RfD is an estimate (with
uncertainty spanning perhaps an order of magnitude) of a daily exposure to the human
population (including sensitive subgroups) that is likely to be without an appreciable risk of
deleterious noncancer effects during a lifetime. The inhalation RfC is analogous to the oral RfD,
but provides a continuous inhalation exposure estimate. The inhalation RfC considers toxic
effects for both the respiratory system (portal-of-entry) and for effects peripheral to the
respiratory system (extrarespiratory or systemic effects). It is generally expressed in units of
mg/m3.
The carcinogenicity assessment provides information on the carcinogenic hazard
potential of the substance in question and quantitative estimates of risk from oral exposure and
inhalation exposure. 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 are presented in three ways. The slope factor is
the result of application of a low-dose extrapolation procedure and is presented as the risk per
mg/kg-day. The unit risk is the quantitative estimate in terms of either risk per |ig/L drinking
water or risk per |ig/m3 air breathed. Another form in which risk is presented is a drinking water
or air concentration providing cancer risks of 1 in 10,000; 1 in 100,000; or 1 in 1,000,000.
Development of these hazard identification and dose-response assessments for methyl
ethyl ketone has followed the general guidelines for risk assessment as set forth by the National
Research Council (1983). EPA guidelines that were used in the development of this assessment
may 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), Guidelines
for Developmental Toxicity Risk Assessment (U.S. EPA, 199 la), Guidelines for Reproductive
Toxicity Risk Assessment (U.S. EPA, 1996), Guidelines for Neurotoxicity Risk Assessment (U.S.
1
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EPA, 1998a), Draft Revised Guidelines for Carcinogen Assessment (U.S. EPA, 1999),
Recommendations for and Documentation of Biological Values for Use in Risk Assessment (U.S.
EPA, 1988), (proposed) 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), Science Policy Council
Handbook: Peer Review (U.S. EPA, 1998b, 2000a), Science Policy Council Handbook: Risk
Characterization (U.S. EPA, 2000b), Benchmark Dose Technical Guidance Document (U.S.
EPA, 2000c) and Supplementary Guidance for Conducting Health Risk Assessment of Chemical
Mixtures (U.S. EPA 2000d).
The literature search strategy employed for this compound was based on the CASRN and
at least one common name. At a minimum, the following data bases were searched: RTECS,
HSDB, TSCATS, CCRIS, GENE-TOX, DART/ETIC, EMIC, TOXLINE, CANCERLIT, and
MEDLINE. For this toxicological review, updated literature searches for 1987 to July 2003
were conducted for MEK. Literature searches were also conducted from 1991 to July 2003 for
2-butanol and from 1965 to July 2003 for 3-hydroxy-2-butanone and 2,3-butanediol. Any
pertinent scientific information submitted by the public to the IRIS Submission Desk was also
considered in the development of this document.
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2. CHEMICAL AND PHYSICAL INFORMATION RELEVANT TO ASSESSMENTS
Methyl ethyl ketone (MEK) is also known as 2-butanone, butanone, ethyl methyl ketone,
and methyl acetone. Some relevant physical and chemical properties of MEK are listed below
(ATSDR, 1992; CRC, 1994; HSDB, 1999; NTP, 2002):
CAS registry number: 78-93-3
Chemical formula: C4H8O
Molecular weight: 72.11
Density: 0.805 g/mL @ 20°C
Vapor pressure: 77.5 mm Hg @ 20°C
Water solubility: 275 mg/mL @ 20°C
Conversion factor: 1 ppm = 2.95 mg/m3, 1 mg/m3 = 0.340 ppm @ 25°C, 760 mm Hg
At room temperature, MEK is a clear liquid with a fragrant mint-like odor. It is
flammable, with a flash point of-3°C. MEK is strongly reactive with a number of chemicals and
chemical classes, including potassium tert-butoxide, chloroform, hydrogen peroxide, and strong
oxidizers (e.g., chlorosulfonic, sulphuric, and nitric acids). It can also react with bases and
strong reducing agents. Vigorous reactions occur with chloroform in the presence of bases, and
explosive peroxides are formed when added to hydrogen peroxide and nitric acid. ACGIH
(2001) recommends an 8-hour time-weighted average threshold limit value (TWA-TLV) of 200
ppm (590 mg/m3) MEK. Similarly, the Occupational Safety and Health Administration (OSHA)
has promulgated an 8-hour permissible exposure limit (PEL) of 200 ppm (590 mg/m3) MEK
(OSHA, 1993).
MEK is used as a solvent in the application of protective coatings (varnishes) and
adhesives (glues and cements), in magnetic tape production, in smokeless powder manufacture,
in the dewaxing of lubricating oil, in vinyl film manufacture, and in food processing. Its use as a
component in adhesives used to join PVC pipes is a potential route for entry of the chemical into
potable water (ATSDR, 1992). It is also commonly used in paint removers, cleaning fluids,
acrylic coatings, pharmaceutical production, and colorless synthetic resins, and as a printing
catalyst and carrier (Merck Index, 2001). MEK may be found in soil and water in the vicinity of
some hazardous waste sites. MEK has been detected as a natural component of numerous foods,
including: raw chicken breast, milk, nuts (roasted filberts), cheese (Beaufort, Gruyere, and
cheddar), bread dough and nectarines at concentrations ranging from 0.3 to 19 ppm (ATSDR,
1992; HSDB, 1999; WHO, 1992). MEK is also found in tobacco smoke and volatile releases
from building materials and consumer products (ATSDR, 1992). WHO (1992) estimated levels
of daily MEK intake from different sources as follows: foodstuffs - 1,590 |ig/day; drinking
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water (2 liters) - 3.2 jig/day; rural outdoor air - 36 jig/day; urban outdoor air <760 jig/day; and
tobacco smoke < 1,620 ng/day.
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3. TOXICOKINETICS RELEVANT TO ASSESSMENTS
3.1. ABSORPTION
3.1.1. Oral Exposure
Case reports provide qualitative evidence that MEK is absorbed by the gastrointestinal
tract following oral exposure in humans; however, they do not provide information regarding the
extent of absorption following ingestion. For example, a woman accidentally ingested an
unknown quantity of MEK and presented with symptoms of metabolic acidosis and a blood
concentration of 95 mg/100 mL (13.2 mM) MEK (Kopelman and Kalfayan, 1983). A man who
intentionally ingested 100 mL of liquid cement containing a mixture of acetone (18%), MEK
(28% or about 37 mg/kg), and cyclohexanone (39%) was treated by gastric lavage 2 hours after
ingestion. Three hours later, he had a plasma level of about 110 |ig/mL MEK (Sakata et al.,
1989).
Experimental data from rodents indicate that orally administered MEK is absorbed from
the gastrointestinal tract and rapidly eliminated. Oral administration (gavage) of 1,690 mg/kg of
MEK to four male Sprague-Dawley rats resulted in a mean peak plasma concentration of 94.1
mg/100 mL after 4 hours that decreased to 6.2 mg/100 mL 18 hours after exposure (Dietz and
Traiger, 1979; Dietz et al., 1981). Thrall et al. (2002) reported mean peak concentrations in
exhaled air 1 hour after an oral gavage dose of 50 mg/kg MEK to three male F344 rats, providing
further support that MEK is absorbed from the digestive tract.
3.1.2. Inhalation Exposure
Data from humans and rats suggest that MEK is well absorbed during inhalation
exposure due to its high blood/air solubility ratio (Perbellini et al., 1984; Sato and Nakajima,
1979; Thrall et al., 2002). Perbellini et al. (1984) investigated the uptake and kinetics of MEK in
groups of industrial workers occupationally exposed to MEK. In one group, the concentration of
MEK in environmental air was compared to MEK in the alveolar air of exposed workers (n = 82)
by simultaneous collection of air samples into glass tubes via instantaneous sampling methods
and gas chromatography (GC) analysis. Most of the measurements were made at environmental
concentrations at or below 100 ppm. The alveolar air concentration of MEK in the exposed
workers was highly correlated with the environmental air concentration and averaged 30% of the
latter. From these survey results, the investigators estimated a pulmonary retention of 70% in
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workers exposed to concentrations less than 300 ppm for 4 hours. Perbellini et al. (1984)
presented a physiologically-based mathematical model for MEK that suggests that steady-state
concentrations are reached within 8 hours when exposures are between 50 and 100 ppm,
depending on the physical work load. In a controlled exposure experiment, pulmonary uptake in
volunteers ranged from 51 to 55% of the inspired quantity at 200 ppm MEK for 4 hours in an
exposure chamber (Liira et al., 1988). Liira et al. (1990a) found the pulmonary retention of
MEK in five human volunteers similarly exposed to MEK to be 55.8 ± 9.1%. Exercise increased
the pulmonary uptake of MEK due to the greater ventilatory rate (Liira et al., 1988). Liira et al.
(1990b) and Imbriani et al. (1989) reported that human inhalation exposure to MEK exhibited
dose-dependent saturation. Dick et al. (1988) exposed 24 volunteers (12 men and 12 women) to
MEK at 200 ppm for 4 hours and reported that alveolar breath samples (exhaled air) reached
steady-state concentrations by 2 hours, stabilizing at 5-6% of the exposure concentration. There
is no apparent explanation for the much lower pulmonary retention reported by Dick et al.
(1988) as compared to Liira et al. (1988, 1990a).
Kessler et al. (1988) reported a pulmonary retention of 40% for rats exposed to
concentrations less than or equal to 180 ppm for up to 14 hours.
3.1.3. Dermal Exposure
The percutaneous absorption of MEK appears to be rapid (Munies and Wurster, 1965;
Wurster and Munies, 1965). These authors reported that MEK was present in the exhaled air of
human subjects within 2.5-3.0 minutes after application to normal skin of the forearm, and the
concentration of MEK in exhaled air reached a plateau in approximately 2 hours. The rate of
absorption was slower when MEK was applied to dry skin, where a plateau for the concentration
of MEK in expired air was attained in 4-5 hours. By contrast, absorption of MEK to moist skin
was very rapid. MEK was detected in expired air in measurable concentrations within 30
seconds after application of MEK to the skin of the forearm, and a maximum concentration in
expired air was achieved in 10-15 minutes, decreasing thereafter. Munies and Wurster (1965)
concluded that the rapid percutaneous absorption of MEK is related to its olive oil-water
partition coefficient of 0.93, as reported by GC analysis.
The percutaneous absorption data of Munies and Wurster (1965) have been used to
calculate the following minimum rates of percutaneous penetration of MEK: 0.46 |ig/cm2/minute
for dry or normal skin and 0.59 |ig/cm2/minute for moist skin (JRB Associates, 1980 as cited in
WHO, 1992). Ursin et al. (1995) also studied the in vitro permeability of MEK through living
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human skin. Ursin et al. (1995) measured the permeability of various solvents, including MEK,
through a 0.64 cm2 sample of living skin tissue separating a two-chamber diffusion cell. All skin
samples were first calibrated for relative permeability using tritiated water. The authors
concluded that MEK has a permeability rate of 53±29 g/m2/hour, which is equivalent to
approximately 0.0066 cm/hour (Ursin et al., 1995) or approximately 88.3 |ig/cm2/minute [53
g/m2-hour) x (1 hour/60 minute) x (100 |ig/l g) x (1 m2/104 cm2) = 88.3 |ig/cm2-minute]. The
permeability absorption values from these studies differ by 2 orders of magnitude. The values
reported by Munies and Wurster (1965) may be low because the analysis was based solely on the
amount of MEK exhaled from the lungs, thereby not considering all routes of MEK elimination
(WHO, 1992).
Brooke et al. (1998) studied the dermal uptake of MEK from the vapor phase. Groups of
four volunteers were exposed for 4 hours to MEK in an inhalation chamber either 'whole body'
or via the 'skin only' at 200 ppm MEK. For skin-only exposures, volunteers wore air masks that
delivered room air. Uptake was assessed by monitoring levels of MEK in blood, single breath,
and urine following exposure. Brooke et al. (1998) reported that dermal absorption of MEK
contributed approximately 3-3.5% of the total body burden.
3.2. DISTRIBUTION
No studies were located regarding the distribution of MEK following oral or dermal
exposure in humans or animals. In a study of MEK-exposed industrial workers (n = 23),
Perbellini et al. (1984) compared the concentration of MEK in venous blood to alveolar air.
Samples were collected simultaneously toward the end of the work shift and analyzed by gas
chromatography-mass spectrometry (GC/MS). The level of MEK in the blood was significantly
correlated with the environmental concentration, indicating rapid transfer from the lungs to the
blood. Information on the distribution of MEK following inhalation exposure in humans also
comes from an examination of postmortem tissues reported by Perbellini et al. (1984). The
distribution of MEK in human tissues was examined in two solvent-exposed workers who died
suddenly of heart attacks at the workplace (Perbellini et al., 1984). Postmortem determinations
of the MEK tissue/air solubility ratio for human kidney, liver, muscle, lung, heart, fat, and brain
revealed similar solubility in all these tissues, with the tissue/air ratio ranging from 147 (lung) to
254 (heart) (Perbellini et al., 1984). The available data suggest that MEK does not accumulate in
fatty tissues in humans. Blood/tissue solubility ratios for several tissues approach unity
(Perbellini et al., 1984). Since the results have also been repeated in rats (Thrall et al., 2002),
7
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MEK is not expected to accumulate in any particular tissue (Perbellini et al., 1984).
3.3. METABOLISM
The available evidence indicates that the metabolism of MEK is similar in humans and
experimental animals. As shown in Figure 1, the majority of MEK is metabolized to 3-hydroxy-
2-butanone, which is subsequently metabolized to 2,3-butanediol. A small portion is reversibly
converted to 2-butanol. Evidence supporting common metabolic pathways for MEK in humans
and experimental animals is presented below.
In humans exposed to airborne MEK, 2-butanol and 2,3-butanediol have been identified
as MEK metabolites in serum, while 3-hydroxy-2-butanone and 2,3-butanediol have been
identified as urinary metabolites of MEK (Perbellini et al., 1984; Liira et al., 1988, 1990a).
From a study of the kinetics of inhaled MEK in human volunteers (200 ppm for 4 hours), it was
estimated that 3% of the absorbed dose was exhaled as unchanged MEK, 2% of the absorbed
dose was excreted in urine as 2,3-butanediol, and the remainder of the absorbed dose entered
into mainstream intermediary metabolism and was transformed to simple compounds such as
carbon dioxide and water (Liira et al., 1988). Results from this study suggest that MEK is
rapidly and nearly completely metabolized in humans exposed to 200 ppm MEK for 4 hours.
In humans, MEK has also been identified as a minor but normal constituent of urine, as a
constituent in the serum and urine of diabetics, and in expired air. Its production in the body has
been attributed to isoleucine catabolism (WHO, 1992). MEK was detected in the blood of more
than 75% of the participants of the general population in the Third National Health and Nutrition
Examination Survey (NHANES III) (Ashley et al., 1994; Churchill et al., 2001). Median blood
levels were 5.4 ppb. Investigators looked for associations between MEK blood levels and self-
reported chemical exposures as collected via NHANES questionnaire. Blood MEK levels were
positively associated with mean daily alcohol intake, and were generally not associated with
other environmental exposure variables (Churchill et al., 2001).
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o
I-LC—C C-
l I
I 11-
MEK
(reduction)
OH
hLC—C C—CH,
3 H H2
2-butanol
-CH,
\
(oxidation)
O
X— C - C - ChL
3 s
3-hydroxy-2-butanone
(reduction)
OH OH
H3C—CH CH CH3
2,3-butanediol
Figure 1. Proposed pathways for methyl ethyl ketone metabolism
Source: Adapted from DiVincenzo et al. (1976).
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In rats and guinea pigs, the metabolism of MEK may follow one of two pathways (Dietz
et al., 1981; DiVincenzo et al., 1976). The majority of MEK is oxidized by the cytochrome P450
monooxygenase system (P450IIE1 and IIB isozymes) to the primary metabolite,
3-hydroxy-2-butanone (3H-2B), which is subsequently reduced to 2,3-butanediol (2,3-BD)
(Dietz and Traiger, 1979; Traiger et al., 1989; Brady et al., 1989; Raunio et al., 1990). A small
portion of absorbed MEK is reduced to 2-butanol, which is rapidly oxidized back to MEK.
Based on the data from Traiger and Bruckner (1976), Dietz et al. (1981) established that
approximately 96% of an administered oral dose of 2-butanol is oxidized in vivo to MEK within
16 hours of oral administration. Dietz et al. (1981) reported that no significant difference in area
under the curve (AUC) of MEK blood concentration was observed after oral dosing of rats with
either 1,776 mg/kg 2-butanol or 1,690 mg/kg MEK (10,899±842 or 9,868±566 mg-hour/liter,
respectively).1 Peak concentrations of MEK and its downstream metabolites were similar
whether MEK or 2-butanol were administered (Dietz et al., 1981), with a shift of approximately
4 hours to reach peak concentrations when MEK was administered:
Peak Blood Concentration
MEK
3H-2B
2,3-BD
Administration of
1,776 mg/kg 2-butanol
0.78 mg/ml at 8 hr
0.04 mg/ml at 12 hr
0.21 mg/ml at 18hr
Administration of
1,690 mg/kg MEK
0.95 mg/ml at 4 hr
0.027 mg/ml at 8 hr
0.26 mg/ml at 18hr
Dietz et al. (1981) provides further support for the rapid conversion of orally administered 2-
butanol to MEK. Ultimately, 2-butanol and MEK are metabolized through the same
intermediates as shown in Figure 1.
DiVincenzo et al. (1976) identified the metabolites of aliphatic ketones in the serum of
guinea pigs after administering a single dose of methyl n-butyl ketone, methyl isobutyl ketone,
or MEK. The hepatic cytochrome P450-mediated metabolism of MEK (Figure 1) produced
hydroxylated metabolites (3-hydroxy-2-butanone and 2,3-butanediol) that were eliminated in the
urine (DiVincenzo et al., 1976). Male Sprague-Dawley rats given a single oral dose of MEK at
1,690 mg/kg exhibited blood concentrations of MEK and metabolites 4 hours after dosing as
1 On a molar basis, the administered concentrations of 2-butanol and MEK are 0.024 and
0.023 mol/kg, respectively.
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follows: MEK (94.1 mg/100 mL), 2-butanol (3.2 mg/100 mL), 3-hydroxy-2-butanone (2.4
mg/100 mL), and 2,3-butanediol (8.1 mg/100 mL) (Dietz and Traiger, 1979; Dietz et al., 1981).
After 18 hours, blood concentrations of the parent compound and metabolites were: MEK (6.2
mg/100 mL), 2-butanol (0.6 mg/100 mL), 3-hydroxy-2-butanone (1.4 mg/100 mL), and
2,3-butanediol (25.6 mg/100 mL) (Dietz and Traiger, 1979).
Interestingly, the data of Dietz et al. (1981) demonstrated a peak blood concentration of
MEK approximately 4 hours after oral administration of Sprague-Dawley rats to 1,690
mg/kg MEK, while Thrall et al. (2002) found peak concentrations in exhaled air at 1 hour after
oral gavage of 50 mg/kg MEK to F344 rats. Thrall et al. (2002) concluded that the differences in
MEK dose level (approximately 35-fold), rat strain used, and overnight fasting may explain the
discrepancy between these findings.
Gadberry and Carlson (1994) showed that the in vitro hepatic oxidation of 2-butanol to
MEK is inducible by pretreatment with ethanol (an inducer of P450IIE1) and phenobarbital (an
inducer of P450IIB and IVB), but not beta-naphthaflavone (an inducer of P450IA1). By
contrast, in vitro studies showed that 2-butanol oxidation in the lung was not inducible by any of
the treatments. A daily dose of 1.4 mL/kg MEK for 3 days increased the amounts of ethanol-
and phenobarbital-inducible cytochrome P450 isoforms (P450IIE1 and P450IIB) as
demonstrated by in vitro assays (Raunio et al., 1990). Because MEK is an inducer of
microsomal P450 activity, repeated MEK exposure may enhance the body's capacity for
metabolism of subsequent exposures.
3.4. ELIMINATION AND EXCRETION
In human studies involving acute inhalation exposure, the urinary excretion of MEK and
metabolites and the exhalation of unchanged MEK have been estimated to account for only a
small percentage (0.1-3%) of the absorbed dose (Perbellini et al., 1984; Liira et al., 1988). The
remainder of the absorbed dose is expected to have undergone rapid transformation to carbon
dioxide and water through intermediary metabolic pathways (Liira et al., 1988). Nevertheless,
the presence of unchanged MEK in urine has been proposed as a marker of exposure since strong
positive correlations have been reported between MEK levels in urine and MEK levels in air
(Perbellini et al., 1984; Liira et al., 1988; Imbriani et al., 1989; Sia et al., 1991; ACGffl, 2001).
11
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MEK is rapidly cleared from the blood with a reported plasma half-life in humans of
49-96 minutes, exhibiting a biphasic elimination: t1/2 alpha = 30 minutes and t1/2 beta =81
minutes (Liira et al., 1988). Dick et al. (1988) collected blood samples from 20 volunteers (sex
not specified) who were exposed to 100 or 200 ppm MEK for 4 hours. Blood samples were
obtained from each subject at 2 and 4 hours from the start of exposure and 15 and 20 hours post
exposure. Assuming first-order kinetics, Dick et al. (1988) estimated an elimination half-life of
49 minutes for MEK. MEK was not detected in blood at 20 hours post exposure. Given the
rapid clearance of MEK demonstrated by Liira et al. (1990b) and Dick et al. (1988), it is unlikely
that MEK would accumulate with chronic exposure.
Based on the strong correlation between urinary MEK concentration and environmental
exposure, a biological exposure index of 2 mg/L MEK in urine measured at the end of the work
shift has been adopted to monitor occupational exposure to MEK (ACGIH, 2001).
3.5. PHYSIOLOGICALLY-BASED PHARMACOKINETIC (PBPK) MODELS
Physiologically-based pharmacokinetic (PBPK) models of MEK are available for humans
(Liira et al., 1990b; Leung, 1992) and rats (Dietz et al., 1981; Thrall et al., 2002). PBPK models
are unavailable for other species. The structural differences and limited data sets used to
calibrate and test the rat and human models limits their application. The human PBPK model
(Liira et al., 1990b; Leung, 1992) was developed to describe the dose-dependent elimination
kinetics of MEK in humans following inhalation exposure to low concentrations of MEK. Liira
et al. (1990b) exposed two men in an inhalation chamber for 4 hours in separate exposures to 25,
200, or 400 ppm MEK. Venous blood samples were taken during each exposure and for 8 hours
thereafter. The metabolism of MEK was assumed to occur only in the liver and was described
by Michaelis-Menten kinetics. The model, which is based on the spreadsheet model of Johanson
and Naslund (1988), contained eight compartments describing the kinetics of MEK in lungs, GI
tract, liver, richly perfused tissue, poorly perfused tissue, fat, muscle, and blood (see Table 1 for
12
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model parameters). The elimination rate for MEK was calculated by the following equation:
elimination rate = Vmax x Ch/(Km+Ch)
where:
Ch = MEK concentration in hepatic venous blood;
Vmax =30 (imol/minute (obtained by applying best fit of simulated curves to experimental MEK
blood concentration); and
Km = 2 jiM (obtained by applying best fit of simulated curves to experimental MEK blood
concentration).
Liira et al. (1990b) reported that model predictions were similar to observed blood
concentrations of MEK in 17 male volunteers exposed to 200 ppm. The authors also concluded
that the kinetic constants were fairly representative of healthy male subjects.
Research utilizing rats (Dietz and Traiger, 1979; Dietz et al., 1981) identified the
pathways of MEK metabolism and permitted a calculation of rate constants for the elimination of
MEK and its metabolites from the blood as well as for the metabolic transformations. The data
were used as the basis for a PBPK model for MEK (Dietz et al., 1981) to predict blood
concentrations of 2-butanol and its metabolites. More specifically, the model was used to predict
concentrations of MEK (i.e., 2-butanone), 3-hydroxy-2-butanone, and 2,3-butanediol in Sprague-
Dawley rats after oral administration of 2-butanol or MEK, as well as after intravenous
administration of 3-hydroxy-2-butanone or 2,3-butanediol.
The model contains two compartments (in the blood and the liver) where metabolism
occurs. The differential equations are based upon a perfusion-limited model, and account for:
(1) the elimination of 2-butanol and its metabolites from the blood at rates linearly proportional
to blood concentrations, (2) transport between the blood and liver compartments, and (3)
metabolic conversions in the liver. Metabolic conversions were described with Michaelis-
Menten saturation kinetics and included rates for bidirectional conversions between 2-butanol
and MEK, unidirectional conversion of MEK to 3-hydroxy-2-butanone, and bidirectional
conversions between 3-hydroxy-2-butanone and 2,3-butanediol. Kinetic constants in the model
were estimated by successive curve fitting of submodels to in vivo blood concentration data from
groups of 5 rats following: (1) a single gavage administration of 1,690 mg/kg MEK, (2) a single
gavage administration of 1,776 mg/kg 2-butanol, and (3) intravenous injections of 3-hydroxy-2-
butanone and 2,3-butanediol at 400 or 800 mg/kg. Equations describing the metabolic
conversion of MEK to 3-hydroxy-2-butanone included a competitive inhibition of its conversion
13
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that was attributed to the presence of the competitive substrate, 2-butanol. In addition, a
distribution coefficient was included to account for the unexpectedly low observed concentration
of 3-hydroxy-2-butanone in the blood. The authors hypothesized that this was due to
partitioning, binding, or altered transport rates from the liver. The "adjustments" resulted in an
improved fit between model simulations and experimentally observed blood concentrations of
MEK and 2-butanol following oral administration of 1,690 mg/kg MEK, but predicted blood
concentrations of 3-hydroxy-2-butanone and 2,3-butanediol were about 20-30% lower than the
observed values.2 There were no comparisons reported for model predictions with data not used
to derive the model parameters.
Thrall et al. (2002) developed a PBPK model for MEK in F344 rats, from experimentally
determined partition coefficients using in vitro vial equilibration technique and in vivo
measurements of MEK uptake in rats exposed to 100 to 2,000 ppm MEK in a closed,
recirculating gas uptake system. The model included both a saturable metabolic pathway
described by Michaelis-Menten kinetic constants and a nonsaturable first-order pathway. The
model provided adequate predictions (based on visual inspection) of exhaled MEK
concentrations following inhalation, intravenous, intraperitoneal, or oral administration of MEK
to rats. One notable difference between the Thrall et al. (2002) and Dietz et al. (1981) models is
the peak exhaled breath concentrations following oral gavage. Dietz et al. (1981) found peak
MEK concentrations in blood 4 hours after oral gavage (1,690 mg/kg MEK), whereas the Thrall
et al. (2002) study found peak MEK concentrations in exhaled air 1 hour after oral gavage (50
mg/kg MEK).
The Thrall et al. (2002) model could be extended to humans by substituting human
parameter values for rat parameter values. Use of such a model for risk assessment purposes
would still be dependent upon sufficient validation or comparisons of model predictions with
relevant human data. This has not been carried out to date.
In summary, three PBPK models have been developed based on a limited number of data
sets in rats and humans. The predictive capabilities of these models have not been adequately
tested, and none of the models were parameterized for rats and humans to sufficiently support an
2 The model parameters for the Dietz et al. (1981) model are not provided in Table 1
because relatively few of the values were provided by the authors. The rate constants that were
provided were not readily interpretable in the framework shown in Table 1. The physiological
constants appropriate for converting the available parameters to reflect the equivalent framework
were also not available.
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extrapolation of rat dose-response data to humans based upon an equivalent internal human dose
metric. Data to support the use of the PBPK models for route-to-route extrapolation are also
limited or not available.
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Table 1. Kinetic parameters used for PBPK models for MEK kinetics in humans and rats
Parameter
Body weight (kg)
Blood flows to tissues at
rest, L/min and (% of
cardiac output)3
Tissue volume, L and (%
body weight)
Tissue/air partition
coefficient13
Ventilation at rest (L/hr)
Hepatic metabolism0
Tissue/Kinetic
Parameter
Lungs
GI tract
Liver
Richly perfused tissues
Poorly perfused tissues
Fat
Muscle
Lungs
GI tract
Liver
Richly perfused tissues
Poorly perfused tissues
Fat
Muscle
Lungs
GI tract
Liver
Richly perfused tissues
Poorly perfused tissues
Fat
Muscle
Blood
Alveolar
Pulmonary
Vmax, mg/h-kg
Km, mg/L
First-order rate constant
(h-1)
Human model,
Liira et al. (1990b)
70
5.05
1.2 (21)
1.6 (28)
2.1 (37)
0.1 (2)
0.25 (4)
0.5 (9)
2.0
2.4
1.5
2.1
12.5
14.5
16.5
103
107
107
107
107
162
103
125
403
672
1.85
0.14
_
Rat model,
Thrall et al.
(2002)
0.25
d
-
(25)
(51)
(20)
(4)
-
-
-
(4)
(5)
(74)
(8)
-
_
-
152
-
-
101
185
138.5
5.4
-
5.44
0.63
4.1
a Cardiac output for humans taken to be the total of the blood flows, or 5.75 L/min.
b Tissue air partition coefficient as reported by the autopsy study by Fiserova-Bergerova and Diaz (1986).
0 Human metabolic parameters were reported by Liira et al. (1990b) as Vmax=30 ^mol/minute and K,,, =2 ^
d Parameter not used in model or not reported.
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4. HAZARD IDENTIFICATION
4.1. STUDIES IN HUMANS - EPIDEMIOLOGY, CASE REPORTS, CLINICAL
CONTROLS
All dose conversions made in this chapter are made assuming conditions of standard
temperature and pressure.
4.1.1. Oral Exposure
Kopelman and Kalfayan (1983) described a case report of nonoccupational, acute toxicity
from ingestion of MEK. A 47-year-old woman who inadvertently ingested an unknown amount
of MEK was unconscious, hyperventilating, and suffering from severe metabolic acidosis upon
hospital admission. Her plasma concentration of MEK was 950 mg/L. After a complete and
uneventful recovery, she was discharged from the hospital.
4.1.2. Inhalation Exposure
4.1.2.1. Acute Exposure
As with other small molecular weight, aliphatic, or aromatic organic chemicals used as
solvents (e.g., acetone or toluene), acute inhalation exposure to high concentrations of MEK
vapors is expected to cause reversible central nervous system depression; however, evidence for
such effects in humans is limited to a single case report (Welch et al., 1991). In an extensive
series of studies involving 4-hour exposure of human subjects to 200 ppm (590 mg/m3) MEK,
National Institute for Occupational Safety and Health (NIOSH) investigators found no
statistically significant increase in reported symptoms of throat irritation, nor did they find
marked performance changes in a series of tests of psychomotor abilities, postural sway, and
moods (Dick et al., 1984, 1988, 1989, 1992).
Welch et al. (1991) reported that a 38-year-old male worker exposed to paint base
containing MEK and toluene in an enclosed, unventilated garage exhibited neurological
symptoms. Exposure occurred at an unknown concentration of MEK for an acute, but
unspecified, period of time. Initial symptoms included nausea, headache, dizziness, and
respiratory distress. Over the next several days, the subject experienced impaired concentration,
memory loss, tremor, gait ataxia, and dysarthria. Subsequent MRI evaluation revealed fluid
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accumulation in the left parietal area. The condition was diagnosed as toxic encephalopathy
with dementia and cerebellar ataxia. Some neurological deficits persisted for more than 30
months following the acute exposure. It is not clear from this report whether the central nervous
symptom effects were due to exposure to MEK, toluene, or a combination of solvents.
In a series of studies by NIOSH investigators (Dick et al., 1984, 1988, 1989), volunteers
(male and female) underwent a single 4-hour exposure to 200 ppm (590 mg/m3) MEK, after
which the following neurobehavioral tests were conducted: psychomotor tests (choice reaction
time, visual vigilance, dual task, and memory scanning), postural sway, and a profile of mood
states. No statistically significant changes in neurobehavioral performance were observed (Dick
et al., 1984, 1988, 1989). Dick et al. (1984, 1988) evaluated the performance of 16-20
volunteers on three performance tasks before, during, and after MEK exposure. Dick et al.
(1989) evaluated 12 male and 13 female volunteers for neurobehavioral performance changes
and biochemical indicators during and after MEK exposure. In a more recent study by Dick et
al. (1992), exposure of 13 men and 11 women (ages ranged from 18 to 32 years) to 200 ppm
MEK for 4 hours in an environmental chamber found no statistically significant increase in
airway irritation reported by volunteers. Ingested ethanol (95%, 0.84 mL/kg) was used as a
positive control for neurobehavioral effects. The volunteers were evaluated by the same battery
of psychomotor tests noted for the earlier studies, a sensorimotor test, and a test of mood to
measure neurobehavioral effects. Additionally, chemical measurements of MEK concentrations
(venous blood and expired breath) and reports of sensory and irritant effects were recorded.
MEK exposure produced statistically significant performance effects on 2 of 32 measures
(choice reaction time in males only and percent incorrect responses for dual task in females
only). Given the number of comparisons performed, the number of statistically significant
associations was consistent with the number expected by chance alone. The authors concluded
that the observed effects of MEK exposure could not be attributed directly to chemical exposure.
Muttray et al. (2002) exposed 19 healthy male volunteers to 200 ppm (590 mg/m3) MEK
or filtered air for 4 hours in a crossover study design. Mucociliary transport time was measured,
as well as collection of nasal secretions for cytokines (tumor necrosis factor-alpha and
interleukins 6, 8, and 1-beta). The study also assessed acute symptoms via a 17-part
questionnaire that assessed irritation of mucous membranes, difficulties in breathing, and pre-
narcotic symptoms. The volunteers did not report nasal irritation. The only statistically
significant (p = 0.01) change was a 10% increase in mucociliary transport time (median values
were 660 seconds for sham exposure as compared with 600 seconds after exposure to MEK), an
indicator of subclinical rhinitis. The biological significance of this effect is not clear.
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In an earlier study, ten volunteers were exposed to several concentrations of MEK for 3
to 5 minutes to determine a concentration that would be satisfactory for industrial exposure and a
concentration that would be "unpleasant" or objectionable. Volunteers exposed to 100 ppm (295
mg/m3) MEK reported only slight nose and throat irritation, while mild eye irritation was
reported by some subjects at 200 ppm (590 mg/m3), and exposure to 300 ppm (885 mg/m3) was
"conclusively rejected" as an 8-hour exposure (Nelson et al., 1943).
4.1.2.2. Case Studies of Long-term Human Exposure to MEK
Although MEK is a widely used industrial solvent, evidence that MEK may induce
general solvent-like effects such as peripheral or central nerve fiber degeneration in humans is
restricted to a small number of case reports and occupational studies. Three case studies
demonstrated adverse effects following repeated exposure to MEK. First, Seaton et al. (1992)
reported that a maintenance fitter was exposed to MEK for 2-3 hours/day for 12 years.
Exposure was via both dermal and inhalation routes. The worker had developed slurred speech,
cerebral ataxia, and sensory loss in his arms and on the left side of his face. Nuclear magnetic
resonance imaging showed severe cerebellar and brainstem atrophy; however, nerve conduction
studies were normal. A survey of his work area revealed peak MEK concentrations in excess of
1,695 ppm (5,000 mg/m3) during some operations and 10-minute concentrations of
approximately 305 ppm (900 mg/m3).
Callender (1995) reported that a 31-year-old male engineer developed severe chronic
headache, dizziness, loss of balance, memory loss, fatigue, tremors, muscle twitches, visual
disturbances, throat irritation, and tachycardia after working for 7 months in a quality assurance
laboratory where he was exposed daily to MEK and fumes from burning fiberglass material.
Personal protection equipment and formal safety training were not provided. Based on a
physical examination, neuropsychological tests (Poet Test Battery and WHO Neurobehavioral
Core Test Battery), electroencephalographic tests, evoked brain potential tests, nerve conduction
velocity tests, rotational and visual reflex testing, vestibular function testing, and SPECT and
MRI scans of the brain, the patient was diagnosed with chronic toxic encephalopathy, peripheral
neuropathy, vestibular dysfunction, and nasosinusitis. Information concerning the exposure
levels and subsequent possible progression or regression of these conditions was not provided.
In a third case, a 27-year-old man developed multifocal myoclonus, ataxia, and postural
tremor after occupational exposure (through dermal and inhalation pathways) over a 2-year
period to solvents containing 100% MEK (Orti-Pareja et al., 1996). The actual exposure levels
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are unknown. The patient reported symptoms of dizziness, anorexia, and involuntary muscle
movement, beginning about 1 month prior to admission. Neurological examination confirmed
multifocal myoclonus, ataxia, and tremor. Symptoms of solvent toxicity disappeared after 1
month of cessation of exposure and treatment with clonazepam and propranolol. Symptoms did
not reappear after withdrawal of the drugs.
4.1.2.3. Occupational Studies of MEK Exposure
Several occupational studies examined the effects of chronic exposure to MEK. WHO
(1992) reported the results of an occupational study by Freddi et al. (1982) of 51 Italian workers
chronically exposed to MEK. The authors reported that MEK exposure was associated with
slightly, but not statistically significant, reduced nerve conduction velocities (distal axonopathy)
and other symptoms such as: headache, loss of appetite and weight, gastrointestinal upset,
dizziness, dermatitis, and muscular hypotrophy, but no clinically recognizable neuropathy
(Freddi et al., 1982). In addition, a brief report of dermatoses and numbness of fingers and arms
in workers was reported following chronic exposure in a factory producing coated fabric (Smith
and Mayers, 1944 as cited in WHO, 1992). MEK concentration in the factory was estimated to
be 300-600 ppm (885-1,770 mg/m3) in the apparent absence of other solvents (Smith and
Mayers, 1944 as cited in WHO, 1992). In both of these reports, the exposure concentration and
duration are uncertain; thus, they are of limited utility in supporting an association between
MEK exposure and persistent neurological impairment for dose-response assessment.
Oleru and Onyekwere (1992) examined the relative impacts of exposures to MEK,
polyvinyl chloride, leather dust, benzene, and other chemicals for four operations (plastic,
leather, rubber, and tailoring) at a Nigerian shoe factory that had been in existence for 30 years.
MEK exposure occurred only in the leather unit where 43 workers were exposed to leather, dyes,
MEK, and other unspecified solvents that were used to preserve leather. The concentration of
MEK in the shoe factory was not measured. The workers were monitored for pulmonary
function (forced ventilatory capacity and forced expiratory volume). The data were used to
determine obstructive, restrictive, and mixed lung diseases among the study cohort (smoking
status was assessed). The pulmonary function results were compared against prediction
equations for nonindustrially exposed subjects. The subjects were given a questionnaire that
assessed tiredness, headache, sleep disorder, dizziness, and drowsiness. The mean age of the
MEK-exposed cohort was 32.8±4.03 years, and the mean duration of employment was 10.3±4.03
years. Incidences of self-reported symptoms of neurological impairment were elevated among
the leather workers (MEK-exposed subgroup) compared with a referent group of tailors
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(controls). Odds ratio (OR) analysis revealed that the following neurological indices were
statistically significant: headache (27/43, OR = 6.2, p<0.005), sleep disorder (15/43, OR = 4.1,
p<0.01), dizziness (15/43, OR = 16.6, p<0.005), and drowsiness (11/43, OR = 5.2, p<0.05). The
authors did not report 95% confidence intervals for the odds ratios. Although the frequency of
reported chest pain was statistically different from the reference population (p<0.05), the authors
found that pulmonary toxicity (restrictive lung disease as determined by pulmonary function
tests) was not statistically different from controls when age was considered. Association of the
neurological effects reported by Oleru and Onyekwere (1992) with a specific chemical such as
MEK is complicated by concurrent exposure to multiple solvents (including hexacarbon solvents
whose neurotoxicity is reportedly exacerbated by MEK). In addition, the lack of a measured
airborne concentration of MEK limits the utility of the data for use in dose-response assessment.
Mitran et al. (1997) and Mitran (2000) reported the results of a cross-sectional health
study of workers in three Romanian factories exposed to MEK, acetone, or cyclohexane. The
MEK group was composed of 41 exposed and 63 controls from a cable factory where a laquer
containing MEK was applied as a coating. The mean age of the exposed subjects was 36±9.2
years and the mean length of exposure was 14±7.5 years. Workers were exposed to reported
concentrations of 51-116 ppm (149-342 mg/m3) MEK during an 8-hour shift. The control
subjects were similar in age (36±12.3 years) and were reported to be matched for physical effort
required for completion of work tasks, shift characteristics, and socioeconomic factors. Study
participants completed a questionnaire about memory and subjective symptoms of neurological
impairment, responded to questions about alcohol consumption, submitted to a clinical
examination, submitted samples for identification of biological exposure markers, and underwent
motor nerve function tests (conduction velocity, latency, amplitude, and duration of response
following proximal and distal stimulation) and psychological tests. Psychological testing
included tests for reaction times to auditory and visual stimuli, distributive attention, the
Woodworth-Mathews personality questionnaire for psychoneurotic tendencies, and the labyrinth
test to identify quality of attention. Nerve conduction testing was performed on the median and
ulnar nerves of the arm of the dominant hand and the peroneal nerve of the ipsilateral leg.
Several neurotoxic symptoms were reported more frequently by MEK-exposed workers
than control workers (Mitran et al., 1997). Percentages of MEK-exposed and control workers
reporting neurotoxic symptoms were: 17% vs. 4.7% for mood disorders; 28% vs. 17% for
irritability; 31% vs. 9.5% for memory difficulties; 19% vs. 6% for sleep disturbances; 41% vs.
7.8% for headache; and 24% vs. 7.8% for numbness of the hands or feet. Also reported more
frequently by MEK-exposed workers than control workers were symptoms of ocular irritation
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(41% vs. 7% in controls); upper respiratory tract irritation (28% vs. 11%); and various types of
bone, muscle, or joint pain (e.g., 31% vs. 15% for muscular pains). In psychological tests,
MEK-exposed workers were reported to have shown more "behavioral changes, such as
emotional lability, low stress tolerance, and a tendency of hyperreactivity to conflict," but the
data were not sufficiently reported by Mitran et al. (1997) to allow an independent assessment of
the results. The only other information concerning these tests was a statement indicating that
diffuse somatic neurotic changes were the dominant findings in exposed workers. Statistically
significant decreases in mean nerve conduction velocities for the median, ulnar, and peroneal
nerves in the MEK-exposed group were observed when compared with control means by 22, 28,
and 26%, respectively (Mitran et al., 1997). Other statistically significant nerve conduction
variables that were different in the MEK-exposed group included: increased proximal and distal
latencies in the median nerve, increased proximal and distal latencies and decreased proximal
amplitude in the ulnar nerve, and increased proximal latency and decreased distal amplitude in
the peroneal nerve.
The Mitran et al. (1997) report has several weaknesses that limit its ability to support an
association between long-term occupational exposure to MEK at concentrations below 200 ppm
(590 mg/m3) and persistent neurological impairment. The report does not provide information
regarding important methodological details including: (1) criteria for selecting and matching the
exposed and control workers (important confounding variables that can influence nerve
conduction include the type of work [e.g., office vs. physical work], alcohol and tobacco
consumption habits, and height and weight); (2) protocols for assessing exposure levels
experienced by the workers; and (3) protocols used in the nerve conduction tests (e.g., it is not
clear whether the exposed and control subjects were tested at the same location and time and
under the same environmental conditions).
Two reviews (memorandum dated June 27, 2002, from William Boyes and David Herr,
U.S. EPA to Susan Rieth, U.S. EPA; Graham, 2000) of the Mitran et al. (1997) report have noted
that the differences in mean nerve conduction velocities between the two groups could be
explained if the control subjects were tested under higher temperatures. Second, although there
were statistically significant increases in self-reported neurological symptoms in the MEK-
exposed group (e.g., numbness of hands and feet), the reviewers noted that the reliability of self-
reported symptoms is widely recognized as suspect and subject to bias. Confidence in these
findings would be increased if the study had demonstrated a correlation between subjects
reporting symptoms and subjects with poor or subnormal nerve conduction velocity results, but
this type of analysis was not presented. Third, the reviewers observed that the report provides no
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indication of increasing response (either in prevalence of self-reported symptoms or nerve
conduction results) with increasing indices of exposure. Confidence in the symptomatological
and nerve-conduction findings would be increased if such dose-response relationships were
demonstrated. Fourth, the pattern of changes in nerve conduction variables in the MEK-exposed
group was not considered to be consistent with patterns demonstrated for compounds such as
hexane and methyl n-butyl ketone (MnBK), which are well-known to cause peripheral
neuropathy. A U.S. EPA memorandum dated June 27, 2002, from William Boyes and David
Herr to Susan Rieth noted that, for this type of peripheral neuropathy, the distal latency of the
peroneal nerve would be expected to be the most affected; however, the mean distal latency of
the peroneal nerve in the MEK-exposed group was not different from that of the control group.
Finally, the reviewers noted that the Mitran et al. (1997) results are only supported by
inconclusive case reports of neuropathies in a few MEK-exposed individuals and are not
consistent with results from well-conducted studies of animals. For example, a study of rats
exposed to concentrations as high as 5,000 ppm (14,750 mg/m3) MEK, 6 hours/day, 5 days/week
for up to 90 days looked for, but did not find, evidence for nerve fiber degeneration or gross
neurobehavioral changes induced by MEK (Cavender et al., 1983, also reported in Toxigenics,
1981).
In summary, the human case reports and studies by Oleru and Onyekwere (1992) and
Mitran et al. (1997) provide limited and equivocal evidence that repeated exposure to MEK in
the workplace increases the hazard for persistent neurological impairment. The available
occupational studies are limited by inadequate characterization of exposure, multiple solvent
exposure, and study design problems.
Potential for Carcinogenic Effects in Humans
Several epidemiological studies evaluated the potential for carcinogenic effects in
humans associated with MEK exposure. Two retrospective epidemiological mortality studies
conducted by Alderson and Rattan (1980) and Wen et al. (1985) reported that deaths due to
cancer were less than expected in industrial workers chronically exposed to MEK in dewaxing
plants. Spirtas et al. (1991) and Blair et al. (1998) found no clear evidence of increased cancer
risk from occupational exposure to MEK, but some evidence suggests an increased risk between
multiple solvent exposure, which included MEK as a component, and certain cancers among
workers in a degreasing plant. A case-control study of lymphoblastic leukemia in children and
parental exposure to MEK (Lowengart et al., 1987) was considered exploratory and
inconclusive.
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In a historical prospective mortality study of 446 male workers in two MEK dewaxing
plants, the number of observed deaths (46) was below the number expected (55.51), based on
national mortality rates for the U.K. (Alderson and Rattan, 1980). The average follow-up was
13.9 years. Mortality due to cancer was less than expected (13 observed; 14.26 expected),
although there was a significant increase in the number of deaths from tumors of the buccal
cavity and pharynx (2 observed; 0.13 expected). Also, there were significantly fewer deaths
from lung cancer (1 observed; 6.02 expected). Although statistically significant increases in the
incidence of buccal or pharyngeal neoplasms was observed, the findings were regarded by the
authors as due to chance since there were a small number of individuals affected, the researchers
failed to include tobacco use in the study, and the number of separate comparisons between
observed and expected rates. In view of the small number of individuals affected, the authors
concluded that there was no clear evidence of cancer hazard in these workers.
A retrospective cohort study of 1,008 male oil refinery workers occupationally exposed
to MEK in a lubricating-dewaxing solvent mixture (also containing benzene, toluene, hexane,
xylene, and methyl isobutyl ketone) demonstrated a lower overall mortality for all causes,
including cancer, than expected based on mortality data from the U.S. population (Wen et al.,
1985). The increased incidence of buccal and pharyngeal neoplasms reported by Alderson and
Rattan (1980) was not confirmed in this study. Although a statistically significant elevated risk
of mortality from bone cancer was reported (SMR=10.34, 95% CI: 2.1-30.2, 3 observed), the
investigators questioned the validity of this finding because two of the three observed bone
cancers were not primary bone cancers and thus appeared to have been misclassified. The
number of prostate cancer deaths was increased (SMR=1.82, 95% CI: 0.78-3.58, 8 observed, 4.4
expected), but the increase was not statistically significant. The risk of prostate cancer tended to
increase with increasing duration of employment in the lube oil department, but not among
workers in the solvent-dewaxing unit where the exposure to solvents (including MEK) primarily
occurred. Thus, these epidemiological studies (Alderson and Rattan, 1980; Wen et al., 1985)
showed no clear relationship between occupational exposure to MEK and the development of
neoplasms in humans.
A retrospective cohort mortality study was conducted of aircraft maintenance workers
employed for at least one year at Hill Air Force Base, Utah (Spirtas et al., 1991; with 10 years of
follow-up reported by Blair et al., 1998). The MEK-exposed workers were from a total cohort of
14,457 subjects (222,426 person-years for male workers, and 45,359 person-years for female
workers). The numbers of MEK-exposed workers were reported as 32,212 person-years for
male workers and 10,042 person-years for female workers. Associations with cancer mortality
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were also evaluated for 26 other specific chemical categories. Trends in mortality were assessed,
although the data on MEK were limited due to a particular focus on potential carcinogenic risks
posed by trichloroethylene. In general, the risks of mortality due to multiple myeloma, non-
Hodgkin's lymphoma, and breast cancer were elevated for the entire cohort; the authors
examined the relationship between the incidence of these cancers and several solvents (including
MEK).
Spirtas et al. (1991) reported a significantly increased standard mortality ratio (SMR) for
multiple myeloma among women exposed to MEK (SMR = 904, 95% CI: 109-3267, 2
observed), but not among men (SMR = 96, 95% CI: 2-536, 1 observed). The MEK-exposed
subcohort was compared to age- and gender-matched incidences of multiple myeloma among the
population of Utah. The authors applied an alternate analytical method by Thomas-Gait (TG),
which adjusted for age at entry into follow up and competing causes of death to account for the
small number of unexposed subjects in the subcohort. According to the TG analyses, the
association was not statistically significant among women for multiple myeloma and exposures
to MEK (n = 2, chi-square = 1.6, p = 0.204).
In the 10-year follow-up study, Blair et al. (1998) compared the mortality due to multiple
myeloma, non-Hodgkin's lymphoma, and breast cancer among the MEK-exposed subcohort and
internal referents (study subjects without occupational solvent exposure). During the 10-year
follow-up period, one additional death due to multiple myeloma occurred in a female subject.
The risk for multiple myeloma among females was elevated but was not statistically different
from controls (relative risk = 4.6, 95% CI: 0.9-23.2, 3 observed). The finding is consistent with
an earlier report by Spirtas et al. (1991) where the TG analysis was applied. As reported by the
authors of the original (Spirtas et al., 1991) and follow-up (Blair et al., 1998) studies, the small
number of cases and exposures to multiple solvents complicate attempts to relate the mortality
excess for multiple myeloma to specific causes. In addition, given the multiple comparisons
performed, some positive associations would be expected by chance alone. Thus, these studies
(Spirtas et al., 1991; Blair et al., 1998) provide insufficient evidence that MEK is responsible for
elevated risk of cancer.
In an exploratory case-control study, Lowengart et al. (1987) examined the relationship
between acute lymphoblastic leukemia in children and parental exposure to MEK that occurred
one year prior to conception until shortly before the diagnosis of leukemia. The mothers and
fathers of children diagnosed with leukemia and individually matched controls (n = 123 matched
pairs) were interviewed regarding occupational and home exposure to MEK, chlorinated
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solvents, spray paints, dyes, pigments and cutting oils, personal and family medical history, and
lifestyle habits associated with leukemia. The study investigators reported a statistically
significant positive trend for risk of childhood leukemia based on father's frequency of use of all
of the chemicals examined, including MEK. The authors reported an odds ratio for MEK that
appeared elevated, but not statistically so, for the period of paternal exposure after birth of the
child and acute lymphoblastic leukemia (OR = 3.0, 95% CI = 0.75-17.23; 9 exposed cases/3
exposed controls). There was no statistically significant association between a father's exposure
one year before pregnancy or during pregnancy and leukemia in the child. No significant
associations between leukemia and mothers' exposures to specific substances were found,
although few mothers reported occupational exposure to the industrial solvents evaluated in the
study. The investigation is considered an exploratory study, given that exposure levels were
judged according to questionnaires only. Factors that could be confounding covariates such as
other chemical exposures and personal lifestyle were not taken into account in the statistical
analysis. The authors did not provide a biological rationale for why an elevated risk of
childhood leukemia would be associated only with father's exposure after birth, but noted the
possibility that recall bias could have influenced results (i.e., the possibility of better recall of
more recent exposures). Thus, the findings of this study cannot be used to reliably examine the
existence of an association between MEK and cancer.
In summary, the retrospective cohort studies of worker populations exposed to MEK
(four studies of three different worker cohorts) provide no clear evidence of a cancer hazard in
these populations. Because of various study limitations (including sample size, small numbers of
cases, and multiple solvent exposures), these studies are not adequate to support conclusions
about the carcinogenic potential of MEK in humans. A case-control study examining the
association between paternal exposures to several solvents, including MEK, and childhood
leukemia is exploratory in nature and cannot be used to reliably support the existence of any
such association. Overall, the epidemiologic evidence from which to draw conclusions about
carcinogenic risks in the human population is inconclusive. Although there is some suggestion
of increased risk for some cancers (including bone and prostate) and multiple solvent exposure
that includes MEK, there is no clear evidence for a relationship between these cancers and MEK
exposure alone.
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4.2. PRECHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN
ANIMALS-ORAL AND INHALATION
4.2.1. Oral Exposure
Information on the toxicity of MEK in experimental animals following oral exposure is
limited to a few acute studies (see Section 4.4.1.1). No subchronic or chronic toxicity studies of
MEK in experimental animals were located. Since 2-butanol is a metabolic precursor of MEK
(Traiger and Bruckner, 1976), oral toxicity data on 2-butanol were evaluated to determine
whether data gaps in the MEK oral exposure data base could be addressed by oral studies with 2-
butanol. Similarly, the data base for the MEK metabolites, 3-hydroxy-2-butanone and 2,3-
butanediol, were reviewed. No oral repeat-exposure animal studies or human exposure data
were located for 2,3-butanediol. A 2-generation drinking water study of 2-butanol and a 13-
week drinking water study with 3-hydroxy-2-butanone, however, provide information relevant to
an assessment of the potential health effects of repeated exposure to MEK (see Section 4.3 for
the 2-butanol study).
Gaunt et al. (1972) exposed CFE rats (15/sex/group) to 3-hydroxy-2-butanone in drinking
water (0, 750, 3,000, or 12,000 ppm) for 13 weeks. According to the authors, the exposures are
equivalent to mean intakes of 0, 80, 318, or 1,286 mg/kg-day for males and 0, 91, 348, or 1,404
mg/kg-day for females. Additional groups of 5 rats of each sex were exposed to 0, 3,000, or
12,000 ppm 3-hydroxy-2-butanone in their drinking water for 2 or 6 weeks. All rats were
weighed weekly throughout the study and water and food consumption were measured once
weekly over a 24-hour period. Urine was collected during the final week of treatment for
appearance, microscopic constituents, glucose, bile salts, and blood. Also, a urine concentration
test measured the specific gravity and volume of urine produced during a 6-hour period of water
deprivation. At the end of the study, the animals were sacrificed and specimens of all major
organs and tissues were examined histologically. Also, blood cell counts and blood chemistry
were determined at the end of the exposure period. No animals died during the study, and all
appeared normal. The 12,000-ppm rats showed a statistically significant (5-6%) reduction in
body weight gain compared to controls at weeks 8 and 13 (study termination) for both sexes. In
addition, a statistically significant increase in relative liver weight was observed among 12,000
ppm rats of both sexes exposed for 13 weeks (6.5% increase for males and 8.4% for females
when compared to controls). The increased relative liver weight was not accompanied by
changes in liver histology or in the activities of liver enzymes (LDH, SGPT, or SGOT), and was
likely an adaptive response to the hepatic metabolism of 3-hydroxy-2-butanone. Slight, but
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statistically significant, anemia was observed in both sexes of 12,000 ppm rats after 13 weeks of
exposure (in males and females hemoglobin decreased by 4.9 and 4.2% as compared to controls
and red blood cell count decreased by 5.4 and 8.3% with corresponding increases in
reticulocytes, respectively). At study termination, the mean hemoglobin concentrations for all
rats were 14.3, 13.8, 14.4, and 13.65 g/100 mL for 0, 750, 3,000, and 12,000 ppm, respectively.
No other statistically significant effects were noted among rats exposed to 3-hydroxy-2-butanone
compared with the controls. In this study, 3,000 ppm (318 mg/kg-day) was a NOAEL, and
12,000 ppm (1,286 mg/kg-day) was a LOAEL for slight anemia in CFE rats exposed to
3-hydroxy-2-butanone in drinking water for 13 weeks.
4.2.2. Inhalation Exposure
No chronic toxicity studies or cancer bioassays of inhalation exposure to MEK in
experimental animals were located, although a number of less-than-lifetime inhalation toxicity
studies have been reported. Since 2-butanol is a metabolic precursor of MEK (Traiger and
Bruckner, 1976), inhalation toxicity data on 2-butanol were evaluated to determine whether the
data gaps in the MEK inhalation exposure data base could be addressed by toxicity studies with
2-butanol. Similarly, the data bases for MEK metabolites (3-hydroxy-2-butanone and 2,3-
butanediol) were reviewed. No repeat-exposure animal inhalation studies or human exposure
data were located for 3-hydroxy-2-butanone or 2,3-butanediol. No chronic or subchronic
inhalation toxicity studies with 2-butanol were found; however, a developmental inhalation
toxicity study has been conducted (Nelson et al., 1989, 1990) (see Section 4.3.2.2).
Several repeat exposure inhalation studies of MEK in animals (all involving whole body
chamber exposures) have been reported. Many of these studies have focused on the possible
neurotoxicity of MEK, including the development of peripheral and central nerve fiber
degeneration.
Cavender et al. (1983) exposed male and female Fischer 344 rats (15/sex/group) in a
whole body dynamic air flow chamber to MEK 6 hours/day, 5 days/week for 90 days. The
reported time-weighted average exposure concentrations (by gas-liquid chromatography) of
MEK were 0, 1,254, 2,518, or 5,041 ppm (0, 3,700, 7,430, or 14,870 mg/m3). The results of this
study are also reported in a Toxic Substances Control Act (TSCA) Section 4 submission by
Toxigenics (1981). All rats were observed twice daily for clinical signs and mortality. Food
consumption and body weight were determined weekly. At the end of the 90-day exposure
period, the eyes of each animal were examined by ophthalmoscopy, and neurological function
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(posture, gait, tone and symmetry of facial muscles, and pupillary, palpebral, extensor-thrust and
cross-extensor thrust reflexes) was evaluated. Clinical pathology evaluations, including
urinalysis, hematology, and serum chemistry were performed at sacrifice for 10
animals/sex/group. At the study termination, 10 animals/sex/group were subject to routine gross
pathology and histopathology. For routine histopathology, all tissues commonly listed on
standard National Toxicology Program (NTP) protocols were examined microscopically. Organ
weights were obtained for the brain, kidneys, spleen, liver, and testes. Special neuropathological
studies were conducted on the remaining five male and five female rats from each group,
including examination of Epon sections of the medulla and the sciatic nerve for pathologic
changes, and evaluation of teased nerve fiber preparations of the tibial nerve (minimum of 50
individual nerve fibers/animal) by light microscopy for evidence of neuropathy.
Cavender et al. (1983) reported no signs of nasal irritation and no deaths during the 90-
day study. Transient depressions in body weight gain compared to the control were seen in high
dose (5,041 ppm) male and female rats early in the study. While statistically significant, the
reductions did not exceed 8% of the control group weights for males or females. There were no
treatment-related effects on food consumption or in the ophthalmological studies in any MEK-
exposed rats. The evaluation of neurological function (i.e., assessments of posture, gait, facial
muscular tone or symmetry, and four neuromuscular reflexes) revealed no abnormalities
(Toxigenics, 1981). At all exposure concentrations, female rats exhibited statistically significant
(p<0.05) dose-dependent increases in absolute liver weight when compared with controls.
Relative liver weight was statistically increased in the 5,041 ppm females only when compared
on a liver-to-brain weight basis (24% increase compared to controls) or liver-to-body weight
basis (13% increase). In males, absolute and relative liver weights increased by 27% in the
5,041 ppm rats only. Other statistically significant differences in organ weights in 5,041 ppm
female rats included decreased brain weights (absolute-5%, relative-9%), decreased spleen
weights (absolute-5%), and increased kidney weights (relative-11%). Kidney weights were
significantly increased (relative-6%) in 5,041 ppm male rats. Differences in the serum
chemistry values for female rats in the 5,041 ppm exposure group included significant increases
in serum potassium, alkaline phosphatase and glucose, and a significant decrease in SGPT
activity compared to controls. No differences in serum chemistry between MEK-exposed males
and control animals were observed. The only statistically significant difference in hematology
parameters included higher mean corpuscular hemoglobin (average weight of hemoglobin per
erythrocyte) in 5,041 ppm male and female rats, and higher mean corpuscular hemoglobin
concentration (average hemoglobin concentration per erythrocyte) in 5,041 ppm females; the
increase corresponded to a slight but not statistically significant decrease in number of red blood
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cells. Hemoglobin concentrations were similar in the control and exposed groups. With the
exception of larger urine quantity in 5,041 ppm males, no urinalysis parameters were
significantly different in MEK-exposed rats.
Routine gross and histopathological examinations and the special neuropathology studies
revealed no lesions that could be attributed to MEK exposure. Thus, while the increase in
absolute liver weights in 5,041 ppm rats and altered serum enzyme activities in 5,041 ppm
female rats indicated possible liver damage, no histopathological lesions in the liver were
observed. The authors stated that the response may have been the result of a physiological
adaptation mechanism. While decreased brain weights in the 5,041 ppm females suggest
possible effects of MEK exposure on brain tissue, no histopathological lesions of the brain were
observed and neurological function tests revealed no abnormalities.
Minimal to mild lesions of the upper or lower respiratory tract were noted in all control
and MEK-exposed rats. The lesions were coded as chronic respiratory disease and consisted of
"multifocal accumulation of lymphoid cells in the bronchial wall and peribronchial tissues with
occasional polymorphonuclear cells (eosinophils) in the perivascular areas of small veins"
(Toxigenics, 1981). Because the bronchial epithelium remained intact and exudates were not
present in bronchial lumens, the lesions were considered pathologically insignificant. In
addition, the authors reported an increased prevalence of nasal inflammation (including
submucosal lymphocyte infiltration and luminal exudate) across the control and all exposure
groups. There was no difference in the character or severity of lesions among the control and
three treatment groups. The authors suggested that the pulmonary lesions were secondary to
mycoplasma infection; unfortunately, no infectious agent was cultured to verify this etiology.
While there is no indication that respiratory lesions are related to MEK exposure, the possibility
exists that the outcome of the study may have been confounded by exposure to an unidentified
infectious agent. The presence of lesions in the respiratory tract of all animals exposed via
inhalation also prevents obtaining an unconfounded determination of any portal-of-entry effects.
In summary, review of the Cavender et al. (1983) findings reveals effects remote to the
respiratory tract in the 5,041 ppm animals that are of uncertain biological significance, i.e.,
reduced body weight gain, statistically significant increases in relative liver weight (males and
females) and altered serum liver enzymes (females), and decreased brain weight (females). As
noted previously, reported liver effects are more likely indicative of a physiological adaptive
response than toxicity. While the finding of decreased brain weight observed in female rats
raises concerns, it is difficult to interpret. Generally, with a brain weight reduction of 5%, one
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might expect evidence of corresponding pathology; however, no treatment-related brain
pathology was observed in this study. The fact that the reduction in brain weight relative to
controls was observed in only one sex also raises questions about the relevance of the finding.
Thus, while the reduction in brain weight at 5,041 ppm is noteworthy, its biological significance
is uncertain at this time.
LaBelle and Brieger (1955) exposed a group of 25 adult rats (strain and sex not specified)
and 15 guinea pigs (strain and sex not specified) to 235±26 ppm (693±77 mg/m3) MEK 7
hours/day, 5 days/week for 12 weeks. A control group was included, but the number of control
animals was not reported. At the end of the study, 15 rats were examined for histopathology
(organs examined were not specified) and hematology (hemoglobin, erythrocyte, leukocyte,
neutrophil, lymphocyte, and monocyte counts). The remaining 10 rats were reserved for growth
studies. Growth study results demonstrated that 12 weeks of exposure to 235 ppm MEK reduced
body weight gain (mean body weight was 95 g for exposed vs. 135 g for control); however,
neither statistics nor standard deviation on the mean were provided. No adverse effects were
reported for the exposed guinea pigs that could be attributed to MEK exposure. Information on
the guinea pigs is only presented qualitatively in the study. In addition, the authors reported a 4-
hour LC50 of 11,700±2,400 ppm (34,515±7,080 mg/m3) in rats exposed to MEK when narcosis
preceded death. The study is inadequate for use in dose-response assessment since the study is
poorly reported, only one exposure concentration was used in the chronic portion of the study,
and relatively few toxicological parameters were measured.
Saida et al. (1976) found no evidence of peripheral neuropathy (as indicated by paralysis)
following continuous exposure of 12 Sprague-Dawley rats (sex not specified) to 1,125 ppm
(3,318 mg/m3) MEK for 16, 25, 35, or 55 days. Control animals were housed under similar
environmental conditions without MEK exposure. At the end of the exposure period, rats were
sacrificed and the sciatic nerve and foot muscle were excised. Spinal cord and dorsal root
ganglion specimens were taken from the same rats. Additional studies were carried out with up
to 5 months of exposure; no information regarding experimental procedures or endpoints
evaluated was provided. No abnormal clinical findings were observed in animals exposed to
MEK for any of the exposure periods (up to 55 days), although clinical observations were
limited to the nervous system and the clinical data collected was only minimally described.
Quantitative histology (neurofilaments/|J,m2; frequency of inpouching of myelin sheath and
denuded axons/mm2) showed no abnormality in rats exposed for up to 55 days. Although the
authors reported that no abnormalities were observed in rats exposed as long as 5 months, no
further details were provided.
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Male Wistar rats (8 per group) were exposed to 0 or 200 ppm (0 or 590 mg/m3) MEK 12
hours/day for 24 weeks (Takeuchi et al., 1983). Body weight and neurotoxicity endpoints
(motor nerve conduction velocity, distal motor nerve latency, and tail nerve conduction velocity)
were measured prior to exposure and every 4 weeks thereafter. After 24 weeks of exposure, the
tail nerve from 1 rat per group was isolated for histopathology. The authors reported a slight
increase in motor nerve conduction velocity and mixed nerve conduction velocity and a decrease
in distal motor latency at 4 weeks of exposure, although no difference was observed after 8, 12,
16, 20, or 24 weeks. Microscopic examination of the tail nerves revealed no histopathological
lesions after 24 weeks.
Garcia et al. (1978) examined behavioral effects of MEK exposure in rats. An increase in
response rate (lever pressing to obtain a food reward) was reported in a group of six adult
Sprague-Dawley rats (sex unspecified) exposed to MEK at concentrations between 25 and 800
ppm (74 and 2,360 mg/m3) for 2 hours at approximately weekly intervals (the total number of
exposures was not stated). Results at these exposure concentrations were not further reported.
An increase in response rate (lever pressing to obtain a food reward) was also reported in a group
of four rats exposed to 25 ppm for 6 hours compared to preexposure values for the same animals
(Garcia et al., 1978). The effect persisted in some animals for several days. No statistics or
standard deviation in the response rate was reported. The small number of measurements and
variability in postexposure response rates complicate the interpretation of these findings.
Geller et al. (1979) studied behavioral effects in four male baboons (2-years-old) exposed
continuously by inhalation to 100 ppm (295 mg/m3) MEK for 7 days. Operant conditioning
behavior conducted during exposure was compared to preexposure test scores. The operant
behavior selected was a match-to-sample discrimination task. The experiment was designed to
compare the performance of each baboon during exposure to its performance during a clean air
exposure period in the same chamber immediately prior to each MEK exposure period. No
effects on performance of the test in terms of the ability to discriminate visual stimuli were
noted. Although reaction time increased, the extent varied considerably among the four animals.
In two of the four baboons, response times returned to preexposure control values by day 7 of
exposure. The exposure also increased the response time in a delayed "match to sample" task.
This effect, however, was transient and disappeared during the course of repeated exposure. The
authors suggested that this could be an early manifestation of the narcosis observed in rats in the
acute toxicity (LC50) study by LaBelle and Brieger (1955). Thus, Geller et al. (1979) found only
transient neurological effects of MEK in primates at the concentrations studied. It should be
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noted that each baboon was exposed to four different chemicals: acetone, MEK, methyl isobutyl
ketone, and MEK plus methyl isobutyl ketone (in that order).
Couri et al. (1974) exposed 4 cats, 4 rats, 5 mice, and an unknown number of chickens to
1,500 ppm (4,425 mg/m3) MEK 24 hours/day, 7 days/week for 7-9 weeks with no apparent
adverse effects. No paralysis was seen in any of the animals, and MEK did not alter the
histology of the nerves. In a dose range-finding study, an unknown concentration of MEK
reportedly produced a statistically significant elevation in plasma cholinesterase levels in mice,
rats, and chickens. The study was poorly reported and many experimental details required to
evaluate study adequacy were not provided.
In addition to possible neurological effects, portal-of-entry and pulmonary effects of
inhaled MEK have been studied. Five male Wistar rats were exposed to MEK (initially at
10,000 ppm, then reduced to 6,000 ppm) 8 hours/day, 7 days/week for 15 weeks (Altenkirch et
al., 1978). The concentration of MEK was reduced from 10,000 ppm (29,493 mg/m3) to 6,000
ppm (17,696 mg/m3) due to severe irritation of the upper respiratory tract. The authors also
reported that all animals in the MEK-exposed group were somnolent during exposure. The death
of all of the rats at week 7 was attributed to bronchopneumonia rather than MEK exposure. The
authors did not comment on possible connections between bronchopneumonia susceptibility and
MEK exposure.
Toftgard et al. (1981) exposed 4 male Sprague-Dawley rats to 800 ppm (2,360 mg/m3)
MEK for 6 hours/day, 5 days/week for 4 weeks, and examined changes in enzymatic activity in
rat liver. Increased absolute and relative liver weight when compared to controls (p<0.05) and
slight reductions in the in vitro metabolic capacity of liver microsomes were reported in rats
exposed to MEK.
In an earlier experiment intended to assess the effects of MEK on hepatic microsomal
enzyme activity, Couri et al. (1977) continuously exposed an unreported number of young male
Wistar rats to 750 ppm (2,210 mg/m3) MEK for 7 or 28 days. After 7 days of exposure, there
was a significant (p<0.005) reduction in hexobarbital sleep times (16.0±2.4 minutes for exposed
vs. 26.0±2.4 minutes for control). In the group exposed for 28 days, the reduction in sleep times
was less marked (the 28-day results were not reported quantitatively). The results are consistent
with an earlier study (Raunio et al., 1990) reporting that pretreatment with MEK can induce
hepatic detoxification capacity.
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In summary, a number of less than lifetime inhalation studies of MEK have been
conducted. In the 90-day inhalation study by Cavender et al. (1983), the only observed effects
were decreased body weight gain, increased liver weight, altered enzyme levels, and decreased
brain weight at 5,041 ppm (14,870 mg/m3). Other studies of shorter duration have largely
focused on neurological endpoints; many of these studies used either small numbers of animals
or one exposure concentration. Data from these repeat inhalation exposure studies provide no
evidence for MEK-induced nerve degeneration or other persistent neurological effects. Evidence
is available suggesting that MEK can potentiate nerve degeneration produced by certain alkanes
that can be metabolized to gamma-diketones, including n-hexane (Altenkirch et al., 1978) and
MnBK (Saida et al., 1976). The evidence is summarized in Section 4.4.4.
4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES
4.3.1. Studies in Humans
No studies were located that examined the potential for MEK to induce developmental
effects in humans after inhalation or oral exposure. Only one occupational study (Lemasters et
al., 1999) is available that addresses the potential reproductive toxicity of MEK in humans. The
investigators studied the male reproductive effects of solvent and fuel exposure at an aircraft
maintenance facility. The study included 50 males who were exposed to a mixture of solvents
and jet fuel on an Air Force installation and a control group of 8 unexposed men. In this
prospective study, each subject was evaluated before the first exposure and at 15 and 30 weeks
after exposures had begun. Industrial hygiene sampling and expired breath samples were
collected to determine jet fuel exposure as measured by total naphthas, benzene,
1,1,1-trichloroethane, MEK, xylenes, toluene, and methylene chloride. Sperm parameters
(concentration, motility, viability, morphology, morphometrics, and stability of sperm
chromatin) were evaluated. Expired breath sampling demonstrated that exposures were
generally low; all mean measures were below 6 ppm, which is less than 10% of the OSHA
standard for all measured chemicals except benzene. Among the subjects, sheet metal workers
had the highest mean breath levels for total solvents (24 ppb) and fuels (28.3 ppb). Mean values
for most sperm measures remained in the normal range throughout the 30-week exposure period.
When jobs were analyzed by exposure groups, some adverse changes were observed. The paint
shop group, for example, had a significant decline in motility (19.5%) at 30 weeks. The authors
noted a lack of a dose-response association for the observed spermatogenic changes. The study
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is limited, since the exposure concentration and duration are unknown. Also, the results are
confounded by exposure to other solvents and chemicals in the workplace.
4.3.2. Studies in Animals
4.3.2.1. Oral Exposure
No studies concerning reproductive or developmental toxicity of MEK exposure by the
oral route are available. A study is available, however, of the reproductive and developmental
toxicity of 2-butanol, a metabolic precursor of MEK (Cox et al., 1975). As described in Section
3.3, data from rats suggest that almost all (96%) of an administered dose of 2-butanol is
converted to MEK (Traiger and Bruckner, 1976), and that both chemicals are metabolized
through the same intermediates (DiVincenzo et al., 1976; Dietz et al., 1981) as shown in Figure
1. Thus, toxicity data from oral exposure to 2-butanol are considered relevant to MEK.
Cox et al. (1975) conducted a multigeneration reproductive and developmental toxicity
study of 2-butanol, which is quantitatively converted to MEK in the body. The study did not
include statistical analyses of the results, although all collected data were fully reported. The
results are also presented in abstract form by Gallo et al. (1977). Weanling FDRL-Wistar stock
rats (30/sex/group) were given 2-butanol in drinking water at 0, 0.3, 1, or 3% solutions and a
standard laboratory ration ad libitum. Weekly food consumption, fluid intakes, and body
weights were examined to determine the efficiency of food utilization and to calculate the
average daily intake of 2-butanol, which was reported by the authors for the initial 8 weeks of
the study (intake was not reported for subsequent weeks) as: 0, 538, 1,644, and 5,089 mg/kg-day
(males) and 0, 594, 1,771, and 4,571 mg/kg-day (females) for the 0, 0.3, 1, and 3% solutions,
respectively. After 8 weeks of initial exposure, FO males and females from each exposure group
were mated to produce Fl A litters which were delivered naturally and nursed through 21 days of
lactation. F1A litters with more than 8 pups were randomly culled to 8 pups per litter on day 4
after birth. Pup and dam weights were recorded on days 4 and 21 after birth. Various indices of
reproductive performance were recorded (e.g., number of successful pregnancies, litter size,
number of live pups at birth and end of lactation). Because increased mortality and decreased
body weight occurred in the Fl A litters at the 3% dose level (see below), all high-dose parents
and Fl A offspring were given drinking water without 2-butanol between days 10 and 21 of
lactation and then 2% 2-butanol for the remainder of the experimental protocol. The average
daily intake in mg/kg-day at the 2% (initially 3%) exposure level was not reported by the study
investigators; therefore, average daily intakes of 3,384 mg/kg-day in males and 3,122 mg/kg-day
35
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in females were estimated based on a linear regression analysis of the reported average intakes
for males and females at drinking water concentrations of 0, 0.3, 1, and 3%.
After a 2-week post-lactation period, the FO females were remated with males of their
respective exposure groups to produce FIB litters. The FIB pregnancies of 20 pregnant rats per
group were terminated on gestation day 20. Data recorded included numbers of corpora lutea,
implant sites, and resorptions, number of live and dead fetuses, and the sex and weight of live
fetuses. FIB fetuses were also examined for skeletal and visceral malformations and variations.
Selected male and female F1A rats (30 of each sex per exposure group) continued on
their respective treatment protocols (0, 0.3, 1, or 2% 2-butanol) and mated at 12 weeks of age to
produce F2 litters that were delivered and nursed through day 21 of lactation. Indices of second-
generation reproductive performance were assessed, as were F2 pup weights at days 4 and 21.
At day 21 of lactation, FIA adults were sacrificed. Major organs and tissues (35 in all) from 10
male and 10 female FIA rats per exposure group were examined histopathologically, and the
liver and kidneys from all 30 FIA rats per sex/group were examined histopathologically.
At the highest exposure level (3%), net parental (FO) body weight gain was reduced
compared with controls both in males (229 vs. 269 g in controls) and females (130 vs. 154 g in
controls) during the 8 weeks of initial exposure. No differences were found in the efficiency of
food utilization. Following birth of the first litter (FIA) of the parental generation, various
reproduction and lactation responses were measured. The study authors reported no effects on
reproductive parameters. Analysis of FO male rate copulatory success (based on data in
Appendix II) suggests a possible impact of 3% 2-butanol on male reproductive performance.
The incidence of male FO rats that did not successfully copulate with FO females was: 0% (1/30),
0.3% (2/30), 1% (0/30), and 3% (6/30). Data from which to determine copulatory failure were
not provided for other generations. In addition, reduced body weight gain in this high-dose
group could have contributed to copulatory success. For these reasons, the biological
significance of these data for the FO generation males is uncertain.
When compared to the control group, the following effects were noted in the FIA litters
from the high-dose (3%) group: reductions in the mean number of pups/litter born alive (8.46 vs.
10.3), the mean number of pups/litter alive before culling at 4 days (8.12 vs. 10.3), the mean
number of pups/litter alive at 21 days (6.85 vs. 7.68), the mean body weight/pup after culling at
4 days (8.3 g vs. 10.7 g, from Appendix II of Cox et al., 1975), and the mean body weight/pup at
21 days (30 g vs. 49 g, from Appendix II of Cox et al., 1975). The high-dose mean FIA body
36
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weights at 4 and 21 days represent 22 and 39% decreases, respectively, when compared to
control values. The litter mean body weights were decreased relative to control at postnatal days
4 and 21 (5 and 4% for the 0.3% group, and 7 and 10% for the 1% group, respectively). Mean
body weights and associated standard deviations were calculated from the individual litter means
in Appendix II of the Cox et al. (1975) report and are summarized in Table 2.
Table 2. Mean F1A litter body weight on days 4 and 21 in rats exposed to 2-butanol in
drinking water"
Doseb
in mg/kg-
day (%)
0
594 (0.3%)
1,771 (1%)
4,571 (3%)
Number
of litters,
day 4
29
27
30
26
Mean litter
body weight,
day 4 (g)
10.7
10.2
10.0
8.3
Standard
deviation,
day 4
1.1
1.3
1.3
1.8
Number
of litters,
day 21
28
27
30
26
Mean litter
body weight
(g), day 21
49
47
44
30
Standard
deviation,
day 21
3.8
3.9
4.8
11.9
"Mean litter body weights and standard deviations were calculated from individual F1A litter body weight means
in Appendix II of Cox et al. (1975). Body weights were measured to the nearest 0.1 g. From the best available
copy of the report, however, the value to the right of the decimal point could not be read clearly for day 21 body
weight values. Therefore, day 21 body weight data in EPA tabulations may not, in all cases, be correctly reported
to the nearest gram (e.g., 10.0 and 10.9 would be indistinguishable and both treated as 10).
b Doses are average daily intake for female rats for the initial 8 weeks of the study as reported by the authors.
Source: Adapted from Cox et al. (1975).
During the second pregnancy, the high-dose FO dams receiving 2% 2-butanol exhibited
reduced weight gain (94 g) compared to control, 0.3% and 1% dams (gains of 113, 111, and 120
g, respectively). The FIB fetuses of high-exposure dams showed a 10% reduction in average
fetal weight compared with controls (3.74±1.01 g vs. 4.14±1.45 g, respectively). Standard
deviations were calculated from the individual animal data in the appendix of the Cox et al.
(1975) report. No differences in average fetal weight were observed at 0.3% (4.16 g) and 1%
(4.38 g). The difference in the mean fetal weights of the adjusted high-dose (2%) and control
groups was not statistically significant (p>0.05) using a t-test, but when the FIB fetal weight
data were fit by linear dose-response models, log-likelihood ratio tests indicated that mean body
weights significantly decreased with increasing dose levels (see Appendix B, output B-3 for
statistical test results).
The incidence of nidation, early fetal death, and late fetal death did not appear to be
affected in the FIB litters of any exposure group compared with controls (Cox et al., 1975). The
FIB fetuses in the 2% group showed increases in skeletal variations (missing sternebrae, wavy
37
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ribs, and incomplete vertebrae ossification) when compared with the 1% dose group. When
compared with control incidences, however, no differences were apparent (see Table 3). The
investigators provided no explanation for the consistently lower responses observed in the 1%
(mid-dose) group.
Table 3. Incidence of skeletal variations in FIB fetuses
Skeletal variation
Missing sternebrae
Wavy ribs
Incomplete ossification
Incidence (%), fetal basis
[litter basis]
0
51/235(22%)
[10/29 (34%)]
41/235 (17%)
[17/29 (59%)]
56/235 (24%)
[17/29 (59%)]
594 mg/kg-d
(0.3%)
14/211(7%)
[9/27 (33%)]
29/211 (14%)
[14/27 (52%)]
56/211 (27%)
[20/27 (74%)]
1,771 mg/kg-d
(1%)
1 1/254 (4%)
[2/30 (7%)]
20/254 (8%)
[10/30 (33%)]
23/254 (9%)
[10/30 (33%)]
3,122 mg/kg-d
(2%)
46/217(21%)
[13/29 (45%)]
35/217 (16%)
[17/29 (59%)]
69/217 (32%)
[18/29(62%)]
Source: Adapted from Cox et al. (1975).
F2 pups from the high-dose group (2%) showed a reduction in the mean pup body weight
at postnatal day 4 (9.5 g vs. 10.0 g in the control) and in mean pup body weight at day 21 (35 vs.
40 g in the control). Mean body weights of F2 pups in the 0.3 and 1% groups were similar to
controls at day 4 (9.7 and 9.6 g) and day 21 (39 and 39 g). Although body weight reductions in
the high-dose F2 pups were not as great as those observed in the high-dose Fl A pups, a
continued decrease in body weight occurred in the high-dose pups at days 4 and 21 (reductions
of 5% at day 4 and 13% at day 21 when compared with F2 controls).
No exposure-related changes in organ weights or increased incidence of lesions were
found in the adult Fl A rats sacrificed 21 days after the F2 birth, with the exception of specific
histopathologic changes in the kidneys that were most prominent in males (Cox et al., 1975).
Microcysts in the tip of the renal papilla were reported for rats receiving 2% 2-butanol, but not in
control rats; however, the incidence was not reported. Slight to mild hydropelvis was also
observed among control and 2-butanol-exposed rats, although no dose-related effect was
observed. Other changes included tubular cast formation and foci of tubular degeneration and
regeneration. Incidences of male F1A rats with these types of kidney changes were 0/30, 1/30,
1/30, and 8/30 for the control through high-dose groups, respectively. A similar increased
incidence was not observed in females. The findings are consistent with the pattern for early
38
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stages of a2u-globulin-associated rat nephrotoxicity as described by the Risk Assessment Forum
(U.S. EPA, 1991b). Testing was not conducted, however, to demonstrate the presence of the
protein a2u.
In summary, the results of the Cox et al. (1975) study demonstrate that the administration
of 2-butanol in drinking water at concentrations as high as 3% did not affect reproductive
performance in rats (with the possible exception of male rat copulatory success), but produced
maternal toxicity accompanied by developmental effects at the highest exposure level.
Decreased maternal weight gain, decreased Fl A pup survival, and decreased Fl A pup weights at
days 4 and 21 were observed in the groups exposed to 3% 2-butanol in drinking water. At the
next lower dose (1%) in this same generation, only reductions in F1A pup weights (7 to 10% at
days 4 and 21) were observed; however, no similar reductions in body weight were observed in
subsequent generations at the 1% dose level. The following effects were noted at the 2% level (
the adjusted high-dose level administered following F1A postnatal day 21): decreased maternal
body weight gain during the second pregnancy of the FO dams (body weight gain was not
measured during the first, FO, pregnancy nor during the F1A pregnancy), decreased FIB fetal
weights when pregnancy was terminated at gestation day 20, and decreased F2 pup weights at
days 4 and 21. Developmental endpoints were not affected at the 0.3% 2-butanol exposure
levels in any of the generations. 2-Butanol treatment produced an increase in the incidence of
kidney lesions in high-dose (3,384 mg/kg-day) male rats (Fl A generation) that were exposed
from gestation through 12 weeks after birth, mating, and gestation and lactation of the F2
generation; no other treatment-related histopathologic lesions were observed in adult rats. Thus,
Cox et al. (1975) identified a LOAEL of 3,122 mg/kg-day (2% solution) and a NOAEL of 1,771
mg/kg-day (1% solution) based on decreased FIB fetal weights and decreased Fl A and F2 pup
body weights. The maternal LOAEL in this study was 3,122 mg/kg-day (2% solution) based on
decreased weight gain, and the NOAEL was 1,771 mg/kg-day (1% solution).
It should be noted that the Cox et al. (1975) study protocol, although consistent with U.S.
Food and Drug Administration (FDA) guidelines available at the time that the study was
conducted, did not include the evaluation of certain parameters routinely measured in studies of
more current design. Deficiencies included: lack of measurements of estrous cyclicity, sperm
parameters, weights of uterus, epididymides, seminal vesicles, and brain; and less than complete
clinical chemistry/hematology and histopathology. Water consumption was recorded in FO and
Fl A rats prior to mating, but not during gestation and lactation. Consequently, more accurate
measures of offspring exposure could not be developed. Statistical analyses were not performed
by study investigators. In addition, changes in the drinking water concentration of high-dose
39
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animals during the last 2 weeks of FO lactation of Fl A litters from 3% to 0% and then to 2% 2-
butanol introduces some uncertainty in the exposure of high-dose animals.
4.3.2.2. Inhalation Exposure
No studies were located that specifically assessed the reproductive toxicity of inhaled
MEK. Although no tests for reproductive function were performed, histological examination of
the reproductive organs from rats of both sexes and mammary glands of female rats exposed
subchronically to MEK at concentrations as high as 5,000 ppm (14,750 mg/m3) revealed no
exposure-related lesions (Cavender et al., 1983). The data base on developmental toxicity of
MEK by inhalation consists of several well-conducted studies.
Schwetz et al. (1974) exposed groups of 21-23 pregnant Sprague-Dawley rats (in whole
body dynamic exposure chambers) to 1,000 or 3,000 ppm (2,950 or 8,850 mg/m3) MEK vapor,
respectively, for 7 hours/day on gestation days 6-15. Sperm positive vaginal smear was
designated as gestation day 0. Forty-three rats exposed to filtered room air served as controls.
Another control group of 47 pregnant rats was sham exposed. The average measured
concentrations in this study were 1,126 or 2,618 ppm (3,322 or 7,723 mg/m3). The following
endpoints were used to assess exposure-related effects: maternal body weight, food intake, liver
weight, SGPT activity levels, number of implantations, litter size, fetal anomalies, incidence of
resorptions, and fetal body measurements. No evidence of maternal toxicity or change in the
number of resorptions was reported at any concentration. Small, but statistically significant,
decreases in fetal weight and crown-rump length were observed at 1,126 ppm, but not at 2,618
ppm. In the 1,126 ppm exposure group, mean litter weight decreased by 5% and crown rump
length decreased by 3% when compared with air controls. Among 4 litters exposed to 2,618
ppm, 4 fetuses had rare gross malformations, two acaudate fetuses had an imperforate anus, and
2 fetuses had brachygnathia. No gross malformations were found in fetuses from the control or
1,126 ppm exposure groups. A statistically significant increase in the percentage of litters with
fetuses exhibiting gross anomalies at 2,618 ppm was observed when compared with controls
(19% vs. 0%; p<0.05). The malformations had not been observed previously in more than 400
historical control litters of this rat strain. The percentage of litters with specific skeletal
variations (e.g., delayed ossification of skull or sternebrae) were not significantly different from
control percentages in the 1,126 ppm group, but the 2,618 ppm group showed a statistically
significant increase in the percentage of litters with sternebral skeletal variations (43% vs. 11%
in concurrent controls; p<0.05). A statistically significant increase in the percent of litters
exhibiting any skeletal anomaly was observed at 1,126 ppm (95%) but not at 2,618 ppm (81%),
40
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when compared with the control (58%). Percentages of litters with specific soft tissue anomalies
(e.g., subcutaneous edema or dilated ureters) were not significantly elevated in either exposure
group. A statistically significant increase in the percent of litters with any soft tissue anomaly
was observed at 2,618 ppm (76%) but not at 1,126 ppm (70%), when compared with the controls
(51%).
The Schwetz et al. (1974) results indicate that 2,618 ppm was an adverse effect level for
developmental effects in the absence of maternal toxicity, predominately on the strength of the
findings for rarely occurring gross malformations that were not seen in the 1,126-ppm exposure
groups or controls. The biological significance of the developmental findings for the 1,126-ppm
exposure group is not clear. The decreased fetal weight and crown rump length reductions were
very small (3-5% decrease), and statistical significance was not demonstrated for these variables
at the higher exposure level. Likewise, the increased incidence of litters with any skeletal
anomalies at 1,126 ppm (i.e., "total skeletal anomalies") was not statistically demonstrable at
2,618 ppm, and no incidences of specific skeletal anomalies were significantly elevated at 1,126
ppm. Thus, for this study, 1,126 ppm (7 hours/day on gestation days 6-15) is designated as a
NOAEL, and 2,618 ppm is established as a LOAEL for developmental effects. Also, the highest
exposure level, 2,618 ppm, is identified as a NOAEL for maternal toxicity for this study.
Deacon et al. (1981) attempted to repeat and improve upon the Schwetz et al. (1974)
study. Deacon et al. (1981), also reported as Dow Chemical Corporation (1979), included an
additional, lower exposure level (400 ppm). Groups of 26, 19, 19, and 18 Sprague-Dawley dams
were exposed (in whole body dynamic exposure chambers) to nominal MEK concentrations of 0,
400, 1,000, and 3,000 ppm, respectively, for 7 hours/day on gestation days 6-15. The number of
animals in the treatment groups are slightly smaller than the 20 animals/group recommended in
current protocols. Average measured concentrations of MEK during the experiment were 412,
1,002, and 3,005 ppm (1,215, 2,955, and 8,865 mg/m3). Dams exposed to 3,005 ppm MEK
exhibited maternal toxicity consisting of a slight decrease in weight gain (326 g for 3,005 ppm
group vs. 351 g for control; p<0.05 at gestation day 16), and increased water consumption on
days 15-17 (82 mL/day for 3,005 ppm group vs. 69 mL/day for control; p<0.05 at gestation day
16) (Dow Chemical Corporation, 1979). None of the exposure levels produced statistically
significant effects on: the incidences of pregnancy or resorption, the average number of
implantations or live fetuses per dam, fetal weight, or length. No statistically significant
differences in the incidence of external or soft-tissue alterations were observed between the
exposed and control groups. Differences in the incidence of litters with two skeletal variations
occurred in the 3,005 ppm exposure group when compared with the controls. The incidence of
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extra ribs was 2/26 for control litters, compared with 0/19, 0/19, and 6/18 for 412, 1,002, and
3,005 ppm litters, respectively. This finding was statistically significant at the high dose under
Fisher's Exact test. The respective incidences of delayed ossification of the cervical centra were
22/26, 15/19, 16/19, and 18/18, and were not statistically significant by Fisher's Exact test.
Thus, this study found maternal toxicity (decreased weight gain) and fetal toxicity (increased
incidence of extra ribs) at 3,005 ppm MEK 7 hours/day on gestation days 6-15 (LOAEL), but
not at 412 or 1,002 ppm (NOAEL). The study thereby corroborates the developmental effect
levels reported by Schwetz et al. (1974).
A subsequent inhalation developmental toxicity study in CD-I mice (Schwetz et al.,
1991; also reported as Mast et al., 1989 and NTP, 1990) verified the fetal effect levels
established by the two developmental inhalation studies in Sprague-Dawley rats (Schwetz et al.,
1974; Deacon et al., 1981). Groups of 10 virgin Swiss CD-I mice and 33 sperm plug-positive
(gestation day 0) females were exposed to mean concentrations of 0, 398±9, 1,010±28, and
3,020±79 ppm (0, 1,174±27, 2,980±83, and 8,909±233 mg/m3) MEK by inhalation (in whole
body dynamic exposure chambers) for 7 hours/day on gestation days 6-15 and were sacrificed
on day 18 of gestation. At 0, 398, 1,010, or 3,020 ppm MEK, the number of gravid/mated mice
were 26/33, 23/33, 26/33, and 28/33, respectively. In the dams, a slight, concentration-related
increase in liver-to-body-weight ratio was observed. The increase achieved statistical
significance at 3,020 ppm (increase of approximately 7% when compared with the control).
Maternal body weight gain was similar across all groups. Two statistically significant
developmental effects were observed, including: a 5% decrease in mean fetal weight (per litter)
at 3,020 ppm in males when compared with controls and a 4% decrease for all fetuses combined
when compared with controls; and a positive trend for an increased incidence of fetuses with
misaligned sternebrae with increasing exposure level (incidences were 31/310, 27/260, 49/291,
and 58/323 for the control through 3,020 ppm exposure groups, respectively). No statistically
significant trend was found for the increased incidence of litters containing fetuses with
misaligned sternebrae with increasing exposure level. For female fetuses at 3,020 ppm, the
extent of the reduction in litter mean body weight (approximately 4%) was equivalent to the
reduction noted in all fetuses and males, but it did not achieve statistical significance due to the
relatively low fetal weight among female controls. No increase in the incidence of intrauterine
death was observed in any of the exposed groups. No statistically significant increases in the
incidence of malformations occurred, although 4 malformations (cleft palate, fused ribs, missing
vertebrae, and syndactyly) were observed in exposed groups that were not seen in the control
group or in contemporary control data. Based on the absence of both maternal and
developmental toxic effects, a NOAEL of 1,010 ppm was established. Developmental and
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maternal LOAELs were established at 3,020 ppm (7 hours/day on gestation days 6-15) for small,
but statistically significant decreased fetal weight among males, increased incidence of
misaligned sternebrae, and an increased maternal liver-to-body-weight ratio. The results are in
accord with the developmental effect levels established by earlier studies in rats (Schwetz et al.,
1974; Deacon et al., 1981).
While two other studies (Stoltenburg-Didinger et al., 1990; Stoltenburg-Didinger, 1991)
involve inhalation exposure of rats to MEK during gestation (21 days) and lactation (21-30
days), their main focus was to compare the relative susceptibility of adult and juvenile rats to
MEK potentiation of n-hexane peripheral neuropathy. The studies are not useful to assess the
developmental toxicity of MEK alone since the available reports do not clearly describe details
of the experimental conditions or the results from the groups exposed to MEK alone.
Consequently, these particular studies are not further discussed in this document.
Three inhalation developmental studies in rodents demonstrated that exposure (7
hours/day on gestation days 6-15) to approximately 3,000 ppm (8,850 mg/m3) MEK caused
developmental toxicity in the presence of maternal toxicity in rats (Deacon et al., 1981) and mice
(Schwetz et al., 1991), and developmental toxicity in the absence of maternal toxicity in one rat
study (Schwetz et al., 1974).
Additional information relevant to the developmental toxicity of inhaled MEK is
provided by the developmental inhalation toxicity study of Nelson et al. (1989, 1990), wherein
the effects of exposure to industrial alcohols, including butanol isomers, were examined. Nelson
et al. (1989, 1990) exposed gravid Sprague-Dawley rats by inhalation to 2-butanol at 0, 3,500,
5,000, or 7,000 ppm (0, 10,605, 15,150, or 21,210 mg/m3) for 7 hours/day on gestation days
1-19. At these exposure concentrations, the number of gravid/mated rats were 15/16, 16/16,
14/15, and 11/15, respectively. Dams were sacrificed on gestation day 20 (sperm positive
vaginal smear was gestation day zero), and fetuses were serially removed, weighed, sexed, and
examined for external malformations. The frequency of visceral malformations and variations
was determined in one-half of the fetuses, and the frequency of skeletal deviations was
determined in the other half. Maternal toxicity was exhibited in the dams at all three exposure
concentrations as statistically significant reductions in weight gain and food consumption (see
Table 4). The authors reported narcosis (impairment of locomotor activity) at 5,000 ppm and
above.
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Table 4. Maternal and fetal effects in 2-butanol-exposed rats
Endpoint
Exposure Concentration (ppm)
0
3,500
5,000
7,000
Maternal Effects
Day 20 weight gain (g)
(estimated from a graph)
Food consumption (week 3) (g)
105
126±15
80*
113±13
76*
112±17*
22*
99±11*
Fetal Effects
# resorptions/litter
Fetal weight (g)
Male
Female
1.5±1.3
3.3±0.23
3.1±0.22
1.6±1.4
3.1±0.22
2.9±0.20
1.5±0.9
2.7±0.25*
2.6±0.23*
3.8±2.2*
1.5±0.12*
1.4±0.18*
* Significantly different from control (p < 0.05).
Source: Adapted from Nelson et al. (1989).
Inhalation exposure to 2-butanol also produced statistically significant dose-related
effects on certain fetal developmental indices. A statistically significant increase in the number
of resorptions per litter was reported at 7,000 ppm (3.8±2.2) compared with the control
(1.5±1.3). Fetal weights were reduced in all 2-butanol-exposed groups; differences were
statistically significant when compared to the control at 5,000 and 7,000 ppm (see Table 4).
External fetal malformations were not observed. A statistically significant increase in the
incidence of pooled skeletal variations was observed at 7,000 ppm (100%) when compared to
controls (32%). The authors did not report the nature of skeletal variations observed or the
incidence of individual variations. Occasional visceral variations were seen; however, the
authors did not attribute these to 2-butanol treatment. Although marked maternal toxicity was
observed at 7,000 ppm (including weight gain that was less than 25% of the control), the
increase in resorptions and skeletal variations at this concentration cannot necessarily be
attributed to a direct effect of 2-butanol exposure. The types of developmental effects induced
by inhalation exposure to 2-butanol at concentrations below 7,000 ppm during gestation are
generally similar to those identified for inhalation exposure to MEK by Schwetz et al. (1974,
1991) and Deacon et al. (1981), and for oral exposure to 2-butanol by Cox et al. (1975). Body
weight reductions were observed in 2-butanol-exposed rats and in MEK-exposed rats (Schwetz
et al., 1974) and mice (Schwetz et al., 1991). No increase in the incidence of variations,
however, was present in 2-butanol-exposed rats at concentrations that were associated with
various skeletal variations in MEK-exposed rats and mice (Schwetz et al., 1974, 1991; Deacon et
44
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al., 1981). Considering that the ability to detect a change in fetal weight (a continuous variable)
is much greater than for other (dichotomous) fetal endpoints, changes in fetal weight are often
observed at doses below those producing other signs of developmental toxicity (U.S. EPA,
1991a). Because the Nelson et al. (1989) study of 2-butanol included only 15-16 animals per
group compared to the approximately 25 animals per group included in the MEK developmental
toxicity studies (Schwetz et al., 1974, 1991; Deacon et al., 1981), it is possible that the 2-butanol
study did not have sufficient power to detect anomalies.
To assess whether the magnitude of developmental effects associated with the inhalation
of 2-butanol and MEK were similar, fetal weight changes observed in 2-butanol- and MEK-
exposed animals were compared. Figure 2 shows the relationship between fetal weight
(expressed as percent change from control) and exposure concentration for 2-butanol [based on
data for male rat fetuses from Nelson et al. (1989)] and for MEK [based on data for rat fetuses
from Schwetz et al. (1974) and mouse fetuses from Mast et al. (1989)/Schwetz et al. (1991)].
Although the range of exposure concentrations used in the 2-butanol study exceeded the range of
exposure concentrations used in the MEK studies, visual inspection of the graph demonstrates
that the dose-response curves for 2-butanol and MEK are consistent.
A summary of key repeat exposure reproductive and developmental toxicity studies in
animals exposed to MEK and 2-butanol is available in Table 5.
4.4. OTHER STUDIES
4.4.1. Acute Toxicity Data
4.4.1.1. Oral Exposure
Oral LD50 values for MEK include 5,522 and 2,737 mg/kg in rats (Smyth et al., 1962 and
Kimura et al., 1971, respectively) and 4,044 mg/kg in mice (Tanii et al., 1986). A single gavage
dose of 15 mmol/kg MEK (1,082 mg/kg) in corn oil produced no deaths or histological
alterations in the livers of male Fischer 344 rats, but produced tubular necrosis in the kidneys
(Brown and Hewitt, 1984).
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Figure 2. Comparison of Fetal Body Weight Changes in Animals Exposed
to MEK or2-Butanol during Gestation
60
-| 50
o
O
E
2 40
D)
30
T3
O
00
s20
cc
CD
b
CD
Q
10
Schwetz etal., 1974
(MEK-exposed rats)
-Mast etal., 1989
(MEK-exposed mice)
-Nelson etal., 1989
(2-butanol-exposed
male rats)
1000 2000 3000 4000 5000 6000
Exposure Concentration (ppm)
7000
8000
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4.4.1.2. Inhalation Exposure
LaBelle and Brieger (1955) reported a 4-hour LC50 for MEK of 11,700±2,400 ppm
(34,515±7,080 mg/m3) in rats. Several studies describe the behavioral effects of acute inhalation
exposure of mice to MEK (Section 4.2.2.). Glowa and Dews (1987) exposed a group of 12
adult, male CD-I mice to air concentrations of MEK that were increased at 30-minute intervals
until the mice failed to respond to a visual stimulus (response to a visual stimulus and the
response rate were used as indicators). The concentrations for each 30-minute period were 300,
1,000, 3,000, 5,600, and 10,000 ppm (885, 2,950, 8,850, 16,520, and 29,500 mg/m3) MEK with a
total exposure time of 2 hours. No effects were observed at 300 ppm, while a slight decrease in
response rate was observed at 1,000 ppm and a 75% decrease in response rate was observed at
3,000 ppm. Most mice (incidence not reported) ceased to respond at 5,600 ppm, and all failed to
respond at 10,000 ppm. The response rate returned to the control value 30 minutes after
exposure ended. The EC50 (concentration expected to elicit a 50% decrease in response rate) was
calculated to be 2,891 ppm (SD = 689 ppm). From these results, an EC10 (concentration
estimated to elicit a 10% decrease in response rate) was calculated and dose-response estimates
were derived. The concentrations of MEK producing a 10% decrease in response rate in 0.1, 1,
and 10% of a population were calculated to be 17, 66, and 300 ppm, respectively (Glowa and
Dews, 1987).
The EC50 established by Glowa and Dews (1987) for response to a visual stimulus in CD-
1 mice (2,891 ppm) is similar to an EC50 for behavioral effects induced by MEK in Swiss mice.
Groups of 10 adult male Swiss mice were exposed via whole-body inhalation chamber to MEK
at 0, 1,602, 1,848, 2,050, or 2,438 ppm (0, 4,726, 5,452, 6,048, or 7,192 mg/m3) for 4 hours
(DeCeaurriz et al., 1983). Immediately after exposure, mice were subjected to the behavioral
despair swimming test, where the decrease in total time of immobility during the first 3 minutes
in a water bath was used as an indication of behavioral toxicity. MEK exposure produced a
statistically significant (p<0.05) decrease in immobility in the behavioral despair swimming test
at all exposure concentrations tested. Based on this data, the authors calculated a 50% decrease
in immobility (ID50) for MEK of 2,065 ppm. No other observations of the effects of inhalation
exposure of mice to MEK were reported in this study.
4.4.2. Genotoxicity
MEK is not mutagenic as indicated by a number of conventional short-term assays for
genotoxic potential. A battery of in vitro tests showed that MEK was not genotoxic in the
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following assays: Salmonella (Ames) assay with or without metabolic activation, the
L5178/TK+/" mouse lymphoma assay, and the BALB/3T3 cell transformation assay. MEK did
not induce unscheduled DNA synthesis in rat primary hepatocytes (O'Donoghue et al., 1988).
MEK also tested negative in a battery of in vitro tests (Salmonella, chromosome aberration, and
sister chromatic exchange) conducted by the National Toxicology Program (NTP, undated).
MEK was not mutagenic in Salmonella typhimurium strains TA98, TA100, TA1535, or TA1537
in the presence or absence of rat hepatic homogenates (Florin et al., 1980; Douglas et al., 1980;
Zeiger et al., 1992). No induction of micronuclei was found in the erythrocytes of mice
(O'Donoghue et al., 1988) or hamsters (WHO, 1992) after intraperitoneal injection with MEK.
The only evidence of mutagenicity was mitotic chromosome loss at a high concentration in a
study on aneuploidy in the diploid D61, M strain of the yeast Saccharomyces cerevisiae
(Zimmermann et al., 1985); the relevance of this positive result to humans is unknown. Low
levels of MEK combined with low levels of nocodazole (another inducer of aneuploidy) have
also produced significantly elevated levels of aneuploidy in the S. cerevisiae test system (Mayer
and Goin, 1987).
4.4.3. Carcinogenicity
As discussed in Section 4.2, no cancer bioassay is available from which to assess the
carcinogenic potential of MEK in experimental animals by the oral or inhalation routes. In a
skin carcinogenesis study designed to investigate the contribution of sulfur compounds to tumor
induction by unrefined mineral oils, groups of 10 to 15 male C3H/He mice received dermal
applications of various solvent mixtures, some containing MEK (Horton et al., 1965). Mice
received 50 mg of a solution containing: (1) 25% MEK, 70% dodecylbenzene, and 5% benzyl
disulfide; (2) 29% MEK, 70% dodecylbenzene, and 0.8% 2-phenylbenzothiophene; or (3) 17%
MEK, 50% dodecylbenzene, and 33% decalin twice a week for 1 year. No skin tumors
developed in the groups of mice treated with solvents containing 25% MEK with 5% benzyl
disulfide (a weak accelerant for skin tumors in C3H mice). After 27 weeks, a single skin tumor
developed in 1 of 10 mice treated with the solution containing 29% MEK, and after 51 weeks, a
skin tumor developed in 1/15 mice treated with the solution containing 17% MEK and other
solvents. This study is an inadequate test of MEK carcinogenicity due to concomitant exposure
to sulfur-containing chemicals and dodecylbenzene (which are expected to accelerate the rate of
skin tumor formation).
Using mechanism-based structure-activity relationship (SAR) analysis, it was determined
that MEK is unlikely to be carcinogenic based on the lack of any structural features/alerts
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indicative of carcinogenic potential (Woo et al., 2002).
4.4.4. MEK Potentiation of Peripheral Neuropathy from Chemicals Metabolized to
Gamma-Diketones
A number of studies in experimental animals demonstrate that MEK potentiates the
effects of known neurotoxicants (e.g., n-hexane, MnBK, and 2,5-hexanedione) (Saida et al.,
1976; Altenkirch et al., 1978; Takeuchi et al., 1983). Saida et al. (1976) found peripheral
neuropathy in rats after 25 days of continuous exposure to MEK and MnBK at concentrations of
1,125 ppm (3,319 mg/m3) MEK and 225 ppm MnBK. In contrast, rats exposed to 225 ppm
MnBK alone developed peripheral neuropathy after 66 days. In a study with n-hexane and
MEK, Altenkirch et al. (1978) reported that the onset of clinical and morphological effects was
shortened and that the extent and severity of lesions in the peripheral and central nervous
systems increased at an exposure of 9,000 ppm n-hexane and 1,000 ppm (2,950 mg/m3) MEK as
compared to 10,000 ppm n-hexane alone. Altenkirch et al. (1982) also examined nervous system
response to n-hexane and a mixture of n-hexane and MEK. Animals exposed continuously to
500 ppm n-hexane alone displayed hind limb paralysis after 9 weeks, as well as axonal lesions in
peripheral nerves. In rats treated with a mixture of n-hexane (300 ppm) and MEK (200 ppm)
similar clinical and pathological signs of neuropathy occurred one week earlier. Takeuchi et al.
(1983) reported that distal motor nerve latency was significantly reduced at 4 weeks of exposure
to 100 ppm n-hexane plus 200 ppm (590 mg/m3) MEK. While this effect did not persist, it was
not seen with exposure to 100 ppm n-hexane alone or 200 ppm MEK alone. In addition, tail
nerve conduction velocity in rats exposed to a mixture of 100 ppm n-hexane and 200 ppm MEK
was statistically reduced as compared to control at 20 and 24 weeks of exposure, an effect that
was not seen with exposure to n-hexane alone at 100 ppm. Microscopic examination of the tail
nerves revealed no histopathological lesions after 24 weeks.
Evidence in humans that MEK has the capacity to interact with other solvents is less
clear. In a series of studies in human volunteers by Dick et al. (1984, 1988, 1989, 1992), MEK-
exposed groups (at 100 ppm) that were coexposed to relatively low levels (also around 100 ppm)
of several other solvents including acetone, methyl isobutyl ketone and toluene, for 4 hours
exhibited no evidence of neurotoxic interactions. Altenkirch et al. (1977) reported the
occurrence of polyneuropathies in juveniles who sniffed glue thinner following the change in
composition of the thinner from one containing n-hexane and other solvents to one that included
MEK in the composition. A recent review (Noraberg and Arlien-Soborg, 2000) reported
possible interactions following occupational exposure to mixtures of organic solvents containing
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MEK, although because of the nature of the exposures, the studies cannot be used to establish a
causal relationship between cases of neuropathy and specific chemical exposures. For example,
Dyro et al. (1978) reported three cases of polyneuropathy in shoe factory workers exposed to
MEK, acetone and toluene; the potential for dermal contact was noted but not further
characterized. Allen et al. (1974) found evidence of neuropathies in 79 of 1,161 employees in a
fabric plant where workers were regulatory exposed to methyl butyl ketone and MEK. Air
concentrations of MEK reached levels as high as 5,000 mg/m3 and employees washed their
hands with these solvents. Upon removal of methyl butyl ketone from the plant and efforts to
reduce solvent exposure, no new cases of neuropathy developed. Whether there were any
interactive effects between MEK and methyl butyl ketone cannot be ascertained. Fagius and
Gronquist (1978) performed a study of polyneuropathy in 42 steel plant workers exposed to 18
solvents, including MEK. Three possible cases of polyneuropathy were found (and none in a
referent population). Measurement of 11 neurological tests revealed only weak and inconclusive
evidence of decrements in peripheral nerve function in the solvent exposed population. Chia et
al. (1993) investigated neurobehavioral effects in workers exposed to MEK, cyclohexanone,
tetrahydrofuran and toluene in a video tape manufacturing facility in Singapore. Three of 7
neurobehavioral tests (indicative of visual motor control and recent memory impairment)
revealed statistically significant differences between the exposed group and matched controls,
although no dose-effect relation was observed. The possibility of extensive skin contact with the
solvents was noted by the authors.
None of the available occupational studies involving multiple chemical exposures as
discussed above provide information adequate to establish whether MEK interacts with other
neurotoxic solvents in humans. Further, the studies do not provide information to establish the
lower limit of MEK exposure that may result in potentiation of effects by known neurotoxicants.
From the review by Noraberg and Arlien-Soborg (2000), however, it appears that neurotoxicity
was observed only in worker populations exposed to solvent mixtures where reported MEK air
concentrations reached levels at or above the TLV (200 ppm; 590 mg/m3).
The mechanism by which MEK potentiates the neurotoxicity of hexacarbon solvents is
not entirely clear, although it appears to involve the biotransformation of these solvents to their
toxic metabolites (such as 2,5-hexadione (2,5-HD), which is the putative moiety responsible for
inducing neural damage associated with n-hexane exposure) (DiVincenzo et al., 1976; van
Engelen et al., 1997; Ichihara et al., 1998). In the case of 2,5-HD, the potentiation effect appears
to be due to the increased persistence of 2,5-HD in blood, probably due to inhibition by MEK of
2,5-HD phase II biotransformation that alters the metabolism and elimination of 2,5-HD (van
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Engelen et al., 1997; Zhao et al., 1998; Yu et al., 2002). Consistent with this proposed
mechanism of potentiation are the findings of Abdel-Rahman et al. (1976), wherein blood
samples from rats exposed continuously to 400 ppm MnBK for 6 days revealed no detectable
levels of MnBK, whereas MnBK was present at detectable levels (reported as 9.5 mg%) in blood
samples from rats exposed continuously to a mixture of MnBK (225 ppm) and MEK (750 ppm).
In contrast to n-hexane and MnBK, MEK is not metabolized to a gamma-diketone (a diketone in
which the two carbonyl groups are separated by two carbons) (DiVincenzo et al., 1976). The
difference is significant, since the gamma-diketones (in contrast to MEK's metabolites) have
been associated with distal neurofilamentous axonopathy (Graham, 2000).
In general, potentiation of the neurotoxicity of other solvents by MEK has been
demonstrated in experimental animals only at relatively high concentrations (>1,000 ppm or
2,950 mg/m3) where induction of hepatic enzymes (liver enzymes that are responsible for
toxifying the gamma-diketones) is postulated as the mode of action. Studies of MEK
potentiation of neurotoxicants at lower exposure concentrations have generally not been
performed. One exception is Takeuchi et al. (1983) in which reversible potentiation of n-hexane
neurotoxicity was observed at 200 ppm (590 mg/m3) MEK. The work of van Engelen et al.
(1997) provides some insight into the lower limits of interactive effects of MEK and n-hexane in
humans. Volunteers were exposed to n-hexane (approximately 60 minutes) with or without
coexposure to MEK (200 or 300 ppm) for 15.5 minutes, while the concentration-time course of
n-hexane (in exhaled alveolar air) and its metabolite 2,5-HD (in serum) were measured.
Coexposure to 200 ppm MEK did not affect the concentration-time course of exhaled n-hexane
or the rate of formation of serum 2,5-HD; however, MEK significantly decreased the rate of 2,5-
HD formation (approximately 3-fold) at 300 ppm. Coexposure to 300 ppm MEK also
significantly increased the time to reach peak concentration of 2,5-HD (Tmax). At 200 ppm MEK,
there was a trend to higher values of Tmax, but the effect was not statistically significant. The
investigators cautioned that their findings could not be used to predict interactive effects
resulting from chronic exposure. The findings do suggest that, at least following short-term
exposure, significant interactive effects may occur at levels somewhat above the TLV (200
ppm).
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4.5. SYNTHESIS AND EVALUATION OF MAJOR NONCANCER EFFECTS
4.5.1. Oral Exposure
Data on the toxic effects associated with oral exposure of humans to MEK are limited to
a single nonoccupational report of acute toxicity following accidental ingestion of MEK
(Kopelman and Kalfayan, 1983). The report did not indicate any persistent adverse health
effects. In laboratory animals, the data base on toxicity of MEK following oral exposure is
limited to a small number of acute studies. LD50 values for adult mice and rats are 2-6 g/kg
body weight, with death occurring within 1-14 days following a single oral dose (Tanii et al.,
1986; Kimura et al., 1971; Smyth et al., 1962). The lowest, non-lethal acute oral dose producing
an adverse effect is a report of renal tubule necrosis in F344 rats following a single oral dose of
1,082 mg/kg of MEK in corn oil (Brown and Hewitt, 1984).
Subchronic and chronic toxicity studies of oral MEK exposure are not available. Repeat-
dose toxicity data are available, however, for 2-butanol (a metabolic precursor) and 3-hydroxy-2-
butanone (a metabolite). In rats, the majority of an oral dose of 2-butanol is rapidly converted to
MEK (Traiger and Bruckner, 1976; Dietz et al., 1981); both MEK and 2-butanol are transformed
to common metabolites (3-hydroxy-2-butanone and 2,3-butanediol) in the rat (Dietz et al., 1981).
In rats administered similar oral doses of MEK or 2-butanol, the elimination kinetics for the
common metabolites are similar (Dietz et al., 1981).
The oral toxicity data base for 2-butanol consists of a two-generation reproductive and
developmental toxicity study in rats (Cox et al., 1975). The administration of 2-butanol in
drinking water before and during gestation and lactation at concentrations as high as 3% did not
affect reproductive performance (with the possible exception of increased male copulatory
failure), but did result in decreased pup survival and pup body weight gain in Wistar rats. A
concentration of 2% in drinking water caused a reduction in fetal weights when pregnancies
were terminated on gestation day 20 and decreased pup body weights when dams were allowed
to deliver. Cox et al. (1975) reported a LOAEL of 3,122 mg/kg-day (2% solution) and a
NOAEL of 1,771 mg/kg-day (1% solution) for decreased fetal weight and decreased pup body
weight gain. The findings of developmental toxicity in rats exposed orally to 2-butanol is
consistent with similar findings in inhalation developmental toxicity studies of MEK discussed
in Section 4.3.2.2 (Schwetz et al., 1974, 1991; Deacon et al., 1981) and 2-butanol (Nelson et al.,
1989, 1990). Given these observations, it is plausible that the developmental effects produced
by 2-butanol and MEK are caused by MEK or a subsequent metabolite common to both.
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In adult rats, exposure to 3% 2-butanol in drinking water for 8 weeks caused reduced
weight gain in FO males and females (Cox et al., 1975). Fl animals exposed to 2-butanol in
drinking water at concentrations up to 2% for 12 weeks after birth and through mating, gestation,
and lactation of F2 litters were subject to gross and histopathological examination. No exposure-
related changes in organ weights or incidence of histopathologic lesions were observed with the
exception of specific histopathologic changes of the kidney in male rats exposed to 2% 2-
butanol. Changes were consistent with the pattern of early stages of a2u-globulin-associated rat
nephrotoxicity; however, testing needed to demonstrate the presence of a2u was not conducted.
Therefore, the relevance of this finding to humans is uncertain.
The oral toxicity data base for 3-hydroxy-2-butanone consists of a 13-week drinking
water study in rats (Gaunt et al., 1972). Thirteen weeks of drinking water exposure to
3-hydroxy-2-butanone in CFE rats (15/sex/dose) did not produce a toxic effect aside from slight
anemia (decreased hemoglobin concentration and red blood cell count) at the high dose (1,286
mg/kg-day) (Gaunt et al., 1972), an effect that has not been reported following exposure to 2-
butanol (orally; Cox et al., 1975) or MEK (by inhalation; Cavender et al., 1983). In the
Cavender et al. (1983) study, hemoglobin concentrations were unaffected by inhalation exposure
to MEK; at 15,000 mg/m3, there was a statistically significant increase in mean corpuscular
hemoglobin that corresponded to a slight but insignificant decrease in red blood cells. Further,
Gaunt et al. (1972) provides no information concerning the potential for developmental effects
from exposure to 3-hydroxy-2-butanone. This observation further supports the use of 2-butanol,
rather than a metabolite, as a surrogate for MEK.
In summary, information on the effects of MEK following repeat-dose oral exposure is
limited to data for 2-butanol (a metabolic precursor) and 3-hydroxy-2-butanone (a metabolite).
Because of the similarity in the effects of exposure to MEK and 2-butanol, as well as the finding
that 2-butanol is rapidly converted to MEK in rats, 2-butanol is considered to be an appropriate
surrogate for assessing MEK-associated toxicity. A multigeneration reproductive and
developmental toxicity study of 2-butanol by Cox et al. (1975) identified developmental effects
(reduced fetal and pup weight) as the most sensitive lexicologically relevant endpoint.
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4.5.2. Inhalation Exposure
Evidence for neurotoxic effects following inhalation exposure to MEK is limited to a
small number of case reports of neurological impairment in occupationally-exposed humans
(Welch et al., 1991; Seaton et al., 1992; Callender, 1995; Orti-Pareja et al., 1996) and in one
study of problematic design reporting increased incidence of subjectively reported neurological
symptoms in MEK-exposed workers (Mitran et al., 1997; Graham, 2000). A few animal studies
involving a single or limited number of inhalation exposures reported behavioral effects and
narcosis (Nelson et al., 1989, 1990; Glowa and Dews, 1987). Several well-conducted studies in
experimental animals, however, provide no convincing evidence that repeated exposure to MEK,
by itself, is capable of producing persistent neurological effects. No persistent, treatment-related
central or peripheral neural histopathology was observed in rats exposed for 90 days (6
hours/day, 5 days/week) to MEK at concentrations up to 5,041 ppm (14,870 mg/m3) (Cavender
et al., 1983). Repeated exposure of rats and mice to MEK at approximately 3,000 ppm (8,850
mg/m3) (7 hours/day during days 6-15 of gestation) produced no overt neurological effects in the
dams (Schwetz et al., 1974, 1991; Deacon et al., 1981).
Developmental effects following exposure to MEK have been described in experimental
animals, but not humans. Three inhalation developmental studies in rodents demonstrated that
MEK caused developmental toxicity in the presence of maternal toxicity in rats (Deacon et al.,
1981) and mice (Schwetz et al., 1991), and in one rat study (Schwetz et al., 1974) in the absence
of maternal toxicity. These inhalation studies provide evidence for developmental effects
(decreased fetal weight and increased incidence of certain skeletal variants) in rats and mice
exposed to 3,000 ppm MEK, 7 hours/day during gestation, but not at 1,000 ppm and lower. The
observation of developmental delays following inhalation exposure to MEK is supported by the
findings from studies of rats exposed orally (Cox et al., 1975) and by inhalation (Nelson et al.,
1989, 1990) to 2-butanol, a metabolic precursor of MEK.
Available data provide no clear evidence for other systemic effects resulting from
inhalation exposure to MEK. A subchronic inhalation study of MEK found no persistent body
weight changes, gross behavioral changes, or histological changes in major tissues and organs in
rats exposed 6 hours/day, 5 days/week for 90 days to concentrations as high as 5,000 ppm
(14,750 mg/m3) (Cavender et al., 1983). Some changes in organ weight (including increased
liver weight and decreased brain weight) and clinical pathology parameters were observed;
however, these were not supported by histological changes.
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The available data provide no evidence for portal-of-entry effects following inhalation
exposure to MEK. In a series of studies involving numerous volunteers, Dick et al. (1984, 1989,
1992) did not find any reported net effects related to irritation from MEK at exposures up to 200
ppm (590 mg/m3) for up to 4 hours. In an earlier study involving few subjects and unclear
exposure conditions, exposure to 300 ppm (885 mg/m3) MEK was reported as intolerable
(Nelson et al., 1943). Nasal irritation was noted in rats exposed to 6,000 ppm MEK for 15 weeks
(Altenkirch et al., 1978), but not in other studies involving somewhat lower exposure
concentrations. In the only available subchronic animal inhalation study of MEK (Cavender et
al., 1983), no exposure-related upper respiratory irritation could be evaluated in rats exposed up
to 5,000 ppm (14,750 mg/m3) MEK for 90 days (confounding respiratory tract lesions were
likely due to an infectious agent that occurred in all groups in this study including controls). In
addition, respiratory irritation was not reported in dams exposed to 3,000 ppm (8,850 mg/m3)
MEK, 7 hours/day for days 6-15 of gestation (Schwetz et al., 1974, 1991; Deacon et al., 1981).
Results from studies of pregnant rodents exposed by inhalation to MEK indicate that
developmental effects are the most sensitive, lexicologically relevant endpoint for inhalation
exposure to MEK.
4.5.3. Mode of Action Information
The mode of action by which MEK induces toxicity has not been characterized.
4.6. WEIGHT-OF-EVIDENCE EVALUATION AND CANCER CHARACTERIZATION
Under EPA's draft revised cancer guidelines (U.S. EPA, 1999), "data are inadequate for
an assessment of human carcinogenic potential for MEK, because studies of humans
chronically-exposed to MEK are inconclusive, and MEK has not been tested for carcinogenicity
in animals by the oral or inhalation routes. The majority of short-term genotoxicity testing of
MEK has demonstrated no activity, and SAR analysis suggests that MEK is unlikely to be
carcinogenic.
The few available epidemiological studies of MEK-exposed workers provide no clear
evidence of a cancer hazard, but the studies are generally inadequate to discern an association
between MEK exposure and an increased incidence of cancer (Alderson and Rattan, 1980; Wen
et al., 1985; Spirtas et al., 1991; Blair et al., 1998). In these studies, the epidemiological
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evidence is based on a small number of site-specific deaths, and studies that are confounded by
exposure to multiple chemicals. A case-control study examining the association between
paternal exposures to several solvents, including MEK, and childhood leukemia (Lowengart et
al., 1987) is exploratory in scope and cannot be used to reliably support the existence of any such
association. Overall, the epidemiologic evidence from which to draw conclusions about
carcinogenic risks in the human population is inconclusive. Although there is some suggestion
of increased risk for certain cancers (including bone and prostate) involving multiple solvent
exposures that include MEK, there is no clear evidence for a relationship between these cancers
and MEK exposure alone.
No cancer bioassay is available from which to assess the carcinogenic potential of MEK
in experimental animals by the oral or inhalation routes. A skin carcinogenesis study by Horton
et al. (1965) is an inadequate test of MEK's potential carcinogenicity due to concomitant
exposure to chemicals that are expected to accelerate the rate of skin tumor formation.
MEK has not exhibited mutagenic activity in a number of conventional short-term test
systems. In vitro tests showed that MEK was not genotoxic in the Salmonella (Ames) assay with
or without metabolic activation, the L5178/TK+/" mouse lymphoma assay, and the BALB/3T3
cell transformation assay. The tests did not induce unscheduled DNA synthesis in rat primary
hepatocytes, chromosome aberrations, or sister chromatic exchange (Florin et al., 1980; Douglas
et al., 1980; O'Donoghue et al., 1988; NTP, undated; Zeiger et al., 1992). No induction of
micronuclei was found in the erythrocytes of mice (O'Donoghue et al., 1988) or hamsters
(WHO, 1992) after intraperitoneal injection with MEK. The only evidence of mutagenicity was
mitotic chromosome loss at a high concentration in a study on aneuploidy in the diploid D61, M
strain of the yeast Saccharomyces cerevisiae (Zimmerman et al., 1985); the relevance of this
positive result to humans is unknown. In general, studies of MEK yielded little or no evidence
of mutagenicity. SAR analysis suggests that MEK is unlikely to be carcinogenic based on the
absence of any structural alerts indicative of carcinogenic potential (Woo et al., 2002).
4.7. SUSCEPTIBLE POPULATIONS AND LIFESTAGES
4.7.1. Possible Childhood Susceptibility
No specific data are available that assess the potential differences in susceptibility to
adverse effects from MEK exposure between children and adults. At certain stages in their
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development, children have differences in levels of cytochrome P450 enzymes and several phase
II detoxification enzymes (e.g., N-acetyl transferases, UDP-glucuronyl transferases, and
sulfotransferases) relative to adults (Leeder and Kearns, 1997; Vieira et al., 1996). Quantitative
data on the possible contributions of these differences to potential age-related toxicity from
MEK are lacking. Available results from animal inhalation developmental toxicity studies
(Schwetz et al., 1974, 1991; Deacon et al., 1981) suggest that MEK or its metabolites may cross
the placenta and may produce developmental effects. An exploratory case-control study showed
an increased risk (not statistically significant) of childhood leukemia associated with paternal
(but not maternal) exposure to MEK after birth of the child (Lowengart et al., 1987). Given the
nature of the exposure information (self-reported via questionnaire), evidence for possible
childhood susceptibility as provided by this study is considered very limited.
4.7.2. Possible Gender Differences
Available studies in humans and animals provide little evidence of any biologically
relevant gender-related differences in the toxicity of MEK. Human occupational studies have
failed to report sex-related differences in MEK toxicity by any route. The 90-day subchronic
inhalation study by Cavender et al. (1983) suggests that female rats may be slightly more
susceptible to the toxic effects of MEK (decreased absolute and relative brain and liver weight,
as well as altered blood chemistry); however, the differences between the sexes were too small to
specifically identify females as more susceptible to the effects of MEK than males. No gender-
specific susceptibility was observed in offspring in any of the developmental studies (Schwetz et
al., 1974, 1991; Deacon et al., 1981). While a possible effect on male copulatory success was
noted in the Cox et al. (1975) multigeneration reproductive toxicity study, no effects on females
were apparent.
4.7.3. Other
The potential exists for increased susceptibility to neurotoxicity, hepatotoxicity, and renal
toxicity following exposure to MEK in combination with certain other solvents. MEK, for
example, potentiates the neurotoxicity of hexacarbon solvents (n-hexane, MnBK, and
2,5-hexanedione) (Saida et al., 1976; Couri et al., 1977) and the liver and kidney toxicity of
haloalkane solvents (carbon tetrachloride, trichloromethane, and chloroform) (Dietz and Traiger,
1979; WHO, 1992; Brown and Hewitt, 1984). Although the mode by which MEK potentiates
the neurotoxicity of hexacarbon solvents is not entirely clear, it appears to involve alterations in
their metabolism to toxic metabolites. Unlike other ketones, MEK is not metabolized to a
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gamma-diketone and is not, therefore, associated with distal neurofilamentous axonopathy. The
potentiating effects of MEK on the toxicity of other solvents have only been demonstrated at
relatively high exposure concentrations (200-1,000 ppm or 590-2,950 mg/m3).
No data are available concerning susceptibility of other specific lifestages (including
elderly populations) to MEK toxicity. No toxicologic basis exists, metabolic or otherwise, to
suspect that MEK is capable of exhibiting toxicity specific to these lifestages.
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5. DOSE RESPONSE ASSESSMENTS
5.1. ORAL REFERENCE DOSE (RfD)
5.1.1. Choice of Principal Study and Critical Effect
No studies examining the subchronic or chronic effects of oral exposure to MEK in
humans or experimental animals were identified. The repeat-dose oral toxicity data base is
limited to data for 2-butanol (a metabolic precursor) and 3-hydroxy-2-butanone (a metabolite).
The 2-butanol data consist of a 2-generation reproductive and developmental toxicity
study of 2-butanol in the rat (Cox et al., 1975). For 3-hydroxy-2-butanone, a 13-week drinking
water study in rats is available (Gaunt et al., 1972). No in vivo toxicity studies of repeat
exposure (by any route) to 2,3-butanediol (the other main metabolite of MEK) are available.
While the administration of 2-butanol in drinking water before and during gestation and lactation
at concentrations as high as 3% did not affect reproductive performance (with the possible
exception of effects on male copulatory success), it decreased pup survival and pup body weight
gain. A concentration of 2% in drinking water reduced fetal weights when pregnancies were
terminated on gestation day 20 and decreased pup body weight gains when dams were allowed to
deliver. The finding of developmental toxicity in rats exposed orally to 2-butanol is consistent
with similar findings in inhalation developmental toxicity studies of MEK (Schwetz et al., 1974,
1991; Deacon et al., 1981) and 2-butanol (Nelson et al., 1989, 1990) (see Table 5). Given these
observations, it is plausible that the developmental effects produced by 2-butanol and MEK are
caused by MEK or a subsequent metabolite common to both. The only other toxic effect
associated with long-term oral exposure to 2-butanol is renal lesions in male rats at 2% in
drinking water (3,384 mg/kg-day) (Cox et al., 1975).
While data from the 13-week drinking water study with 3-hydroxy-2-butanone in CFE
rats (Gaunt et al., 1972) suggest adverse hematological effects (decreased hemoglobin
concentration and red blood cell count), the effect was not observed in toxicity studies of 2-
butanol (Cox et al., 1975) or MEK (Cavender et al., 1983). The study concerning exposure to
3-hydroxy-2-butanone in drinking water provides no information regarding the potential for
developmental effects, which are the key effects seen with oral and inhalation exposure to 2-
butanol and inhalation exposure to MEK. Thus, the slight anemia produced by oral exposure to
3-hydroxy-2-butanone is inconsistent with the effects seen following inhalation exposure to
MEK, or oral or inhalation exposure 2-butanol. Hence, 3-hydroxy-2-butanone does not appear
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to be an appropriate surrogate for assessing the toxicity of MEK.
Pharmacokinetic and toxicologic data support the use of 2-butanol as an appropriate
surrogate for MEK. Pharmacokinetic findings in rats that support the use of 2-butanol as a
surrogate for MEK include: (1) orally administered 2-butanol was almost completely converted
to MEK and its metabolites within 16 hours; (2) peak MEK blood concentrations occurred at
similar times after administration of 1,776 mg/kg 2-butanol (7-8 hours) or 1,690 mg/kg MEK
(4-5 hours); and (3) common metabolites (3-hydroxy-2-butanone and 2,3-butanediol) were
formed and eliminated with similar kinetics after the administration of 2-butanol or MEK
(Traiger and Bruckner, 1976; Dietz et al., 1981). Comparable pharmacokinetic data for 2-
butanol and MEK in humans are not available; however, evidence for metabolic conversion of
MEK to 2-butanol in humans supports the assumption that rats and humans metabolize 2-butanol
similarly. As discussed in Section 4.3, toxicologic findings that support the use of 2-butanol as a
MEK surrogate include: (1) fetal weight deficits were critical effects in (a) studies of rats
(Schwetz et al., 1974; Deacon et al., 1981) and mice (Schwetz et al., 1991) exposed to MEK by
inhalation during gestation, (b) a two-generation reproductive and developmental toxicity study
in rats exposed to 2-butanol in drinking water (Cox et al., 1975), and (c) a study of rats exposed
to 2-butanol by inhalation during gestation (Nelson et al., 1989); and (2) the relationships
between air concentrations and the degree of fetal weight changes were consistent for MEK and
2-butanol.
Thus, the reproductive and developmental drinking water toxicity study of 2-butanol in
rats (Cox et al., 1975) was selected as the principal study for deriving an RfD for MEK. Cox et
al. (1975) also served as the principal study for the RfD of 0.6 mg/kg-day that was previously
entered in the IRIS data base in 1993. Developmental effects identified in this study included
decreased pup survival and decreased neonatal body weight in F1A pups whose parents were
exposed to 3% 2-butanol in drinking water before mating through day 10 of lactation. Decreased
body weights, with no effect on survival, were observed in FIB fetuses and F1A and F2 pups
that were exposed to 2% 2-butanol in drinking water (see Table 5).
5.1.2. Methods of Analysis
The RfD was derived using benchmark dose analysis of body weight data from offspring
in the rat multigeneration drinking water toxicity study of 2-butanol (Cox et al., 1975). Details
of the benchmark dose modeling results are presented in Appendices B-l through B-5.
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5.1.2.1. Benchmark Dose Modeling
The following data sets from the Cox et al. (1975) study were selected for benchmark
dose modeling: fetal weight data from the FIB generation, and postnatal day 4 and day 21 pup
weights from the F1A and F2 generations. Decreased F1A pup survival observed in the highest
dose group (3% solution) is likely to have confounded the effects on surviving pup body weight
(measured body weights represent survivors rather than all offspring born to 3% dams).
Consequently, this exposure level does not help identify a level of exposure at which a less
severe precursor of frank toxicity might occur. Because of this likely confounding, the modeling
of the F1A pup body weight data did not include data from the high-dose group. Survival of
fetuses or pups was not affected in any dose group in the FIB or F2 generations, so body weight
data from all dose groups (0, 0.3, 1, and 2%) were included in the modeling for these
generations.
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Table 5. Summary of key repeat-exposure reproductive and developmental toxicity studies in animals exposed
to MEK or 2-butanol
Study
Exposure protocol
Effects
LOAEL
NOAEL
Oral studies
Cox etal. (1975)
FO generation
Fl A (first litter)
FIB (second litter)
F2(F1 A offspring)
FDRL-Wistar rats (FO), ~30/grp,
exposed to 0, 0.3, 1, or 3% 2^
butanol in drinking water for 8
weeks prior to mating, during
F1A pregnancy and litter cast.3
Starting at F1A postnatal day 21a
through F2 gestation day 20: FO,
FIB and F2 received 0, 0.3, 1, or
2% 2-butanol in drinking water.
FO: Decreased body weight (3%).
F1A: Decreased pup survival (3%).
Decreased pup weight, days 4
and 21 (3%).
FIB: Decreased fetal weight (2%) .
F2: Decreased pup weight, days 4
and 2 1(2%).
[All pup and fetal weight comparisons
were based on litter means and not
individual pup/fetus data.]
3,122 mg/kg-day
(2%)- fetal/pup
body weight
1,771 mg/kg-day
(1%)
Inhalation studies
Schwetz etal. (1974)
Deacon etal. (1981)
also reported by Dow
Chemical Corporation
(1979)
Nelson etal. (1989,
1990)
Pregnant SD rats, 21-23/grp,
exposed to 0, 1,126, or 2,618
ppm MEK, 7 hours/dav on
gestation days 6-15.
Pregnant SD rats, 18-26/grp,
exposed to 0,412, 1,002, or
3,005 ppm MEK, 7 hours/dav on
gestation days 6-15.
Pregnant SD rats, 15-16/grp,
exposed to 0, 3,500, 5,000, or
7,000 ppm 2-butanol, 7 hours/dav
on gestation days 1-19.
Developmental effects (percentage of
litters with any soft tissue anomaly)
Decreased maternal weight gain.
Increased incidence fetal skeletal
variations.
Decreased maternal weight gain and
food consumption.
Decreased maternal locomotor activity.
Decreased fetal weight.
2,618 ppm
(7,723 mg/m3)
3,005 ppm
(8,865 mg/m3)
5,000 ppm
(15, 150 mg/m3)
1,126 ppm
(3,322 mg/m3)
1,002 ppm
(2,955 mg/m3)
3,500 ppm
(10,605 mg/m3)
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Table 5. Summary of key repeat-exposure reproductive and developmental toxicity studies in animals exposed
to MEK or 2-butanol
Study
Schwetzetal. (1991)
also reported as Mast
etal. (1989) and NTP
(1990)
Exposure protocol
Pregnant CD-I mice, 33/grp,
exposed to 0,398, 1,010, or
3,020 ppm MEK, 7 hours/dav on
gestation days 6-15.
Effects
Decreased body weight in male fetuses
and both sexes combined [based on
litter means].
Increased maternal liver-to-body
weight ratio.
LOAEL
3,020 ppm
(8,909 mg/m3)
NOAEL
1,010 ppm
(2,980 mg/m3)
a In the 3% group, FO dams and Fl A pups received drinking water with no 2-butanol between days 10 and 21 post partum. Thereafter, the concentration was
changed to 2%.
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Continuous data models (linear, polynomial, or power), either with a constant variance or
with variance as a power function of the mean value (using an additional model parameter), were
fit to the data using U.S. EPA Benchmark Dose Software (version 1.3.1). The software was used
to calculate potential points of departure for deriving the RfD by estimating the effective dose at
a specified level of response (EDX) and its 95% lower bound (LEDX). In the case of pup or fetal
body weight, there is no specific decrement that is generally regarded as indicative of an adverse
response. EPA's draft Benchmark Dose Technical Guidance Document (U.S. EPA, 2000c)
recommends, in the absence of some idea of the level of response to consider adverse, selecting
as the benchmark response (BMR) level for continuous data a change in the mean equal to one
control standard deviation from the control mean. Using data from Cox et al. (1975), one
standard deviation from the control mean resulted in BMDs that corresponded to body weights 9
to 26% below the control mean (see Tables 6, 7 and 9) - values generally above the range of
experimental data. Because an aim in BMD modeling is to select a BMD within the range of
observation, other measures of the BMR were examined. A 5% reduction in fetal/pup body
weight relative to the control was a response rate that fell within the range of experimental dose
levels in the Cox et al. (1975) study, and consequently was selected as the benchmark response
(BMR). In addition, an ED10 and LED10 for each endpoint were estimated as a consistent point
of comparison across chemicals, as recommended in the Benchmark Dose Technical Guidance
Document (U.S. EPA, 2000c). These additional measures are provided in Appendix B.
Modeling of F1A Pup Body Weights
The overall means of the individual litter means for Fl A pup body weights and their
standard deviations in the control and two lowest exposure groups were calculated from litter
data shown in Table 6.
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Table 6. Mean litter pup body weight in F1A generation Wistar rats exposed to
2-butanol in drinking water in a two-generation reproductive and developmental
toxicity study
Drinking water
concentration
(% 2-butanol by
weight)
0
0.3
1
Maternal
dose (mg/kg-
day)a
0
594
1,771
Mean of litter means pup
body weight postnatal day 4
(g ± standard deviation)15
[number of litters]
10.7±1.1 [29]
10.2±1.3 [27]
10.0±1.3 [30]
Mean of litter means pup
body weight postnatal day 21
(g ± standard deviation)15
[number of litters]
49±3.8 [28]
47±3.9 [27]
44±4.8 [30]
a Average daily intake of 2-butanol as reported by Cox et al. (1975).
b The data reported herein differ from the summary data in Table 3 of Cox et al. (1975) because data for day 21
could only be discerned to the nearest gram from the best available copy of the study report.
Source: Adapted from Appendix II of Cox et al. (1975).
A linear continuous-variable model using constant variance (BMDS version 1.3.1)
provided an adequate fit to the data (with a goodness-of-fit p value > 0.1) and was used to
establish the ED05 (See Appendix B for benchmark dose software output). Other continuous
variable models (polynomial and power) could not be fit to the data due to lack of degrees of
freedom since the number of dose groups in the modeled data set were equal to or less than the
number of parameters estimated in the models, and thus it was not possible to perform statistical
tests typically used to determine adequacy of model fit. Visual inspection of a plot of the
predicted and observed means also indicated a reasonable fit of the linear model to the data in
the range nearest the point of departure (see Appendix B, outputs B-l and B-2).
The model-predicted ED05 values associated with a 5% decrease in mean Fl A pup body
weight were 1,387 mg/kg-day for day 4 and 878 mg/kg-day for day 21. The corresponding
LED05 values were 803 mg/kg-day for day 4 and 657 mg/kg-day for day 21.
Modeling of FIB Fetal Weights
The overall means of the individual litter means for FIB fetal weights and their standard
deviations in the control and exposed groups were calculated from average fetus weight data for
each litter presented in Appendix III of the Cox et al. (1975) report and are shown in Table 7.
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Table 7. Litter mean fetal weight in FIB generation Wistar rats exposed to 2-butanol in
drinking water in a two-generation reproductive and developmental toxicity study
Drinking water
concentration
(% 2-butanol by
weight)
0
0.3
1
2
Maternal dose
(mg/kg-day)
Oa
594a
l,771a
3,122b
Number of
litters
29
27
30
29
Mean of litter
means fetal weight
(g)
4.14
4.16
4.38
3.74
Standard
deviation
1.45
0.69
1.04
1.01
"Average daily intake of 2-butanol as reported by the authors.
b Calculated based on a linear regression analysis of the reported average intakes and drinking water
concentrations of 2-butanol.
Source: Adapted from Appendix III of Cox et al. (1975).
A constant variance polynomial continuous-variable model (BMDS version 1.3.1)
provided the best fit to the data (as indicated by the lowest AIC with a goodness-of-fit p value >
0.1; see summary of goodness-of-fit statistics in Table 8) and was used to establish the ED05.
Fitting a model that described the variance as a power function of the mean value did not
improve the fit as indicated by the AIC. Visual inspection of a plot of the predicted and
observed means also indicated a reasonable fit of the polynomial model to the data in the range
nearest the point of departure (see Appendix B, output B-3).
Table 8. Benchmark dose modeling results using litter mean body weight data for FIB fetuses
Model
Linear
Polynomial
Power
GOFP
0.15
0.12
0.17
AIC
137.9
136.6
137.0
ED05 (mg/kg-day)
1,969
2,198
2,980
LED05 (mg/kg-day)
896
1,046
1,578
GOFP = Goodness-of-fit p-value for chi-square.
AIC = Akaike's Information Criterion.
ED05 = Benchmark dose calculated by BMDS associated with a 5% decrease in mean fetal weight.
LED05 = 95% lower confidence limit on the ED05 as calculated by BMDS.
Source: Adapted from Cox et al. (1975).
The model-predicted ED05 associated with a 5% decrease in mean FIB fetal weight was
2,198 mg/kg-day. The corresponding LED05 was 1,046 mg/kg-day.
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Modeling of F2 Pup Body Weights
The overall means of the individual litter means for F2 pup body weights at postnatal
days 4 and 21 and their standard deviations in the control and exposed groups were calculated
from litter averages presented in Appendix V of the Cox et al. (1975) report (see in Table 9).
Table 9. Litter mean pup body weight in F2 generation Wistar rats exposed to
2-butanol in drinking water in a two-generation reproductive and developmental
toxicity study
Drinking water
concentration
(% 2-butanol by
weight)
0
0.3
1
2
Maternal
dose (mg/kg-
day)
Oa
594a
1,771"
3,122b
Mean of litter means pup
body weight postnatal
day 4 (g ± standard
deviation)
10.0±1.4
9.7±1.6
9.6±2.3
9.5±1.6
Mean of litter means pup
body weight postnatal
day 21 (g ± standard
deviation)
40±6.1
39±7.8
39±9.4
35±4.7
"Average daily intake of 2-butanol as reported by the authors.
b Calculated based on a linear regression analysis of the reported average intakes and drinking water
concentrations of 2-butanol.
Source: Appendix V of Cox et al. (1975).
A linear continuous-variable model assuming constant variance (BMDS version 1.3.1)
provided the best fit to the day 4 pup body weight (as indicated by the lowest AIC with a
goodness-of-fit p value > 0.1), whereas a polynomial continuous-variable model assuming
constant variance provided the best fit to the day 21 pup body weight. Both models were used to
establish day 4 and day 21 ED05 values (see goodness-of-fit statistics in Table 10.) Fitting
models that described the variance as a power function of the group means did not improve the
fit as indicated by the AIC. Visual inspection of the plots of the predicted and observed means
also indicated a reasonable fit of the selected models to the data in the range nearest the point of
departure (see Appendix B, outputs B-4 and B-5).
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Table 10. Benchmark dose modeling results using litter mean body weight data for F2
pups on postnatal days 4 and 21
Model
GOFP
AIC
ED05 (mg/kg-day)
LED05 (mg/kg-day)
Postnatal day 4
Linear
Polynomial
Power
0.88
0.61
0.26
227.9
227.9
231.9
3,471
3,471
3,471
1,347
1,347
1,347
Postnatal day 21
Linear
Polynomial
Power
0.55
0.48
0.26
512.5
511.8
515.5
1,398
2,056
2,508
851
901
919
GOFP = Goodness-of-fit p-value for chi-square.
AIC = Akaike's Information Criterion.
ED05 = Benchmark dose calculated by BMDS associated with a 5% decrease in mean pup body weight.
LED05 = 95% lower confidence limit on the ED05 as calculated by BMDS.
Source: Adapted from Cox et al. (1975).
The model-predicted ED05 values associated with a 5% decrease in mean pup body
weight were 3,471 mg/kg-day for postnatal day 4 and 2,056 mg/kg-day for day 21. The
corresponding LED05 values were 1,347 and 901 mg/kg-day, respectively.
Comparison of Benchmark Dose Modeling Results
For oral exposure to 2-butanol, developmental effects on body weight from three
generations of the Cox et al. (1975) study were modeled, including: fetal weight from the FIB
generation, and pup body weight at postnatal days 4 and 21 from the F1A and F2 generations.
The LED05 values calculated from modeling these data sets are summarized in Table 11.
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Table 11. Benchmark doses for developmental effects in rats from various generations
and potential points of departure for the RfD
Endpoint
F1A pup body weight, postnatal day 4b
F1A pup body weight, postnatal day 21b
FIB fetal weight, gestation day 20
F2 pup body weight, postnatal day 4
F2 pup body weight, postnatal day 2 1
ED05a
(mg/kg-day)
1,387
878
2,198
3,471
2,056
LED05 a
(mg/kg-day)
803
657
1,046
1,347
901
a ED05: benchmark dose associated with a 5% decrement in litter mean pup or fetal weight compared with control
mean.
LED05: 95% lower confidence limit on the ED.
b The data modeled excluded the high dose (3%).
Source: Adapted from of Cox et al. (1975).
LED05 values from these data sets are within 2-fold of each other, suggesting that all the
modeling results are equally plausible. The lowest point of departure, based on decreased pup
body weight at postnatal day 21 in the F1A generation (LED05 = 657 mg/kg-day) was selected
for derivation of the RfD as the most health protective value.
5.1.2.2. Route-to-route Extrapolation
As an alternative to using 2-butanol data as a surrogate for MEK, consideration was
given to route-to-route extrapolation to derive oral doses from existing inhalation data for
developing an RfD for MEK. Deficiencies in the absorption data, however, preclude the
application of this method for MEK. In humans, the pulmonary retention value of 53% (±2%)
reported by Liira et al. (1988) is based on acute (4-hour) exposure to 200 ppm (590 mg/m3)
MEK. In rats, the pulmonary retention data at similar exposure concentrations (200 ppm in
humans compared to 180 ppm in rats) result from a longer period of exposure, 14 hours (Kessler
et al., 1988). The pharmacokinetic data for MEK indicate that pulmonary retention is
concentration-dependent (Liira et al., 1988), suggesting that absorption is limited by transport to
the metabolizing enzymes in the liver, rather than metabolic capacity. Therefore, it cannot be
assumed that the pulmonary retention value will be the same at exposures across a larger dose
range. Developmental effects of MEK are produced by concentrations that are an order of
magnitude greater than those used to calculate the rat pulmonary retention value. The toxicity of
MEK may be a result of exposure to concentrations that exceed of the capacity for detoxification
by a saturable enzyme mechanism. For this reason, it would be inappropriate to estimate the
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pulmonary retention value at the effect levels identified by the inhalation developmental toxicity
studies of MEK in rodents (Schwetz et al., 1974, 1991; Deacon et al., 1981), precluding the
derivation of an oral RfD for humans based on extrapolation from inhalation effects in animals.
Moreover, the route-to-route extrapolation would also require data on oral absorption of MEK in
humans, and such data are not available. Consequently, these deficiencies in the data preclude
route-to-route extrapolation as a basis for development of an oral RfD for MEK.
Rat PBPK models that include oral, inhalation, and parenteral portals of entry have been
developed recently (Thrall et al., 2002), but human PBPK models with both oral and inhalation
portals of entry have not yet been developed. When appropriate human PBPK models are
developed, the rat and human models could be used to estimate human oral exposure levels
associated with an appropriate internal dose surrogate from the inhalation exposure levels in the
rat developmental toxicity studies for MEK (Deacon et al., 1981; Schwetz et al., 1974).
5.1.3. RfD Derivation - Including Application of Uncertainty Factors
The LED05 of 657 mg/kg-day was used as the point of departure for calculating the RfD.
This point of departure is associated with a 5% decrease in mean postnatal day 21 body weight
of F1A Wistar rat pups exposed to 2-butanol in drinking water (Cox et al., 1975). Because 2-
butanol was used as a surrogate for MEK, a molar adjustment was performed to account for the
different molecular weights of the two chemicals. A total uncertainty factor of 1,000 was
applied to this adjusted point of departure: 10 for extrapolation from animals to humans, 10 for
extrapolation to the most susceptible humans, and 10 for data base deficiencies.
The following molar adjustment of the LED05 value was calculated to account for
differences in the molecular weights of 2-butanol and MEK:
657 mg/kg-day 2-butanol x 72.1066 g/mol MEK
74.1224 g/mol 2-butanol
= 639 mg/kg-day MEK
A 10-fold uncertainty factor was used to account for laboratory-animal to human
interspecies differences. No information is available on the toxicity of MEK in humans exposed
by the oral route. No other information is available to assess possible differences between
animals and humans in pharmacodynamic responses to MEK. Rat and human PBPK models for
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oral exposure to MEK could potentially be used to decrease pharmacokinetic uncertainty in
extrapolating from rats to humans, but such models are not currently available.
A 10-fold uncertainty factor for intraspecies differences was used to account for
potentially susceptible human subpopulations. In the absence of information on the variability in
response of humans to MEK exposure, the default value of 10 was used.
A 10-fold uncertainty factor was used to account for deficiencies in the available MEK
data base. While no oral data are available for MEK, the available pharmacokinetic and
inhalation toxicity data support 2-butanol as an appropriate surrogate for MEK. Nonetheless, the
use of 2-butanol data to estimate the toxicity associated with MEK exposure introduces some
uncertainty in the assessment. Although no chronic studies are available, the data base includes
a two-generation reproductive and developmental toxicity assay wherein rats were exposed to 2-
butanol for 14-18 weeks with observed effects limited to reductions in body weight and
histopathologic changes in the kidney of male rats only. The absence of any other organ-specific
toxicity following a 14-18 week exposure to 2-butanol reduces the uncertainty associated with
the lack of chronic toxicity data for MEK or 2-butanol.
An uncertainty factor to extrapolate from a LOAEL to a NOAEL was not necessary
because BMD modeling was used to determine the point of departure. The dose corresponding
to a 5% decrease in pup weight, relative to control, was selected as the point of departure. There
is no specific decrement in fetal/pup weight that is generally recognized as indicative of an
adverse effect. Further, there were no other effects in the range of the LED05 of 657 mg/kg-day
2-butanol. Therefore, no further adjustments were considered for identifying a level of oral
exposure to MEK associated with a minimal level of risk.
Consistent with EPA practice (U.S. EPA, 1991a), an uncertainty factor was not used to
account for extrapolation from less than chronic results because developmental toxicity
(decreased pup body weight following in utero and neonatal exposure) was used as the critical
effect. The developmental period is recognized as a susceptible lifestage where exposure during
certain time windows are more relevant to the induction of developmental effects than lifetime
exposure.
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The RfD for MEK was calculated as follows:
RfD = LED05 -H UF
= 639 mg/kg-day - 1000
= 0.6 mg/kg-day
5.1.4. Previous Oral Assessment
In the 1993 IRIS assessment of MEK, an RfD of 0.6 mg/kg-day was derived based on the
NOAEL of 1,771 mg/kg-day for decreased fetal birth weight in the FIB generation of Wistar rats
in the multigeneration drinking water study with 2-butanol by Cox et al. (1975). The 1993
assessment stated that "a combined uncertainty factor of 3000 was applied to account for four
uncertainty factors assigned 10 for each factor, including: 10 for inter- and intraspecies
extrapolations, 10 to adjust for subchronic-to-chronic extrapolation since long-term effects in the
dams were not reported in the principal study; 10 for an incomplete data base that included a
lack of both subchronic and chronic oral exposure studies for MEK; and 10 for lack of data for a
second rodent species for either MEK or 2-butanol."
The 1993 and current RfDs differ in the approach used to derive the reference values and
in the application of UFs. The current RfD was derived using BMD methods rather than the
NOAEL as a point of departure. The UFs applied in the 1993 assessment did not conform with
current practices. In particular, it is not current Agency practice to apply a subchronic to chronic
uncertainty factor where the point of departure is based on developmental toxicity. Thus, the
total UF used in the 1993 assessment was 3,000, whereas the total UF applied in the current
assessment is 1,000.
5.2. INHALATION REFERENCE CONCENTRATION (RfC)
5.2.1. Choice of Principal Study and Critical Effect
Several studies examining the health effects of inhalation exposure to MEK exist in
experimentally- and occupationally-exposed humans as well as in experimental animals;
however, many of these are inappropriate for use in dose-response assessment. For example,
many occupational studies are complicated by insufficient data on exposure levels (duration and
concentration) and potential simultaneous exposure to other solvents (often to hexacarbon
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solvents for which MEK may potentiate toxicity). Therefore, it is not possible to identify effect
levels from the available occupational reports for dose-response assessment. As with other small
molecular weight, aliphatic or aromatic chemicals, acute exposure to high concentrations of
MEK results in reversible central nervous system depression. Evidence for this effect in humans
is limited to a few case reports involving combined exposure to MEK and toluene (Welch et al.,
1991; Seaton et al., 1992; Callender, 1995; Orti-Pareja et al., 1996). The only other human data
are from a series of studies involving acute, 4-hour exposures of volunteers (Dick et al., 1984,
1988, 1989, 1992) wherein no exposure-related changes were reported for performance on
psychomotor and mood tests or incidences of irritation.
As discussed in Section 4.5.2, the range of toxic effects in animals resulting from
inhalation exposure to MEK indicates that developmental effects are the most sensitive
toxicologically-relevant endpoints. Inhalation exposure of experimental animals to 3,000 ppm
MEK (7 hours/day on days 6-15 of gestation) resulted in developmental effects in rats and mice
(Schwetz et al., 1974, 1991; Deacon et al., 1981). The most appropriate data for the derivation
of an inhalation RfC for MEK are from inhalation developmental toxicity studies in rats (Deacon
et al., 1981) and mice (Schwetz et al., 1991). The original laboratory reports are available for
both, and the effect levels for the reported developmental effects are consistent (although the
specific endpoints differ). In Sprague-Dawley rats, Deacon et al. (1981) reported fetal toxicity
(increased incidence of skeletal variations) at 3,005 ppm (8,865 mg/m3) MEK 7 hours/day on
gestation days 6-15. In CD-I mice, Schwetz et al. (1991) found reduced fetal weight at 3,020
ppm (8,909 mg/m3) MEK 7 hours/day on gestation days 6-15, and a positive trend for increasing
the incidence of fetuses with misaligned sternebrae. In each case, fetal effects were
accompanied by slight maternal toxicity.
5.2.2. Methods of Analysis
The RfC was derived using benchmark analysis of developmental effects for rats and
mice exposed to MEK during gestation (Deacon et al., 1981; Schwetz et al., 1991). NOAELs of
1,002 and 1,010 ppm (2,955 and 2,980 mg/m3) and LOAELs of 3,005 and 3,020 ppm (8,865 and
8,909 mg/m3) were established for rats and mice, respectively. Details of the benchmark dose
modeling results are presented in Appendix B.
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5.2.2.1. Benchmark Dose Modeling
In Sprague-Dawley rats, Deacon et al. (1981) reported a statistically significant increase
in the incidence of litters with fetuses with extra ribs. In CD-I mice, Schwetz et al. (1991)
identified two statistically significant developmental effects in fetuses exposed to MEK:
decreased fetal weight per litter (continuous data) and a trend for increasing the incidence of
fetuses with misaligned sternebrae with increasing exposure level (dichotomous data). Data
from each of these three endpoints have been analyzed by benchmark dose methods and
examined for toxicological relevance.
Modeling of Incidence of Rat Litters with Fetuses with Extra Ribs
The incidence of extra ribs (litters with an affected fetus) as reported by Deacon et al.
(1981) is shown in Table 12. The incidence of extra ribs in fetal Sprague-Dawley rats exposed
to 3,005 ppm MEK 7 hours/day on gestation days 6-15 was statistically different from control.
Table 12. Incidence of extra ribs (litters with an affected fetus) in Sprague-Dawley rats
exposed to MEK 7 hours/day on gestation days 6-15
Concentration
(ppm)
0
412
1,002
3,005
Number of
fetuses (litters)
329 (26)
237(19)
226 (19)
229 (18)
Number of litters with
fetuses with extra ribs
2(2)
0
0
7(6)
Mean percent of fetuses
with extra ribs per litter
0.6 ±0.3
0±0
0±0
3.1 ±1.8
Source: Adapted from Deacon et al. (1981).
All nested models for dichotomous variables available in EPA's Benchmark Dose
Software (BMDS version 1.3.1) were fit to the data in Table 12. The models - the nested
logistic (NLogistic), NCTR, and Rai and vanRyzin models - allow for the possibility that the
variance among the proportions of pups affected in individual litters is greater than would be
expected if the pups were responding completely independently of each other (U.S. EPA,
2000c). A 5% increase in the incidence of extra ribs was selected as the benchmark response
because it was a response rate that fell within the range of experimental dose levels used in the
Deacon et al. (1981) study. All of the models provided similar fits to the data, based on the
summary results reported in the BMDS output and the detailed examination of graphs and
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goodness-of-fit statistics (summarized in Table 13). Model fits were not improved by the
incorporation of litter size (as a litter-specific covariate) or intra-litter correlations, as determined
by comparisons of AIC values. Since the fits were quite similar, only one set of model output
(the NCTR model, fitting only slightly better than the others) is provided in Appendix B, output
B-6.
Table 13. Benchmark concentration modeling results using litter incidence data for
Sprague-Dawley rat fetuses with extra ribs exposed to MEK during gestation days 6-15
Nested Model
Log-Logistic
NCTR
Rai and vanRyzin
GOFP
0.09
0.51
0.51
AIC
96.5
96.5
96.6
EC05 (ppm)
3,124
3,317
3,353
LEC05 (ppm)
2,993
2,993
2,992
GOFP = Goodness-of-fit p-value for chi-square.
AIC = Akaike's Information Criterion.
EC05 = Benchmark concentration calculated by BMDS associated with a 5% extra risk of affected fetuses per
litter.
LEC05 = 95% lower confidence limit on the EC05 as calculated by BMDS.
Source: Adapted from Deacon et al. (1981).
The model-predicted EC05 value associated with a 5% increased incidence of extra ribs
was 3,317 ppm. The corresponding LEC05 was 2,993 ppm.
Modeling of Decreased Fetal Weight Data in Mice
The full laboratory report from Schwetz et al. (1991) is available in Mast et al. (1989).
The overall means of the individual litter means for fetal weight and their standard deviations in
the control and MEK-exposed groups are shown in Table 14 below.
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Table 14. Litter mean fetal weight (both sexes combined) in CD-I mice exposed to
MEK 7 hours/day on gestation days 6-15
Concentration (ppm)
0
398
1,010
3,020
Number of
litters
26
23
26
28
Fetal weight (mean of litter
means) ( g)
1.35
1.35
1.33
1.29
Standard
deviation
0.07
0.06
0.07
0.08
Sources: Adapted from Mast et al. (1989) and Schwetz et al. (1991).
Models for continuous data (linear, polynomial, and power), either with constant variance
or with variance as a power function of the mean value (using an additional model parameter),
were fit to the data in Table 14 using EPA's Benchmark Dose Software (BMDS version 1.3.1).
Since there is no specific decrement in fetal weight that is generally regarded as indicative of an
adverse effect, a decrease in the mean fetal weight of 1 standard deviation of the control mean
was selected as the benchmark response for this endpoint consistent with the recommendations
of EPA's Benchmark Dose Technical Guidance Document (U.S. EPA, 2000c). This benchmark
response corresponds to a 5% decrease in the mean control group weight for this data set. A
linear continuous-variable model assuming constant variance (BMDS version 1.3.1) provided the
best fit to the data (as indicated by the lowest AIC with a goodness-of-fit p value > 0.1; see
summary of goodness-of-fit statistics in Table 15.) Visual inspection of the graph of the
predicted and observed means also indicated a reasonable fit of the selected model to the data in
the range nearest the point of departure (see Appendix B, output B-7).
Table 15. Benchmark concentration modeling results using litter mean body weight data
Model
Linear
Polynomial
Power
GOFP
0.90
0.66
0.28
AIC
-442.2
-440.2
-438.2
EC (ppm MEK)
3,339
3,330
3,343
LEC (ppm MEK)
2,273
2,273
2,275
GOFP = Goodness-of-fit p-value for chi-square.
AIC = Akaike's Information Criterion.
EC = Benchmark concentration calculated by BMDS associated with a mean fetal weight 1 SD below the control
mean.
LEC = 95% lower confidence limit on the EC as calculated by BMDS.
Sources: Adapted from Mast et al. (1989) and Schwetz et al. (1991).
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The model-predicted EC associated with a mean fetal weight of 1 standard deviation
below the control mean was 3,339 ppm MEK. The corresponding LEG was 2,273 ppm MEK.
Modeling of Misaligned Sternebrae Data in Mice
The other statistically significant effect identified by Schwetz et al. (1991) was an
increased incidence of misaligned sternebrae in CD-I mouse fetuses exposed to MEK. The
complete laboratory report from Schwetz et al. (1991) is available in Mast et al. (1989). A
summary of the incidence of misaligned sternebrae for individual fetuses (Mast et al., 1989) is
shown in Table 16.
Table 16. Total number of fetuses (combined for both sexes) with misaligned
sternebrae per exposure group in CD-I mice exposed to MEK 7 hours/day on gestation
days 6-15
Concentration
(ppm)
0
398
1,010
3,020
Number of fetuses
(litters)
310(26)
260 (23)
291 (26)
323 (28)
Number of fetuses
(litters) with
misaligned sternebrae
31(18)
27(14)
49(18)
58(21)
Mean percent of fetuses
with misaligned
sternebrae per litter
(mean ± SD)
9.7 ±10.4
9.8 ±11.2
17.4 ±16.7
17.5 ±14.9
Sources: Adapted from Mast et al. (1989) and Schwetz et al. (1991).
The nested, dichotomous-variable models available in BMDS version 1.3.1 were fit to the
individual litter data for fetuses with misaligned sternebrae as reported in Appendix F of Mast et
al. (1989). Each model was fit with and without litter size as a covariate. Including litter size as
a covariate made very little difference in the goodness-of-fit statistics, indicating that litter size
was not a significant explanatory variable for changes in the incidence of misaligned sternebrae
(results not shown). Then each model was fit with and without intra-litter correlations. In each
case, the model fit was linear and was better with the intralitter correlations included. All three
nested models provided adequate fits to the data, based on the summary results reported in the
BMDS output (see Appendix B, output 8). A more detailed examination of the graphs and
residuals suggested that a nonlinear model should be considered, since the low- and mid-dose
responses were not fitted by the models as closely as the high-dose response. Allowing the
power parameter in each model to take a value less than one increased the AIC value for each
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model. Therefore, the linear versions already fitted were used. A 10% extra risk for misaligned
sternebrae was selected as the benchmark response since the model and data are most consistent
in this range of the data set. Also, EP A's Benchmark Dose Technical Guidance Document (U.S.
EPA, 2000c) recommends estimation of a 10% BMR for a point of consistent comparison across
chemicals. Further, the nested model did not provide a useful estimate of the lower bound for a
BMR of 5% (the lower bound on the BMC was estimated as essentially zero). Effective
concentrations associated with this BMR and their 95% lower confidence limits (LEC10) are
summarized in Table 17.
Table 17. Benchmark concentration modeling results using individual litter data for
mouse fetuses with misaligned sternebrae exposed to MEK during gestation days 6-15
(without litter size as covariates)
Nested Model
NLOGISTIC
Rai and Van Ryzin
NCTR
GOFP
0.6349
0.5433
0.4877
AIC
937.1
937.2
937.2
EC10
(ppm)
3,197
3,222
3,222
LEC10
(ppm)
1,714
1,789
1,789
GOFP = Goodness-of-fit p-value for chi-square.
AIC = Akaike's Information Criterion.
EC10 = Concentration associated with a 10% extra risk for misaligned sternebrae in fetuses.
LEC10 = 95% lower confidence limit on the EC10.
Source: Adapted from Mast et al. (1989).
Because the three model fits were very similar (Tablel7), an average of the three LEC10
values was calculated as the point of departure. The respective EC10 and LEC10 values calculated
as an average of the three models are 3,214 and 1,764 ppm, respectively.
Comparison of Benchmark Dose Modeling Results
For inhalation exposure to MEK, the following three developmental endpoints from two
species were evaluated: increased incidence of extra ribs in Sprague-Dawley rats (Deacon et al.,
1981), decreased fetal weight, and increased incidence of misaligned sternebrae in CD-I mice
(Schwetz et al., 1991). The EC and LEG values for these developmental endpoints are
summarized in Table 18. Benchmark modeling of the data produced similar points of departure
for the three developmental endpoints observed in the two species (within 2-fold). The lowest
point of departure of 1,764 ppm (5,202 mg/m3) was based on the incidence of misaligned
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sternebrae in mice exposed to MEK by inhalation for 7 hours/day on gestation days 6-15 and
was therefore selected as the most health protective value for derivation of the RfC.
Table 18. Benchmark concentrations for developmental effects in mice and rats and
potential points of departure for the RfC
Endpoint
Increased incidence of extra ribs (rats)
(Deacon et al., 1981)
Decreased fetal weight (mice)
(Schwetz et al., 1991)
Increased incidence of misaligned sternebrae (mice)
(Schwetz et al., 1991)
Benchmark
Response
Level
5%
1 s.d. = 5%
10%
EC, mg/m3
(ppm)a
9,781
(3,317)
9,847
(3,339)
9,478
(3,214)
LEC, mg/m3
(ppm)a
8,826
(2,993)
6,703
(2,273)
5,202
(1,764)
a Sample calculation: (3,317 ppm x 72.1 mg/mmol)/24.45 = 9,781 mg/m3, assuming 25°C and 760 mm Hg.
5.2.2.2. Adjustment to a Human Equivalent Exposure Concentration
By definition, the RfC is intended to apply to continuous lifetime exposures to humans
(U.S. EPA, 1994b). Because the RfC values are often derived from studies using intermittent
and less-than-lifetime exposures, EPA has established guidance (U.S. EPA, 1994b) for adjusting
the exposures to an appropriate human equivalent via a simple concentration (C) x time (t)
relationship (e.g., 8 hours @ 300 ppm = 24 hours @ 100 ppm). For developmental studies, the
Guidelines for Developmental Toxicity Risk Assessment (U.S. EPA, 199 la) and the Reproductive
Toxicity Risk Assessment Guidelines (U.S. EPA, 1996) note that peak exposure may be a more
relevant exposure metric for short half-life compounds, because the toxic effects may be due to
absolute concentration at a specific critical period during fetal development. Some more recent
studies suggest that area under the curve (AUC), the assumption underlying the C x t
relationship, may be a more appropriate metric for some developmental toxicants than peak
exposure. The latter has been demonstrated for certain agents with a short half-life in the body
(U.S. EPA, 2002). In consideration of this information, EPA recommends that adjusted
continuous exposures be used for inhalation developmental toxicity studies as for other health
effects from inhalation exposure (U.S. EPA, 2002).
Duration adjustment is appropriate as the more health-protective procedure, unless there
are pharmacokinetic data suggesting that the adjustment to a continuous exposure equivalent is
inappropriate, or mode of action information suggests that a susceptible period of development is
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specifically targeted (which would suggest that the peak dose may represent the effective dose).
In applying these considerations to MEK, the critical effect is nonspecific developmental toxicity
(developmental delays and variants), which suggests that duration adjustment may be
appropriate. On the other hand, the available pharmacokinetic data indicate that MEK is rapidly
absorbed, distributed, and metabolized, suggesting that duration adjustment may be less
appropriate than peak exposure. Overall, the available pharmacokinetic, pharmacodynamic, and
mechanism of action data for MEK do not provide sufficient evidence to support the use of
either peak exposure level or AUC as the most appropriate metric for internal effective dose.
Thus, it is appropriate to apply a health-protective duration adjustment to time-weight the
intermittent exposures used in the principal study. The LEG of 1,764 ppm (5,202 mg/m3) for
increased incidence of misaligned sternebrae in mice exposed to MEK (7 hours/day on days
6-15 of gestation) as reported by Schwetz et al. (1991) is adjusted from an intermittent exposure
to continuous exposure (7 hours/day to 24 hours/day) as follows:
7 hours I day
LEC(ADJ) = LECx
24 hours I day
= 5,202 mg/m3 x 7/24
= 1,517 mg/m3
The RfC methodology provides a procedure for estimating the human concentration that
corresponds to a given animal exposure concentration, i.e., the human equivalent concentration
or HEC. Because the critical effect of MEK is extrarespiratory, it is appropriate to apply a factor
to account for species differences in blood:air partition coefficients, assuming periodicity was
attained (i.e., the ratio of the coefficients). According to EPA's RfC guidelines (U.S. EPA,
1994b), MEK is a category 3 gas because it is not active in the respiratory tract, is rapidly
transferred between the lungs and blood, and the effects of inhalation exposure are extra-
pulmonary. In humans, reported mean blood:air partition coefficients for MEK from three
studies range from 125 to 202. The value of 125 was reported by Fiserova-Bergerova and Diaz
(1986) using blood collected directly from human volunteers (n=5) and processed immediately.
Perbellini et al. (1984) reported a blood:air partition coefficient of 183 based on blood collected
from two cadavers (delay in blood sample collection and preservation procedures were not
reported), and Sato and Nakajima (1979) reported a blood:air partition coefficient of 202 based
on preserved blood (n=5) collected from a blood bank. Because the blood:air partition
coefficient reported by Fiserova-Bergerova and Diaz (1986) was derived from samples that were
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subject to immediate and minimal processing, most closely resembling the sample processing in
test species, the human blood:air partition coefficient was estimated as 125. In the rat, the
blood:air partition coefficients for MEK have been reported as 138 to 139 (Thrall et al., 2002).
The RfC methodology stipulates that where the animal blood:air partition coefficient is greater
than the human coefficient, a value of one is used for the ratio (U.S. EPA, 1994b). Therefore,
the rat LEC(ADJ) is adjusted to a LEC(HEC) following the default procedure in the guidelines (U.S.
EPA, 1994b) as follows:
Blood: Air Partition Coefficient rat
T T^/^l "S"S
(HEC) = LhC (ADJ) X
Blood: AirPartition Coefficient human
= 1,517 mg/m3 x 1
= 1,517 mg/m3
The LEG (HEC) value of 1,517 mg/m3 for a 10% extra risk of misaligned sternebrae is used
to derive the RfC for MEK.
5.2.2.3. PBPK Modeling
Alternatively, PBPK modeling may be used to reduce uncertainty in the RfC resulting
from extrapolating from mice or rats to humans. PBPK models for rats (Dietz et al., 1981; Thrall
et al., 2002) and humans (Liira et al., 1990b) have been developed to describe the kinetics of
MEK in blood. The existing human model (Liira et al., 1990b) is limited in that it includes only
inhalation as a portal of entry, it was developed based on data from two healthy males, and
comparisons of model predictions with data from other human subjects are not available. With
sufficient model validation, the rat model from Thrall et al. (2002), for which the code is
available, could be used to estimate human equivalent concentrations corresponding to the
benchmark doses developed from the rat inhalation developmental toxicity study by Deacon et
al. (1981). No mouse PBPK model has been developed, however, precluding calculation of
chemical-specific human equivalent concentrations from the mouse inhalation developmental
toxicity study by Mast et al. (1989).
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5.2.3. RfC Derivation — Including Application of Uncertainty Factors
The LEC^EC) of 1,517 mg/m3 (associated with a 10% extra risk of misaligned sternebrae
in CD-I mice exposed to MEK by inhalation 7 hours/day on days 6-15 of gestation; Schwetz et
al., 1991) was used as the point of departure for calculating the RfC. A total uncertainty factor
(UF) of 300 was applied to this point of departure: 3 for interspecies extrapolation, 10 for
susceptible individuals, and 10 for an incomplete data base.
A 3-fold uncertainty factor was used for interspecies extrapolation, since this factor
embodies two areas of uncertainty: pharmacokinetics and pharmacodynamics. In this
assessment, the pharmacokinetic component is addressed by the calculation of the human
equivalent concentration (HEC) according to the procedures in the RfC methodology (U.S. EPA,
1994b). Accordingly, only the pharmacodynamic area of uncertainty remains as a partial factor
for interspecies uncertainty (1005 or approximately 3).
A 10-fold uncertainty factor for intraspecies differences was used to account for
potentially susceptible individuals within the human population. In the absence of information
on the variability in response of humans to MEK exposure, the default value of 10 was used.
Consistent with EPA practice (U.S. EPA, 1991a), an uncertainty factor was not used to
account for the extrapolation from less than chronic results because developmental toxicity
resulting from a narrow period of exposure (gestation days 6-15) was used as the critical effect.
The developmental period is recognized as a susceptible lifestage when exposure during certain
time windows of development are more relevant to the induction of developmental effects than
lifetime exposure.
A 10-fold uncertainty factor was used to account for data base deficiencies. As noted
earlier, the minimum data base requirements for deriving an RfC are satisfied by the Cavender et
al. (1983) study. Inhalation developmental toxicity studies are available in rats and mice
(Deacon et al., 1981; Schwetz et al., 1991). The data base lacks a chronic inhalation toxicity
study and multigeneration reproductive toxicity study. Neurotoxicity is adequately addressed by
the subchronic inhalation study of Cavender et al. (1983), in which animals were examined for
both neurological function and for central nervous system lesions with special neuropathological
procedures. The results from this study indicate that MEK has little, if any, neurotoxic potential
by itself when tested in adult laboratory animals under conditions of high-level repeated
inhalation exposure. Consistent with this finding is a lack of mechanistic evidence for
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neurotoxicity. The MEK data base does not, however, specifically include a developmental
neurotoxicity study.
A UF for extrapolation from a LOAEL to a NOAEL was not necessary because BMD
modeling was used to determine the point of departure. The exposure concentration
corresponding to a 10% extra risk of misaligned sternebrae in CD-I mice (Schwetz et al., 1991)
was selected as the point of departure. There is no specific level of extra risk for a skeletal
variant that is generally regarded as indicative of an adverse effect. Further, there were no
effects in the range of the LECj^c of 1,517 mg/m3 other than those that appeared to be related to
a general delay in growth. Therefore, no further adjustments were considered for identifying a
level of inhalation exposure to MEK associated with a minimal level of risk.
An RfC for MEK is calculated as follows:
RfC -- LEC(HEC)+ UF
= 1,517 mgI'm3 + 300
= 5 mg/m3
As noted in Section 5.2.4., the previous MEK RfC of 1 mg/m3 incorporated a modifying
factor of 3 to account for the lack of unequivocal data for respiratory tract (portal-of-entry)
effects suggested by earlier studies with MEK (Altenkirch et al., 1978). More recent data
concerning the portal-of-entry effects from MEK address the applicability of a modifying factor
to this assessment. Dick et al. (1984, 1988, 1989, 1992) found no evidence of a statistically
significant increase in respiratory tract irritation among humans who were exposed to MEK at
590 mg/m3 for 4 hours. In addition, Oleru and Onyekwere (1992) found no statistically
significant pulmonary effects among MEK-exposed leather workers (mean duration of
employment was approximately 10 years). While these studies do not directly address the
potential for portal-of-entry effects for MEK in continuous lifetime exposure scenarios as animal
studies evaluating histology would, they do address the concerns raised in the 1992 IRIS
assessment. Accordingly, a separate modifying factor to account for possible portal-of-entry
effects is not included in this assessment.
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5.2.4. Previous Inhalation Assessment
The previous RfC for MEK of 1 mg/m3 was entered into IRIS in 1992, prior to the
publication of EPA's RfC methodology (U.S. EPA, 1994b). The RfC was based on the Schwetz
et al. (1991) developmental toxicity study in the mouse. In the previous assessment, a combined
uncertainty factor of 3,000 was applied to a human equivalent concentration NOAEL (2,978
mg/m3), which was not adjusted to a continuous exposure basis. The combined uncertainty
factor accounted for interspecies extrapolation, intrahuman variability, and data base deficiencies
(including a lack of chronic and reproductive toxicity studies). A modifying factor of 3
accounted for the lack of unequivocal data for respiratory tract (portal-of-entry) effects suggested
by earlier studies with MEK (Altenkirch et al., 1978).
The current RfC is based on a BMD approach, rather than the NOAEL/LOAEL approach
used previously, and a combined UF of 300 rather than 3,000. Difference in the UF values are
accounted for by the calculation of an HEC according to the procedures in the RfC methodology,
which supports an interspecies UF of 3 rather than 10, and by more recent data that addressed
earlier portal-of-entry concerns, so that a modifying factor of 3 is no longer considered
necessary.
5.3. CANCER ASSESSMENT
5.3.1. Oral Slope Factor
Not applicable.
5.3.2. Inhalation Unit Risk
Not applicable.
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6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF
HAZARD AND DOSE RESPONSE
6.1. HUMAN HAZARD POTENTIAL
Methyl ethyl ketone (MEK, CASRN 78-93-3) has the chemical formula C4H8O
(structural formula CH3COCH2CH3) and a molecular weight of 72.11 g/mole. At room
temperature, MEK is a clear liquid with a sharp, mint-like odor. MEK is flammable, with a flash
point of-3°C. MEK is strongly reactive with a number of chemical classes, particularly strong
oxidizers. MEK is used as a solvent in the application of protective coatings and adhesives, as a
paint remover, and in cleaning fluids. MEK is a natural component of many foods, and may also
be found in soil and water in the vicinity of some hazardous waste sites. Other sources of
potential exposure include drinking water, tobacco smoke, and volatile releases from building
materials and consumer products (ATSDR, 1992).
Studies of the toxicokinetics of MEK reveal that the chemical is well absorbed by oral
and inhalation routes, does not appear to accumulate in tissues, and undergoes relatively rapid
clearance (on the order of hours) from the body, largely as a result of metabolism. MEK has
been shown to induce microsomal P450 activity.
In general, the available human data do not produce a definitive picture of the possible
adverse effects of long-term human exposure to MEK. Short-term inhalation exposure (4 hours)
to MEK under experimental conditions at or near 200 ppm (590 mg/m3) does not appear to pose
an increased risk of neurologic or irritation symptoms (Dick et al., 1984, 1988, 1989, 1992).
Although some evidence of persistent neurotoxicity is available from case reports of repeated
exposure (especially when MEK exposure occurs in combination with other solvents), the case
for a persistent neurotoxic effect of MEK exposure is not well supported in animal studies that
have focused on the possible neurotoxicity of MEK, including the development of peripheral and
central nerve fiber degeneration. Saida et al. (1976) found no evidence of peripheral neuropathy
(as indicated by paralysis) following continuous exposure of 12 Sprague-Dawley rats to 1,125
ppm (3,318 mg/m3) MEK for periods of 16 to 55 days. Cavender et al. (1983) found no
neurological effects in special neuropathological studies of the medulla (a portion of the brain)
and sciatic and tibial nerves of rats exposed to MEK at concentrations up to 5,041 ppm (14,870
mg/m3) for 90 days. Takeuchi et al. (1983) exposed male Wistar rats (8 per group) to 200 ppm
(590 mg/m3) MEK 12 hours/day for 24 weeks and found no evidence of a persistent effect on
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motor or mixed nerve conduction velocity, distal motor nerve latency, or histopathological
lesions of tail nerves. Couri et al. (1974) exposed 4 cats, 4 rats, 5 mice, and an unknown number
of chickens to 1,500 ppm (4,425 mg/m3) MEK 24 hours/day, 7 days/week for 7-9 weeks with no
apparent adverse neurologic effects.
In experimental animals, the longest exposure study available for characterizing the
health effects of repeated exposure to MEK is the 90-day inhalation study by Cavender et al.
(1983), wherein no toxicity could be attributed to MEK at concentrations as high as 2,518 ppm
(7,430 mg/m3). A two-generation reproductive and developmental toxicity study of Wistar rats
exposed to 2-butanol, a metabolic precursor of MEK, in drinking water, reported no clear
reproductive effects, but found body weight deficits in offspring and kidney histopathologic
lesions in adult male rats at estimated dose levels of approximately 3,000 mg/kg-day (Cox et al.,
1975). In addition, several developmental toxicity studies of rodents (exposed by inhalation 6-7
hours/day during gestation) reported reduced fetal weight and increased skeletal variations at
exposure levels of approximately 1,000 ppm (3,000 mg/m3) MEK (Schwetz et al., 1974, 1991;
Deacon et al., 1981). In the absence of chronic toxicity information for MEK by any route of
exposure, the effects of lifetime exposure to MEK must necessarily remain somewhat uncertain.
Available animal data consistently identify developmental effects in animals exposed to
relatively high levels of MEK. It is therefore reasonable and prudent to state that MEK is a
possible health hazard to humans who are repeatedly exposed to relatively high levels of MEK.
According to EPA's draft revised cancer guidelines (U.S. EPA, 1999), the hazard
descriptor "data are inadequate for an assessment of human carcinogenic potentiar is
appropriate for MEK because cancer studies of humans chronically exposed to MEK are
inconclusive, MEK has not been tested for carcinogenicity in animals by the oral or inhalation
routes, and the majority of short-term genotoxicity testing of MEK has demonstrated no activity.
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6.2. DOSE RESPONSE
6.2.1. Noncancer/Oral
There are no chronic or subchronic oral dose-response data for MEK in humans or
animals. The only relevant data for the oral RfD assessment are derived from a study with 2-
butanol, a metabolic precursor of MEK. The multigeneration reproductive and developmental
toxicity drinking water study by Cox et al. (1975) reported decreased F1A and F2 pup body
weights and decreased FIB fetal weights associated with 2-butanol exposure. Benchmark dose
modeling of F1A pup body weight data (mean of litter means) at postnatal day 21 yields a point
of departure (LED05) of 657 mg/kg-day for 2-butanol (i.e., the lower 95% confidence limit on a
dose producing a mean 5% decrease in body weight compared with control). Molar adjustment
to account for differences in the molecular weights of 2-butanol and MEK yields a point of
departure of 639 mg/kg-day. A combined uncertainty factor of 1,000 was applied to the point of
departure and a chronic RfD of 0.6 mg/kg-day was derived. This RfD is the same as the RfD
from the previous 1993 IRIS assessment. Confidence in the principal study is medium to low.
Although the study was adequately-conducted and the critical effect demonstrated therein was
supported by inhalation studies with MEK, a metabolic surrogate was used in place of MEK and
the highest drinking water concentration was reduced during the study resulting in a need to
estimate the actual exposure dose. Furthermore, certain parameters routinely evaluated in
studies of more current design (e.g., estrous cyclicity, sperm parameters, and uterine weight)
were not measured in Cox et al. (1975). Confidence in the data base is low, due to a lack of
chronic exposure information from any route of exposure for MEK. Consequently, the RfD is
based on developmental toxicity data for 2-butanol, a compound that is rapidly metabolized to
MEK in rats and shows a time-course profile of metabolites following oral administration that is
similar to the profile for MEK. Although similar developmental effects were reported following
oral and inhalation exposure to 2-butanol and by inhalation exposure to MEK, the lack of oral
data for MEK itself and the absence of data in a second species precludes any higher level of
data base confidence. Reflecting the medium to low confidence in the principal study and low
confidence in the data base, confidence in the RfD is low.
6.2.2. Noncancer/Inhalation
In humans, a number of studies examining the toxicity of MEK following inhalation
exposure exist. The available data include case reports, occupational studies, and controlled
short-term tests with volunteers. Uncertainty in exposure levels and multiple chemical exposure
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precludes dose-response assessment using case reports or occupational studies. The majority of
short-term human studies reported no effects after 4 hours of exposure to 200 ppm (590 mg/m3)
(Dicketal., 1984, 1988, 1989, 1992).
In experimental animals, sufficient evidence is available to conclude that developmental
effects may result from inhalation exposure to MEK. The developmental effects occur at
concentrations between approximately 1,000 and 3,000 ppm (3,000 and 9,000 mg/m3) MEK
7 hours/day on days 6-15 of gestation (Schwetz et al., 1974, 1991; Deacon et al., 1981). By
comparison, this concentration range is not far from the range of exposure levels that have been
reported in human case reports of toxicity: 300-600 ppm (885-1,770 mg/m3) (Smith and Mayers,
1944) and 305-1,695 ppm (900-5,000 mg/m3) (Seaton et al., 1992) (see Section 4.1.2.).
In the previous IRIS assessment from 1992, an RfC of 1 mg/m3 was derived based on a
NOAEL of 2,978 mg/m3 for decreased fetal weight in MEK-exposed mice (Schwetz et al.,
1991). In the current assessment, benchmark dose models were employed to derive the point of
departure for the RfC. From the rat and mouse data on developmental effects produced by
inhalation exposure to MEK (exposed 7 hours/day on days 6-15 of gestation), potential points of
departure were derived from data sets from the developmental toxicity studies of Deacon et al.
(1981) and Schwetz et al. (1991). The lowest of the LECs was selected as the point of departure
(1,764 ppm or 5,202 mg/m3; the lower 95% confidence limit on the concentration associated
with a 10% extra risk of misaligned sternebrae in mice) was adjusted from an intermittent
exposure (7 hours/day) to continuous exposure (24 hours/day) and to a human equivalent
concentration (HEC) by accounting for differences in the blood:air partition coefficients that
have been reported for rats and humans (HEC = 1,517 mg/m3). To this HEC, a combined
uncertainty factor of 300 was applied to account for the pharmacodynamic portion of
interspecies uncertainty, susceptible individuals within the human population, and data base
deficiencies, yielding an RfC of 5 mg/m3.
Confidence in the principal study (Schwetz et al., 1991) is high; it is well-designed and it
tested several exposure concentrations over a reasonable range that included maximum tolerated
doses for dams and fetuses. Also, animal studies in a second species (rats) corroborate the effect
level for developmental toxicity. Confidence in the data base is medium. The data base lacks
chronic exposure toxicity information from any route of exposure, and no multigenerational
reproductive toxicity studies are available for MEK itself. The subchronic inhalation study by
Cavender et al. (1983) satisfies the minimum inhalation data base requirements for derivation of
an RfC. Well-conducted studies in experimental animals provide no convincing evidence that
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repeated inhalation exposure to MEK itself (at much higher exposure levels than those in the
workplace) is capable of producing persistent neurological effects. Portal-of-entry concerns are
addressed by studies in human volunteers showing no net irritation following a 4-hour exposure
to 200 ppm (590 mg/m3). Reflecting high confidence in the principal study and medium
confidence in the data base, confidence in the RfC is medium.
6.2.3. Cancer/Oral and Inhalation
Data in both humans and animals are inadequate to evaluate potential associations
between cancer and MEK exposure by any route. Available studies in humans are insufficient to
evaluate the potential carcinogenicity of MEK. In animals, no chronic study exists for MEK by
any route of exposure; short term tests for genotoxicity have generally been negative. Under the
draft revised cancer guidelines (U.S. EPA, 1999), the data are inadequate for an assessment of
human carcinogenic potential of MEK. Accordingly, the data do not support the derivation of
an oral slope factor or inhalation unit risk for MEK.
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APPENDIX A: SUMMARY OF EXTERNAL PEER REVIEW AND
PUBLIC COMMENTS AND DISPOSITION
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APPENDIX A: SUMMARY OF EXTERNAL PEER REVIEW AND
PUBLIC COMMENTS AND DISPOSITION
The support document and IRIS Summary for MEK have undergone both internal peer
review by scientists within EPA and a more formal external peer review by scientists in
accordance with EPA guidance on peer review (U.S. EPA, 1998b, 2000a). Comments made by
the internal reviewers were addressed prior to submitting the documents for external peer review
and are not part of this appendix. The external peer reviewers were tasked with providing
written answers to general questions on the overall assessment and on chemical-specific
questions in areas of scientific controversy or uncertainty. A summary of significant comments
made by the external reviewers and EPA's response to these comments follows. EPA also
received scientific comments from the public. These comments and EPA's response are
included in a separate section of this appendix.
Comments from External Peer Review
A. General Questions
1. Charge Question: Is the document logical, clear and concise? Are the arguments presented
in an understandable manner?
Comments: The external peer reviewers generally considered the document to be clearly
written, with arguments presented in an understandable manner.
2. Charge Question: Are you aware of any other data/studies that are relevant to the assessment
of adverse effects, both cancer and noncancer, from exposure to MEK?
Comments: One reviewer identified a study that demonstrates the potential for MEK to
potentiate the neurotoxic potential of 2-hexanone in rats (Abdel-Rahman et al., 1976). Another
reviewer suggested that the toxicity literature for methyl ethyl ketoxime (MEKO) might be
relevant to the evaluation of the toxicity of MEK. The reviewer indicated that MEKO is
metabolized to hydroxylamine and MEK, and pointed to a two-generation reproductive toxicity
study by Tyl et al. (1996), in which the hydroxylamine was considered responsible for observed
hematopoietic effects and MEK was considered responsible for observed central nervous system
depression. The reviewer noted that there was no evidence of structural or functional
reproductive toxicity and no developmental/postnatal toxicity associated with MEKO exposure.
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Response: The results from Abdel-Rahman et al. (1976) were incorporated into Section 4.4.4. of
the Toxicological Review.
EPA reviewed the literature for MEKO, and considered the pattern of toxicity associated with
this chemical to be substantially different from that associated with MEK. In the rat, the major
target of MEKO toxicity is the erythrocyte; MEKO induces a methemoglobinemia and a
responsive Heinz body anemia. Liver lesions, secondary to the destruction of erythrocytes, have
also been observed in rats and mice exposed to MEKO. NTP (1999) notes that although there is
strong circumstantial evidence that the erythrotoxicity is due to the hydrolysis product,
hydroxylamine, the evidence is not unequivocal. Other effects associated with MEKO include
degeneration of olfactory epithelium and hyperplasia of the urinary bladder transitional
epithelium, neither of which could be attributed to the ketone metabolite (MEK). In a study of
neurotoxic potential, there was no evidence of cumulative neurotoxicity in a 13-week study at
exposure levels that produced effects on the hematopoietic system. MEKO induced liver tumors
in rats and mice. The mechanism of tumor induction has not been well defined, although one
proposed mechanism includes the oxidation of the oxime to nitronates of secondary-nitroalkanes,
which are mutagenic and tumorigenic in rodents. MEKO is weakly genotoxic (Schulze and
Derelanko, 1993; Tyl et al., 1996; NTP, 1999; Volkel et al., 1999; Newton et al., 2001). Given
the differences in the pattern of toxicity between MEK and MEKO and the difficulties associated
with attributing effects to the parent compound or its metabolites (hydroxylamine and MEK), the
literature for MEKO was not considered useful in the qualitative characterization of MEK
toxicity. Further, the available pharmacokinetic data for MEKO does not establish the percent of
MEKO that is metabolized to MEK, and thus does not support a more quantitative analysis of
the exposure level of MEK present as a metabolite of MEKO associated with reported health
effects. Accordingly, the MEKO toxicity literature was not included in the health assessment for
MEK.
B. Reference Dose (RfD)
1. Charge Question: The RfD is based on data for 2-butanol, a metabolic precursor of MEK,
from Cox et al. (1975). Is the use of 2-butanol as a surrogate for MEK adequately supported?
Comments: All of the reviewers agreed that the use of 2-butanol to establish the RfD for MEK
was appropriate given the available data set. One reviewer recommended a more detailed
examination of pharmacokinetics and peak internal concentrations to determine whether a
dosimetric adjustment should be made to equate MEK generated internally from 2-butanol with
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MEK absorbed directly from the environment. Another reviewer suggested using data from
MEKO, which generates MEK in vivo.
Response: EPA believes that the available data from Dietz et al. (1981) and Traiger and
Bruckner (1976) adequately demonstrate that the elimination profiles of MEK and metabolites in
blood following oral administration of MEK and 2-butanol are comparable, in terms of both area
under the curve (AUC) and peak internal dose. Additional information on peak blood
concentrations from Dietz et al. (1981) was added to Section 3.3. of the Toxicological Review to
better describe these elimination profiles.
For the reasons discussed above, MEKO was not considered by EPA to be a suitable surrogate
for assessing MEK toxicity.
2. Charge Question: Reduced pup weight in the F1A generation, particularly at postnatal day
21, served as the critical effect. Do you consider this effect to be a biologically relevant
response?
Comments: Reviewers had different opinions as to whether reduced F1A pup weight on
postnatal day 21 was the most appropriate, biologically relevant response. One reviewer agreed
that the effect was biologically relevant, while another agreed that reduced pup weight on
postnatal day 21 was a biologically relevant response. The reviewer added that the significance
and magnitude of the response made it of questionable use in establishing an RfD. A third
reviewer commented that body weight data from F2 pups on postnatal day 21 is more objectively
justifiable as the critical effect, and has the advantage of being derived from three dose levels
rather than from the truncated 2-dose data set for Fl A neonates. Another reviewer considered
that pup weight at postnatal day 4 is more closely analogous to fetal growth inhibition and
reduced birth weight in humans, although the postnatal day 21 body weight reduction provided
supportive information.
Response: All of the peer reviewers were of the opinion that reduced pup or fetal weight was a
biologically relevant response, but did not agree on which generation or postnatal day from the
multigeneration reproductive toxicity study was most appropriate for evaluating human health.
The LED05 values (95% lower confidence limit on the effective dose, ED) from the five data sets
were within twofold of each other, suggesting that responses in different generations and at
different times in development were reasonably comparable. In the absence of common
recommendations from the external peer reviewers, EPA retained as the critical data set for the
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RfD the F1A pup body weight data on postnatal day 21, since this data set produced the lowest,
and thus most health protective point of departure.
3. Charge Question: Do you agree with the application of a benchmark dose (BMD) approach
to identify a point of departure using data from the Cox et al. (1975) study? Would use of a
NOAEL/LOAEL approach be preferable?
Comments: Three of the reviewers considered BMD methodology the most appropriate way to
analyze the Cox et al. (1975) data. One of the three reviewers suggested that EPA present the
results of a NOAEL/LOAEL approach in addition to the BMD methodology. The fourth
reviewer strongly disagreed with the use of BMD modeling for this particular data set, but also
considered the NOAEL/LOAEL approach to be inappropriate. The reviewer objected to using
the BMD methodology because the choice of a weight reduction as large as 5% implies a
tolerance for potential adverse effects in humans far larger than should be considered acceptable
as a functional equivalent of a NOAEL or LOAEL. The reviewer further observed that a one
standard deviation change in human birth weight corresponds to a 500 g or so shift, or about
15% of approximate average birth weights of 3,400 g. This reduction in birth weight would be
associated with an excess mortality of about 10/1000 and elevated risks for other conditions and
a 5% shift would be expected to cause one-third this level of impact. Overall, the reviewer
objected to the use of fetal growth inhibition as an effect for deriving an RfD since it has been
this reviewer's experience that dose-response relationships for fetal growth inhibition do not
support nonlinearity.
Response: Consistent with the input of 3 out of the 4 reviewers, EPA agrees that a BMD
analysis is the most appropriate method of analysis for deriving the RfD for MEK, and is
preferable to the application of a NOAEL/LOAEL approach. For this reason, a NOAEL/LOAEL
approach is not presented.
EPA does not agree with the reviewer who stated that the selected benchmark response (BMR)
of 5% is too large. In the absence of some idea of a specific level of response to consider
adverse, EPA's "Benchmark Dose Technical Guidance" (U.S. EPA, 2000c) recommends a
change in the mean equal to one standard deviation from the control mean. This level of
response "gives an excess risk of approximately 10% for the proportion of individuals below the
2nd percentile or above the 98th percentile of controls for normally distributed effects." EPA is
not aware of precedents for using BMR values as small as 0.1 or 1% as suggested by the
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reviewer. Furthermore, a BMR of 5%, in this case, is within the range of experimental dose
levels used in the Cox et al. (1975) study, and for that reason is supportable.
4. Charge Question: Are the appropriate uncertainty factors applied? Is the explanation for
each transparent?
Comments: Two reviewers considered that the uncertainty factors have been appropriately
applied. One reviewer indicated that a data base uncertainty factor of 10 is overly cautious, and
suggested that a 3-fold uncertainty factor might be more appropriate. The fourth reviewer
offered an alternative approach to the BMC (or NOAEL) divided by uncertainty factors. The
reviewer also suggested a target RfD be derived that would produce no more than an expected
one-in-one million extra burden of infant mortality, and that this target be tied to day 4 body
weight reduction (in the F1A and F2 litters). This approach would reduce the RfD by about
ninefold from the value of 0.7 mg/kg-day presented in the external review draft.
Response: EPA considers a data base uncertainty factor of 10 to be appropriately cautious for
the MEK RfD, particularly given the lack of oral toxicity data specific to MEK.
At this time, it is EPA's position that the extrapolation from an exposure level associated with a
one-in-one million extra burden of infant mortality is not supportable, since the uncertainty
associated with projecting a dose corresponding to a one-in-one million extra burden level from
the available animal data, i.e., Cox et al. (1975), would be substantial. Thus, EPA is retaining
the current BMD approach and uncertainty factors presented in the external review draft.
C. Reference Concentration (RfC)
1. Charge Question: The RfC for MEK derived in 1993 is based on reduced fetal weight as
reported in the mouse developmental toxicity study of Schwetz et al. (1991). The RfC derived in
the reassessment is based on a different developmental endpoint - increased incidence of
misaligned sternebrae - from the same Schwetz et al. (1991) study. Do you consider an
increased incidence of this skeletal variant to be a biologically relevant endpoint? Would an
alternative endpoint (e.g., reduced fetal weight in the mouse) be more appropriate?
Comments: None of the reviewers offered an opinion that misaligned sternebrae was not a
biologically significant endpoint of toxicity. One reviewer who considered misaligned
sternebrae to be biologically relevant and preferable to the use of reduced fetal weight noted the
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considerable difference of opinion among experts in evaluating increased incidence of
misaligned sternebrae in mouse developmental toxicology. A second reviewer considered
reduced fetal weight in the mouse to be the most appropriate endpoint in this data set because
body weight is a continuous rather than dichotomous variable, and because misaligned
sternebrae were elevated on a per fetus basis rather than a per litter basis (the latter considered
the more appropriate statistical unit). A third reviewer stated that, based on the reviewer's 40
years of experience performing developmental toxicity studies in rodents, fetal weight per litter
is usually the most sensitive indicator of fetal toxicity. The reviewer also stated that the
presentation of skeletal effects data as the number of affected fetuses/group was not appropriate,
since the unit for statistical analysis is the dam or the litter, rather than the fetus. The reviewer
recommended the use of more sophisticated statistical software programs (e.g., SUDAAN) that
are designed to analyze correlated data. The fourth reviewer noted the highly nonlinear dose-
response relationship for misaligned sternebrae, and suggested that fetal weight reduction might
be preferable for addressing human health risk at low doses since it has a more linear dose-
response relationship.
Response: Based on external peer reviewers' comments and further analysis of the data from
Schwetz et al. (1991), EPA gave further consideration to the use of fetal weight data as the basis
for the point of departure for the RfD rather than incidence of misaligned sternebrae. Points that
support the use of fetal weight rather than misaligned sternebrae included the fact that the
incidence of misaligned sternebrae was highly nonlinear with dose, and the incidence showed a
high degree of variability. Mast et al. (1989) also reported a positive trend (not statistically
significant) for reduced ossification of the sternebrae, raising the possibility that misaligned
sternebrae may reflect a more general growth delay. Selection of the critical effect (i.e., reduced
fetal weight or incidence of misaligned sternebrae) was further deliberated during consensus
review. In general, consensus reviewers did not find the arguments for using fetal weight as the
more supportable critical effect to be compelling. Thus, the decision was made to use the more
health protective endpoint, misaligned sternebrae, as the critical effect.
2. Charge Question: Do you agree with the application of a benchmark dose (BMD) approach
to identify a point of departure using data from Schwetz et al. (1991)? Would use of a
NOAEL/LOAEL approach be preferable?
Comments: All four reviewers considered BMD methodology the most appropriate way to
analyze the data for deriving the point of departure for the RfC. One reviewer added that BMD
methodology was appropriate as long as it was applied to the data for the quantal endpoint of
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misaligned sternebrae. Another reviewer suggested presenting the results of the
NOAEL/LOAEL approach in addition to the BMD methodology.
Response: A BMD methodology was retained for the analysis of the RfC, consistent with the
external peer reviewer feedback.
3. Charge Question: The State of California has developed a draft Reference Exposure Level
(REL), which is comparable to EPA's RfC, of 3 mg/m3 based on the Mitran et al. (1997)
occupational study that reported various neurological effects in MEK-exposed workers. Because
of certain critical limitations in this study, it was not selected as the basis for the RfC. Is this
decision adequately supported? More generally, is the weight of evidence for the neurotoxic
potential of MEK adequately described?
Comments: Two of the reviewers agreed that Mitran et al. (1997) suffered from too many
deficiencies to serve as a reliable basis for the RfC. A third reviewer commented that if the
worker study used by the State of California was adequately and appropriately performed and
reported, then it is a valid basis for the RfC, especially since it provided data in humans.
The fourth reviewer appears to have misinterpreted the question as asking if the REL of 3 mg/m3
was comparable to the external review draft RfC of 15 mg/m3. The reviewer also stated that
consistency between the proposed RfC and proposed RfD could be accomplished by developing
and applying PBPK models in rats, mice and humans, such that lexicologically equivalent doses
(based on either Cmax in the plasma or AUC during key periods in development) could be
projected. The reviewer suggested using another calculation based on a test of acute behavioral
toxicity. Specifically, the reviewer recommended using Glowa and Dews (1987) in mice to
define a point of departure based on sophisticated neurobehavioral modeling of interindividual
variability for a sensitive neurobehavioral change.
Two of the four reviewers commented on the adequacy of the description of the weight-of-
evidence for the neurotoxic potential of MEK in humans. One of the two considered the
evidence to be adequately described, while the second did not think the IRIS assessment
adequately addressed or discussed the neurotoxic potential of MEK, but did not further identify
areas where discussion could be improved.
Response: As supported by two of the external peer reviewers, EPA continues to consider the
Mitran et al. (1997) study to be too limited to use for RfC determination. In the absence of
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specific feedback on how to improve the discussion of the neurotoxic potential of MEK, the
relevant sections of the Toxicological Review were not substantially revised.
EPA agrees that the development of a PBPK model for MEK that would support extrapolation
between species and between routes of exposure would inform this assessment. Given the need
for timely completion of the MEK IRIS assessment, it was decided to move forward with
information contained in the published peer reviewed literature available at the time of the
reassessment. A PBPK model that supports interspecies and interroute extrapolation was not
available at the time of the assessment.
An analysis of interindividual variability among mice, as proposed by one of the reviewers, was
not undertaken. The Glowa and Dews (1987) study examined schedule-controlled behavior in
mice exposed to high concentrations of MEK for up to 2 hours. Because developmental toxicity
served as the critical effect for the RfC, and because repeat-dose toxicity studies with MEK do
not provide evidence of neurotoxicity, the proposed analysis of interindividual variability was
not considered appropriate.
4. Charge Question: Has the matter of MEK's capacity to produce interactions with other
toxicants (e.g., n-hexane) been sufficiently acknowledged and accommodated in the assessment?
Comments: In general, the external peer reviewers considered the discussion of MEK's capacity
to interact with other toxicants to be sufficiently addressed. One reviewer directed EPA to a
study by Couri et al. (1977) that reported the potential for MEK to increase potentiation of the
end toxicant, 2,5-hexanedione (2,5-HD) from 2-hexanone. Another reviewer noted that the
mechanism of potentiation as described in the assessment was not completely correct, and noted
that the positive interaction between MEK and 2,5-HD in producing neurotoxicity was best
understood by competitive inhibition by MEK of the enzymes responsible for metabolic
detoxification of 2,5-HD.
Response: EPA reviewed the Couri et al. (1977) study, in which the exposure of rats to MEK in
combination with methyl n-butyl ketone reduced hexobarbital sleep time. Blood levels of 2,5-
HD and 2-hexanone were not measured, nor were other measures of neurotoxicity evaluated. It
was determined that the study did not add appreciably to the current discussion of evidence for
MEK's potentiation of neurotoxicants. The discussion of the mechanism of potentiation was
revised, however, to state that potentiation of n-hexane neurotoxicity appears to be due to the
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increased persistence of 2,5-HD in blood, probably due to the inhibition of 2,5-HD phase II
biotransformation by MEK.
5. Charge Question: Are the appropriate uncertainty factors applied? Is the explanation for
each transparent? Considering the nature of the critical effect, is an additional factor needed to
reduce the point of departure to one that poses minimal health risk.
Comments: Three of the four reviewers considered the assigned uncertainty factors (UFs) to be
appropriate and sufficiently justified. The three reviewers stated that no additional UF was
needed. One of the three reviewers observed that mice typically exhibit spontaneous incidences
of many skeletal effects, such as misaligned sternebrae, so the specific findings may be mouse
specific. One reviewer commented that an additional factor was needed to reduce the point of
departure to one that poses minimal health risk.
Response: Consistent with the comments of the majority of external peer reviewers, the UFs
applied in deriving the RfC were retained.
D. Cancer Weight-of-Evidence Evaluation
1. Charge Question: The weight-of-evidence characterization is discussed in Section 4.6. Have
appropriate criteria been applied from EPA's draft revised Guidelines for Carcinogen Risk
Assessment (U.S. EPA, 1999)?
Comments: All four reviewers were of the same opinion that the 1999 draft revised cancer
guidelines have been appropriately applied in characterizing the cancer weight-of-evidence. One
reviewer noted that the Cox et al. (1975) multigeneration study in rats indicates that the F2
offspring were transferred to a chronic drinking water study of MEK, and questioned whether
the report from that chronic study might be available. A second reviewer noted that the
summary of carcinogenicity incorrectly suggested that MEK can produce bone and prostate
cancer.
Response: The laboratory that performed the multigeneration study (Food and Drug Research
Laboratories, Inc.) is no longer in operation. EPA's literature search did not identify a chronic
drinking water study of MEK in the published literature, and does not have knowledge of an
unpublished version of a chronic study. The summary statement that mentioned bone and
prostate cancers was revised to state: "Although there is some suggestion of increased risk for
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some cancers (including bone and prostate) and multiple solvent exposure that includes MEK,
there is no clear evidence for a relationship between these cancers and MEK exposure alone."
E. Questions Regarding the Study by Cox et al. (1975)
The multigeneration reproductive and developmental toxicity study by Food and Drug Research
Laboratories, Inc. (Cox et al., 1975) serves as the principal study for the RfD. EPA sought
independent peer review of this laboratory report from two external peer reviewers because the
study's findings were not published in the peer-reviewed literature and because the study was
conducted prior to the introduction of Good Laboratory Practices.
1. Charge Question: Was the study design adequate?
Comments: One of the two reviewers considered the study to be adequately designed. A second
reviewer commented that the study was based on early FDA testing guidelines, but did not
consider the design adequate.
Response: EPA recognizes that the Cox et al. (1975) study design, which was based on early
FDA testing guidelines, does not evaluate all endpoints that are covered under current testing
protocols. Major deficiencies are discussed in Section 4.3.2.1. Discussion of other study design
limitations have been added to the Toxicological Review.
2. Charge Question: Were the study findings adequately reported?
Comments: One of the two reviewers indicated that findings were, in general, adequately
reported. The reviewer noted the absence of discussion of male reproductive effects in terms of
mating success and failure, and in particular pointed to data in Appendix II of Cox et al. (1975)
that suggested a potential adverse impact on reproductive performance, but not on the fertility of
high-dose males. The reviewer summarized the high dose male data as follows:
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Male Rat Copulatory Success When Breeding F1A Litters*
Copulation
Failed
Succeeded
Total
Control
1
29
30
0.3%
2
28
30
1%
0
30
30
3%
6
24
30
* Successful mating was judged by the presence of a vaginal plug or sperm in a
vaginal smear. Males that failed, by the end of the second cycle, to mate successfully
according to that measure were replaced by a previously successful male. In Appendix
II of Cox et al. (1975), the second entry of a male rat number was interpreted as
indicative of a mating failure.
Source: Adapted from Cox et al., 1975.
In light of the data on FO male copulatory success, the reviewer also questioned whether the high
dose is a NOAEL for reproductive toxicity for both males and females. The reviewer concluded
with the observation that even if an effect were determined to be statistically or lexicologically
significant, it is unlikely to have altered selection of the critical effect. The second reviewer
noted the following reporting inadequacies: presentation of means/group, with no variance term
(e.g., standard deviation or standard error) and no statistical analyses of any parameter.
Response: A discussion of the incidence of FO male rats that did not successfully copulate with
FO females was added to the Toxicological Review. The identification of the high-dose group as
a NOAEL for reproductive toxicity was qualified with the observation of a possible increase in
male copulatory failure at that dose level.
EPA recognizes that the study presents means/group without a variance term. Because
appendices to the report provided individual litter data, statistical analyses were conducted for
certain data sets and standard deviations were calculated. Standard deviations were included in
the Toxicological Review.
3. Charge Question: Were the authors' conclusions supported by the results?
Comments: One reviewer maintained that the conclusions are supported by the study results. A
second reviewer did not, in all instances, find the conclusions adequately supported. In the
absence of statistical analyses, the reviewer considered the authors' conclusions to be "'cautious'
and very conservative (e.g., effects may have been observed by the authors did not designate
them as 'significant' or necessarily treatment- or concentration-related)."
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Response: EPA conducted independent statistical analysis of the data from Cox et al. (1975) and
based its conclusions on the findings from those analyses. Therefore, the assessment for MEK
was not necessarily limited by instances where the study authors' conclusions were necessarily
"cautious."
4. Charge Question: Are there any notable limitations or deficiencies in this study?
Comments: One reviewer did not identify any notable limitations or deficiencies in the study.
The second reviewer, however, identified the following deficiencies: (1) no water consumption
during gestation and lactation, so maternal intake cannot be correlated with maternal and
offspring effects; (2) no dosed water during the last two weeks of FO lactation of Fl A litters; (3)
the high concentration was changed from 3.0 to 0% during the last two weeks of lactation, to 2%
for the postwean Fl A offspring; (4) no information was presented on the concentration provided
to the FO parents after the weaning of F1A litters; (5) only minimal types of assessment at
minimal times were assessed for critical parameters (e.g., body weights, feed and water
consumption, clinical signs, etc.); (6) limited numbers of F1A adult animals/group (and none of
the FO and F2) were assessed for clinical chemistry/hematology and histopathology; (7) data
were presented as means/group with no variance terms and no statistical analyses; and (8) the
conclusions were "tepid," probably due to the lack of statistical analyses and the very limited
assessments performed, with no summarized information or conclusions for effects at 3%.
Response: Additional discussion of the specific limitations identified by one reviewer was
included in Section 4.3.2.1. of the Toxicological Review.
5. Charge Question: Is EPA's summary of the study in the Toxicological Review and analysis
of the study findings appropriate?
Comments: One reviewer concurred with the use of the Cox et al. (1975) study and the selection
of reduced neonatal body weight as the critical effect, but questioned the designation of 1% 2-
butanol as the LOAEL for neonatal body weight, which the study authors regarded as a no-effect
level. The reviewer stated that day 4 and day 21 body weight reductions in the F1A generation
did not appear to be statistically significant in the mid-dose (1%) group, and the designation of
the 1% solution as the LOAEL was not adequately justified. A second reviewer noted that EPA
focused on the few strengths of the study and ignored its weaknesses. In particular, the reviewer
noted that effects on body weight in male F1A pups on postnatal day 21 occurred in the absence
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of dosing (since treatment was discontinued for the last two weeks of F1A lactation), and
questioned how EPA could assign a dose to this effect.
Response: EPA reconsidered the designation of NOAELs and LOAELs from the Cox et al.
(1975) study. Rather than designate these values for each generation, a NOAEL and LOAEL for
the study as a whole were established. Although the Fl A pup body weights for the low- and
mid-dose groups were 4 to 10% lower than the control, a similar reduction was not observed in
subsequent generations (i.e., FIB and F2). Therefore, the mid-dose group (1%) was considered
to be the NOAEL and the 2% group the LOAEL, consistent with the interpretation of the study
investigators.
EPA was similarly concerned with the interpretation of the Fl A data set at the high dose
resulting from changes in the 2-butanol drinking water concentration during lactation.
Accordingly, EPA did not use pup body weight data from the high-dose group in the dose-
response analysis of F1A pup body weight. Body weight data from the control, low- and mid-
dose groups only were used.
6. Charge Question: Overall, was the study as designed, performed, and reported of sufficient
quality to use as the basis for the RfD?
Comments: One reviewer considered the study to be suitable as the basis for the RfD, while the
second did not. The second reviewer noted that for its time (1973 to 1975) the study was
conducted as well or better than many, and more or less followed FDA multigeneration study
design. In the opinion of this reviewer, however, the study is not supportable when compared
against current guidelines as the basis for the RfD.
Response: EPA acknowledges the limitations of the Cox et al. (1975) study design when
compared against current protocols for multigeneration studies. The study shows dose-related
effects, however, that are consistent across generations and appear to be attributable to 2-butanol
exposure. If this study was deemed to be inadequate, there would be no alternative basis for
derivation of an RfD for MEK. Because the study deficiencies identified by an external peer
reviewer were a function of the age of the study and not inappropriate study conduct, and
because compound-related effects were reported, EPA decided to retain the Cox et al. (1975)
study as the principal study for the RfD.
F. Additional Comments from External Peer Reviewers
A-13
-------
Comment: Several of the reviewers offered editorial comments and suggestions to improve the
clarity of the text.
Response: In most instances, the suggested editorial changes and revisions to the text were
incorporated.
Comment: One reviewer suggested that the EPA create a PBPK model for dosimetic
comparisons across species. The reviewer considered the data on percentage retention as a
function of the duration of exposure to be an excellent basis for calibrating human PBPK
models. Because absorption and therefore metabolism appear to be dose dependent, the reviewer
suggested that PBPK modeling was needed to sort out the likely changes in the relationship
between external and internal exposure levels between the high doses where testing was done
and the lower environmental doses to which humans might be exposed.
Response: PBPK models have been developed for MEK in humans and in the rat, but a mouse
model has not been developed. Development of a mouse model would require parameterization
for the mouse, validation, and calibration against an additional data set. In addition, such a
model would need to be subject to peer review. Given the need for timely reassessment of the
health effects of MEK and associated reference values, EPA decided to move forward with the
published peer reviewed literature at hand at the time of the reassessment. EPA is open to
considering other PBPK models that are developed in the future.
Comment: One reviewer pointed out a conceptual or typographic error in Table 1 (Kinetic
parameters used for PBPK models for MEK kinetics in humans and rats). This reviewer noted
that because the entire cardiac output must pass through the lungs before returning to the heart,
the blood flow to the lungs must be listed as 100%, and the flows to the remaining tissues should
also total 100%.
Response: The table was corrected as suggested by the reviewer.
Comment: Two reviewers offered additional comments on the summary of the Cox et al. (1975)
study. One reviewer recommended that the observations in the Fl A and F2 litters be discussed
in terms of statistical significance (or lack thereof), and that the results of the FO male rat's
copulatory success be noted. The second reviewer disagreed with the LOAEL and NOAEL
designations for fetal growth inhibition, and recommended that the decrease in pup weight would
A-14
-------
better be described as a continuous function of dose with a defined slope than as something that
becomes statistically significant only at defined high doses.
Response: EPA believes that an analysis of the biological significance of the findings of the Cox
et al. (1975) study is more relevant than a statistical analysis of the results. To that end, the
discussion of the biological significance of the study findings was expanded. Further, statistical
analysis was conducted as part of the BMD analysis of the data; BMD software outputs are
provided in Appendix B of the Toxicological Review. While the analysis does not include pair
wise statistical tests, it does include a log-likelihood ratio test that provides an indication of
whether or not pup/fetal body weights changed significantly with increasing dose levels. This
type of analysis appears to be consistent with the recommendations of the second reviewer.
Comments from the Public
Comment: One public commenter concurred with the use of the Cox et al. (1975) study as an
appropriate basis for the RfD, but found inadequate justification for modifying the NOAEL and
LOAEL from that presented previously on IRIS. The commenter noted that the IRIS Summary
(posted in 1993) for MEK considered 1,771 mg/kg-day to be a NOAEL, whereas this dose level
was considered to be a LOAEL in the external review draft based on reduced pup weight
observed in Fl A litters, but not FIB or F2 litters. The commenter also noted that the study was
recently reviewed by EPA scientists as part of an OECD SIDS Dossier and SIAR for 2-butanol,
and in that context 1,771 mg/kg-day was regarded as a NOAEL.
Response: A similar comment was received from one of the external peer reviewers. As
discussed above, the summary of the Cox et al. (1975) study was revised such that a NOAEL and
LOAEL for the study as a whole were identified. Because the body weight reductions compared
to control in the low- and mid-dose groups in Fl A pups were not observed in subsequent
generations (i.e., FIB and F2), a NOAEL and LOAEL for the study as a whole of 1% (1,771
mg/kg-day) and 2% (3,122 mg/kg-day), respectively, were presented.
Comment: A commenter did not consider the 5% decrease in mean pup or fetus body weight to
be sufficiently justified, particularly when 5% is less than one standard deviation and the
purported effect was not observed at 1,771 mg/kg-day in the FIB and F2 litters. The commenter
indicated that it is more scientifically reasonable to use a 10% reduction in pup weight as the
benchmark response. The commenter also considered the inconsistent selection of a benchmark
response rate (either 5% or 10%) to be arbitrary.
A-15
-------
Response: One standard deviation as a benchmark response (BMR) serves as a recommended
point of comparison across assessments, but is generally considered the last choice as the basis
for the BMR if there is no other basis for a biologically relevant degree of change in response.
In the case of pup body weight data from the Cox et al. (1975) study upon which the RfD was
based, there is no specific decrement that is generally regarded as indicative of a biologically
relevant response. EPA considered the use of one standard deviation as the BMR. Using data
from Cox et al. (1975), one standard deviation from the control mean resulted in BMDs that
corresponded to body weights 9 to 26% below the control mean (see Tables 6, 7 and 9) - values
generally above the range of experimental data. Because an aim in BMD modeling is to select a
BMD within the range of observation, other measures of the BMR were examined. A 5%
reduction in fetal/pup body weight relative to the control was a response rate that fell within the
range of experimental dose levels in the Cox et al. (1975) study, and consequently was selected
as the benchmark response (BMR). In addition, an ED10 and LED10 for each endpoint were
estimated as a consistent point of comparison across chemicals, as recommended in the
Benchmark Dose Technical Guidance Document (U.S. EPA, 2000c). These additional measures
are provided in Appendix B. The basis for the selection of a BMR of 5% is discussed in Section
5.1.2.1. of the Toxicological Review.
In the case of the RfC, two data sets from Schwetz et al. (1991) were analyzed by BMD
methods. A decrease in mean fetal weight of one standard deviation of the control mean was
selected as the BMR for this endpoint. The BMR corresponds to an approximately 5% decrease
in mean body weight for the data set, and a BMD generally within the range of experimental
dose levels used in the Schwetz et al. (1991) study. Ten percent extra risk was used as a BMR
for the dichotomous response, misaligned sternebrae, also from Schwetz et al. (1991) because
the BMR corresponds to a BMD within the range of the experimental dose levels. EPA also
tried using 5% extra risk as a BMR in the analysis of misaligned sternebrae, but the nested model
could not provide a useful estimate of the lower bound on the BMD for this BMR (i.e., the lower
bound on the BMD was estimated as essentially zero). The basis for selecting these BMRs is
discussed in Sections 5.2.2.1.2. and 5.2.2.1.3. of the Toxicological Review.
Comment: A commenter stated that decreased fetal birth weight, rather than an increased
incidence of misaligned sternebrae, is the more scientifically sound basis for the RfC. The
commenter noted that misaligned sternebrae represents an anomalous skeletal variation that was
seen in only one of four developmental toxicity studies of MEK. The commenter also observed
that since the control standard deviation is 10%, approximately 15% of the control litters have an
"adverse finding," suggesting that the benchmark response is too restrictive.
A-16
-------
Response: As discussed in the response to external peer reviewer comments, EPA reevaluated
study findings for misaligned sternebrae as reported by Schwetz et al. (1991) and Mast et al.
(1989) and the dose-response analyses for both misaligned sternebrae and fetal weight from this
study. The endpoints were given further consideration as potential critical effects. Based on
deliberations during consensus review, it was decided to use misaligned sternebrae as the
endpoint yielding the more health protective point of departure.
Comment: One commenter stated that EPA did not present an adequate scientific justification
for applying a duration adjustment to the inhalation developmental toxicity study and, at the very
least, the additional conservatism added by the application of this factor should be explicitly
recognized. The commenter pointed to the statement in the Toxicological Review that "MEK is
rapidly absorbed, distributed, and metabolized, suggesting that duration adjustment may be
inappropriate," and to EPA's conclusions that ultimately there was not "sufficient evidence to
argue convincingly for either peak exposure level or area under the curve."
Response: Duration adjustment of the exposure concentrations in the developmental study of
MEK (Schwetz et al., 1991) was performed consistent with recent EPA guidance, A Review of
the Reference Dose and Reference Concentration Processes (U.S. EPA, 2002). The document
recommends that duration adjustment procedures to continuous exposure based on C x t be used
as a default procedure for inhalation developmental toxicity studies as it is for other health
effects from inhalation exposure. The recommendation is based on evidence that shows that
some agents cause developmental toxicity more as a function of peak concentration, whereas the
effects of other agents are related to area-under-the-curve (AUC). The latter is true even of some
developmental toxicants with a short half-life. In the absence of data that supports peak
concentration or AUC as more closely correlated with developmental toxicity, EPA's 2002
review document recommends duration adjustment as the more health protective default
procedure. As noted in the Toxicological Review of MEK, because the data are insufficient to
argue convincingly for either peak exposure level or AUC as the most appropriate metric, the
more health protective procedure (duration adjustment) was applied as a policy matter. The text
of the Toxicological Review was revised to better support this decision.
Comment: One commenter disagreed with EPA's interpretation of the high-exposure level
(5,000 ppm) in the Cavender et al. (1983) study as a LOAEL based on reduced body weight,
increased liver weight, and decreased brain weight. The commenter believed that these organ
weight changes, in the absence of histopathology, should not be considered adverse effects and
that the 5,000 ppm exposure level should be considered a NOAEL. The commenter also noted
A-17
-------
that an OECD SIDS Dossier and STAR for 2-butanol identified a NOAEL of 5,000 ppm on the
basis that the changes in organ weight and clinical pathology parameters were not supported by
histological changes.
Response: EPA gave further consideration to the biological significance of the findings in the
5,000 ppm animals in the Cavender et al. (1983) study, and in particular the organ weight
findings. Although the decrease in brain weight in female high-dose animals is of some concern,
EPA agrees that this effect, in the absence of corresponding histopathology and functional
abnormalities, cannot be clearly characterized as being of toxicological relevance. The text was
revised to further emphasize the difficulty in interpreting the significance of these findings. In
light of these uncertainties, characterization of the 5,000 ppm exposure level as a LOAEL and
the mid-dose group (2,518 ppm) as a NOAEL were dropped.
Comment: A commenter agreed with EPA's characterization of the Mitran et al. (1997) study,
but noted that the California Office of Environmental Health Hazard Assessment (OEHHA) has
withdrawn its draft Reference Exposure Level (REL) based on the study.
Response: According to Dr. Andrew Salmon of Cal EPA's OEHHA (May 27, 2003), the REL
for MEK has not been withdrawn. A draft of the REL is still available; however, OEHHA is not
working to finalize the MEK assessment at this time. One of the primary issues encountered in
reviewing the REL concerned the Mitran et al. (1997) study that serves as the basis for the draft
REL. Cal EPA's science review panel expressed concerns with the findings of the study. Their
concerns are consistent with those raised by EPA in review of the MEK literature.
A-18
-------
APPENDIX B: BENCHMARK DOSE MODELING RESULTS AND OUTPUT
-------
Output B-l: Reduced Pup Body Weight in Wistar Rats, F1A Generation at Postnatal
Day 4 (Cox et al., 1975)
17:0303/032003
Linear Model with 0.95 Confidence Level
500
1000
1500
dose
The form of the response function is:
Y[dose] = beta_0 + beta_l*dose + beta_2*dose^2 + ...
Dependent variable = MEAN
Independent variable = Dose
rho is set to 0
Signs of the polynomial coefficients are not restricted
A constant variance model is fit
Total number of dose groups = 3
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 = 1.52807
rho = 0 Specified
beta_0 = 10.5909
beta 1 = -0.00038168
Variable
alpha
beta_0
beta 1
Parameter Estimates
Estimate Std. Err.
1.48828 0.226961
10.5956 0.193304
-0.00038201 0.000176104
Asymptotic Correlation Matrix of Parameter Estimates
B-l
-------
1.1
1.3
1.3
Likelihoods of Interest
Model Log(likelihood) DF
Al -59.705513 4
A2 -59.191916 6
fitted -60.097751 2
R -62.891337 2
Tests of Interest
Test -2*log(Likelihood Ratio) Test df
Test 1 7.39884 4
Test 2 1.02719 2
Test 3 0.784476 1
The p-value for Test 3 is greater than .05. The model chosen appears to adeguately describe the
data
Benchmark Dose Computation
Specified effect = 0.05
Risk Type = Relative risk
Confidence level = 0.95
BMD =
BMDL =
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Relative risk
Confidence level = 0.95
BMD = 2773.65
BMDL = 1606.17
B-2
-------
Output B-2: Reduced Pup Body Weight in Wistar Rats, F1A Generation at Postnatal
Day 21 (Cox et al., 1975)
Linear Model with 0.95 Confidence Level
500
1000
1500
dose
17:1QQ3,TO2003
Dependent variable = MEAN
Independent variable = dose
rho is set to 0
Signs of the polynomial coefficients are not restricted
A constant variance model is fit
Total number of dose groups = 3
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 = 17.7256
rho = 0 Specified
beta_0 = 48.8619
beta 1 = -0.00278464
Parameter Estimates
Estimate Std. Err.
17.1217 2.62635
48.8645 0.66404
-0.00278278 0.000601426
Asymptotic Correlation Matrix of Parameter Estimates
alpha beta_0 beta_l
alpha 1 -2e-007 2.5e-007
beta_0 -2e-007 1 -0.74
beta 1 2.5e-007 -0.74 1
-------
Likelihoods of Interest
Model Log(likelihood) DF
Al -163.160835 4
A2 -162.157710 6
fitted -163.214728 2
R -172.764782 2
R)
Tests of Interest
Test -2*log(Likelihood Ratio) Test df
Test 1 21.2141 4
Test 2 2.00625 2
Test 3 0.107786 1
A homogeneous variance model appears to be
The model chosen appears to adeguately describe the
Benchmark Dose Computation
Specified effect = 0.05
Risk Type = Relative
Confidence level = 0.95
BMD = 877.979
BMDL = 656.797
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Relative
Confidence level = 0.95
BMD = 1755.96
BMDL = 1313.59
B-4
-------
Output B-3: Reduced Fetal Weight in Wistar Rats, FIB Generation (Cox et al., 1975)
Polynomial Model with 0.95 Confidence Level
3000
17:1603/032003
The form of the response function is:
Y[dose] = beta_0 + beta_l*dose + beta_2*dose^2 + ...
Dependent variable = MEAN
Independent variable = dose
rho is set to 0
The polynomial coefficients are restricted to be negative
A constant variance model is fit
Total number of dose groups = 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
Default Initial Parameter Values
alpha = 1.18178
rho = 0 Specified
beta 0 = 4.07806
beta~l = 0
beta 2 = -1.71632e-007
Variable
alpha
beta_0
beta_l
beta 2
Parameter Estimates
Estimate Std. Err.
1.16541 0.15369
4.25434 0.132747
0 NA
-4.4034e-008 2.57616e-008
Asymptotic Correlation Matrix of Parameter Estimates
alpha beta_0 beta_2
alpha 1 -9.le-008 1.2e-007
beta_0 -9.le-008 1 -0.65
beta~2 1.2e-007 -0.65 1
B-5
-------
Table of Data and Estimated Values of Interest
Dose N Obs Mean Obs Std Dev Est Mean Est Std Dev Chi^2 Res.
29 4.14
27 4.16
30 4.38
29 3.74
Model Descriptions for likelihoods calculated
Model A2 : Yij = Mu ( i ) + e(ij)
Var{e(ij)} = Sigma (1)^2
Likelihoods of Interest
Model Log(likelihood) DF
Al -65.068238 5
A2 -57.686139 8
fitted -66.301668 2
R -67 .746445 2
Tests of Interest
Test -2*log(Likelihood Ratio) Test df
Test 1 20.1206 6
Test 2 14.7642 3
Test 3 2.46686 1
The p-value for Test 2 is less than .05. Consider running a non-homogeneous variance model
The p-value for Test 3 is greater than .05. The model chosen appears to adeguately describe the
data
Benchmark Dose Computation
Specified effect = 0.05
Risk Type = Relative risk
Confidence level = 0.95
BMD = 2197.9
BMDL = 1046.23
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Relative risk
Confidence level = 0.95
BMD = 3108.29
BMDL = 2085.07
B-6
-------
Output B-4: Reduced Pup Body Weight in Wistar Rats, F2 Generation at Postnatal
Day 4 (Cox et al., 1975)
Linear Model with 0.95 Confidence Level
10.5
(.0
c
o
D-
cc
in
500
1000
1500 2000
dose
2500
3000
3500
08:22 03/04 2003
Dependent variable = MEAN
Independent variable = dose
rho is set to 0
Signs of the polynomial coefficients are not restricted
A constant variance model is fit
Total number of dose groups = 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
Default Initial Parameter Values
alpha = 3.09184
rho = 0 Specified
beta_0 = 9.89257
beta 1 = -0.000140383
Variable
alpha
beta_0
beta 1
Parameter Estimates
Estimate Std. Err.
2.98335 0.407875
9.89404 0.249663
-0.000142507 0.00014249
Asymptotic Correlation Matrix of Parameter Estimate
alpha beta_0 beta_l
alpha 1 -4.5e-008 5.4e-008
beta_0 -4.5e-008 1 -0.74
beta 1 5.4e-008 -0.74 1
B-7
-------
Model
Likelihoods of Interest
Model Log(likelihood) DF
Al -111.850740 5
A2 -107.811454 8
fitted -111.977995 2
R -112.478139 2
Tests of Interest
-2*log(Likelihood Ratio) Test df
9.33337 6
8.07857 3
0.254509 2
Benchmark Dose Computation
Specified effect = 0.05
Risk Type = Relative risk
Confidence level = 0.95
BMD = 3471.42
BMDL = 1347.21
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Relative risk
Confidence level = 0.95
BMD = 6942.85
BMDL = 2694.43
-------
Output B-5: Reduced Pup Body Weight in Wistar Rats, F2 Generation at Postnatal
Day 21 (Cox et al., 1975)
Polynomial Model with 0.95 Confidence Level
42
40
Ill
c
o
38
36
34
32
Polynomial
BMDL
BMD
500
1000
1500
dose
2000
2500
3000
17:2603/032003
The form of the response function is:
Y[dose] = beta_0 + beta_l*dose + beta_2*dose^2 + ...
Dependent variable = MEAN
Independent variable = dose
rho is set to 0
The polynomial coefficients are restricted to be negative
A constant variance model is fit
Total number of dose groups = 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
Default Initial Parameter Values
alpha = 52.6945
rho = 0 Specified
beta_0 = 39.6222
beta_l = 0
beta 2 = -6.24081e-007
Variable
alpha
beta_0
beta_l
beta 2
Parameter Estimates
Estimate Std. Err.
50.8889 7.0912
39.7965 0.899513
0 NA
-4.70576e-007 1.85023e-007
Asymptotic Correlation Matrix of Parameter Estimates
alpha beta_0 beta_2
alpha 1 2.4e-008 4.8e-008
beta_0 2.4e-008 1 -0.62
beta 2 4.8e-008 -0.62 1
B-9
-------
Table of Data and Estimated Values of Interest
Dose N Obs Mean Obs Std Dev Est Mean Est Std Dev Chi^2 Res
Model Descriptions for likelihoods calculated
Model A2 : Yij = Mu ( i ) + e(ij)
Var{e(ij)} = Sigma (1)^2
Likelihoods of Interest
Model Log(likelihood) DF
Al -253.632495 5
A2 -247.410830 8
fitted -253.876754 2
R -257.015989 2
Tests of Interest
Test -2*log(Likelihood Ratio) Test df
Test 1 19.2103 6
Test 2 12.4433 3
Test 3 0.488518 1
The p-value for Test 2 is less than .05. Consider running a non-homogeneous variance model
The p-value for Test 3 is greater than .05. The model chosen appears to adeguately describe the
data
Benchmark Dose Computation
Specified effect = 0.05
Risk Type = Relative risk
Confidence level = 0.95
BMD = 2056.33
BMDL = 900.888
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Relative risk
Confidence level = 0.95
BMD = 2908.09
BMDL = 1801.78
B-10
-------
Output B-6: Increased Incidence of Extra Ribs in Sprague-Dawley Rats (Deacon et al.,
1981)
NCTR Model with 0.95 Confidence Level
0.05
"S 0.04
"u
^ 0.03
c
o
"o
E 0.02
u_
0.01
0
0.06 -NCTR
BMDL
BMD
500 1000 1500 2000
dose
2500
3000
3500
16:36 03/05 2003
BMDS MODEL RUN: NCTR Model
The probability function is:
Prob. = 1 - exp[-(alpha + thl*Rij) - (beta + th2*Rij)*Dose^rho],
where Rij is the centralized litter specific covariate.
Restrict Power rho >= 1.
Total number of observations = 82
Total number of records with missing values = 0
Total number of parameters in model = 9
Total number of specified parameters = 6
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
**** We are sorry but Relative Function and Parameter Convergence ****
**** are currently unavailable in this model. Please keep checking ****
**** the web sight for model updates which will eventually ****
**** incorporate these convergence criterion. Default values used. ****
User specifies the following parameters:
thetal = 0
theta2 = 0
phil = 0
phi2 = 0
phi3 = 0
phi4 = 0
Default Initial Parameter Values
alpha = 0.00255273
beta = 2.32553e-013
rho = 3.18392
Warning: Maximum iteration may be not large enough. Iterations reach the maximum.
Parameter Estimates
Variable
Estimate
Std. Err.
B-ll
-------
alpha
beta
rho
0.00252937
1.09266-022
5.87127
0.00183262
3 .334156-019
381.143
P-value
Analysis of Deviance Table
Model Log(likelihood) Deviance Test DF
Full model -29.869
Fitted model -45.2723 30.8066 80
Reduced model -51.542 43.3461 81
AIC: 94.D446 AIC=-2*Log(likelihood)-2*p=96.5446
Goodness of Fit
Dose
400.
400.
400.
400.
400.
400.
400.
1000.
1000.
1000.
1000.
1000.
1000.
1000.
1000.
1000.
1000.
3000.
3000.
3000.
3000.
3000.
3000.
3000.
3000.
3000.
3000.
3000.
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
Litter Size
Est. Prob.
Expected
Observed
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
.0000
4
8
10
11
12
13
14
15
16
17
0
0
0
0
0
0
0
0
0
0
.003
.003
.003
.003
.003
.003
.003
.003
.003
.003
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
.010
.020
.025
.056
.121
.263
.177
.076
.040
.043
0
0
0
0
0
1
0
0
1
0
9
11
12
13
14
15
5
6
9
10
11
12
13
14
15
16
11
12
13
14
15
16
17
19
Chi-square =
34 .20
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.030
0.030
0.030
0.030
0.030
0.030
0.030
0.030
0.030
0.030
0.030
DF = 35
0.040
0.023
0.028
0.061
0.197
0.212
0.038
0.013
0.015
0.023
0.051
113
0.031
0.067
0.072
0.154
0.041
0
0.122
0.244
0.274
1.006
0.366
1.188
1.280
0.914
0.488
0.518
0.579
P-value = 0.5064
To calculate the BMD and BMDL, the litter specific covariate is fixed at the mean litter specific
covariate of control group: 12.653846
Benchmark Dose Computation
Specified effect = 0.05
Risk Type = Extra risk
Confidence level = 0.950000
BMD = 3317.45
BMDL = 2992.69
B-12
-------
Output B-7: Reduced Fetal Weight in CD-I Mice (Schwetz et al., 1991/Mast et al., 1989)
Linear Model with 0.95 Confidence Level
.38
1.36
en i •-"*
c
o
, .-,n
a) 1.32
cr
c
05
Oj
500
1000
1500 2000
dose
2500
3000 3500
16:4303/132003
The form of the response function is:
Y[dose] = beta_0 + beta_l*dose + beta_2*dose^2 + ...
Dependent variable = MEAN
Independent variable = Dose
rho is set to 0
Signs of the polynomial coefficients are not restricted
A constant variance model is fit
Total number of dose groups = 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
Default Initial Parameter Values
alpha = 0.0050202
rho = 0 Specified
beta_0 = 1.35314
beta 1 = -2.09075e-005
Variable
alpha
beta_0
beta 1
Parameter Estimates
Estimate Std. Err.
0.00483473 0.000673703
1.3529 0.00958809
-2.08259e-005 5.75863e-006
Asymptotic Correlation Matrix of Parameter Estimate
alpha beta_0 beta_l
alpha 1 -8.6e-009 6e-009
beta_0 -8.6e-009 1 -0.7
beta 1 6e-009 -0.7 1
B-13
-------
Model Descriptions for likelihoods calculated
Likelihoods of Interest
Model Log(likelihood) DF AIC
Al 223.195553 5 -436.391107
A2 224.250451 8 -432.500903
fitted 223.094409 2 -442.188818
R 216.935622 2 -429.871245
Tests of Interest
Test -2*log(Likelihood Ratio) Test df
Test 1 14.6297 6
Test 2 2.1098 3
Test 3 0.202289 2
The p-value for Test 3 is greater than .05. The model chosen appears
to adeguately describe the data
Benchmark Dose Computation
Specified effect = 1
Risk Type = Estimated standard deviations from the control mean
Confidence level = 0.95
BMD = 3338.74
BMDL = 2272.53
B-14
-------
Output B-8: Incidence of Misaligned Sternebrae in CD-I Mice (Schwetz et al.,
1991/Mast et al., 1989)
0.25
Nested Logistic Model with 0.95 Confidence Level
o
I
0.15
0.05
15:4003/252003
3000
3500
The probability function is:
Prob. = alpha + thetal*Rij + [1 - alpha - thetal*Rij]/
[1+exp(-beta-theta2*Rij-rho*log(Dose
where Rij is the litter specific covariate.
Restrict Power rho >= 1.
Default Initial Parameter Values
alpha = 0.102954
beta = -10.2672
thetal = 0
theta2 = 0
rho = 1
phil = 0.0232129
phi2 = 0.0522594
phi3 = 0.107766
phi4 = 0.0855581
B-15
-------
Variable
alpha
beta
rho
phil
phi2
phi3
phi4
Model
Full model
Fitted model
Reduced model
Parameter Estimates
Estimate
0.102937
-10.2671
1
0.023208
0.052033
0.107815
0.0855465
Analysis of Deviance Table
Log (likelihood) Deviance Test DF
-377.311
-462.552 170.481 97
-478.095 201.568 102
P-value
Dose
0.
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
398.
1010.
1010.
1010.
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
. 0000
. 0000
Lit. -Spec.
Cov .
5 .
7 .
10.
11.
11.
11.
11.
11.
11.
11.
11.
12 .
12.
12.
12.
12.
12 .
13.
13.
14.
14 .
14 .
14 .
15.
15.
16.
6.
8.
8.
8.
8.
10.
10.
11.
11.
11.
11.
11.
12.
12 .
13.
13.
13.
13.
13.
14 .
14 .
15.
15.
^
9 .
9 .
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
. 0000
. 0000
Est.
0.
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0 .
0 .
Litter Data
Litter
. Prob. Size
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.103
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.115
.133
.133
.133
5
7
10
11
11
11
11
11
11
11
11
12
12
12
12
12
12
13
13
14
14
14
14
15
15
16
6
8
8
8
8
10
10
11
11
11
11
11
12
12
13
13
13
13
13
14
14
15
15
^
9
9
chi-squared
Expected Observed Residual
0.
0.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
0.
0 .
0 .
0 .
0 .
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
1.
0.
1.
1.
.515
.721
. 029
.132
.132
.132
.132
.132
.132
.132
.132
. 235
.235
.235
.235
.235
.235
.338
.338
.441
. 441
. 441
. 441
.544
.544
.647
.691
. 921
. 921
. 921
. 921
.152
.152
.267
.267
.267
.267
.267
.382
.382
.497
.497
.497
. 497
. 497
. 612
. 612
.728
.728
. 934
. 200
. 200
0
0
1
4
1
4
0
1
1
1
0
2
1
0
2
0
0
0
1
2
2
4
1
1
1
1
0
0
0
2
0
0
2
1
1
1
1
1
0
4
0
0
1
5
0
2
3
1
2
4
3
1
-0.
-0.
-0.
2 .
-0 .
2.
-1.
-0.
-0.
-0.
-1.
0.
-0 .
-1.
0 .
-1.
-1.
-1.
-0.
0.
0.
1.
-0 .
-0 .
-0.
-0.
-0.
-0 .
-0 .
1.
-0 .
-0.
0.
-0.
-0.
-0 .
-0.
-0 .
-0 .
1.
-1.
-1.
-0.
2.
-1.
0 .
0 .
-0.
0.
2.
1.
-0 .
.7246
. 8397
. 0278
.5634
.1183
.5634
. 0122
.1183
.1183
.1183
. 0122
. 6484
.1995
. 0474
. 6484
. 0474
.0474
. 0802
.2730
. 4308
. 4308
. 9726
.3400
. 4016
.4016
.4584
.7873
. 8737
. 8737
. 0227
. 8737
. 9416
. 6934
.2045
.2045
.2045
.2045
.2045
. 9967
.8879
.0207
.0207
.3390
.3877
. 0207
. 2506
. 8972
.4476
.1675
.6566
.2930
.1439
B-16
-------
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
1010.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
3020.
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
9 .
10.
10.
10.
11.
11.
11.
11.
11.
11.
11.
12.
12.
12.
12 .
12 .
12 .
12 .
12.
13.
13.
14 .
14.
7.
9
9 .
9 .
10.
10.
10.
10.
11.
11.
11.
11.
11.
11.
12 .
12 .
12 .
12 .
12.
12.
13.
13.
13.
14.
14.
14.
15.
15.
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
. 0000
. 0000
.0000
.0000
.0000
.0000
. 0000
. 0000
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0.
0 .
0 .
0 .
0 .
0 .
0.
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0.
0 .
0 .
0 .
0 .
0.
0.
0.
0.
0 .
0 .
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.133
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
.188
10
10
10
11
11
11
11
11
11
11
12
12
12
12
12
12
12
12
13
13
14
14
9
9
10
10
10
10
11
11
11
11
11
11
12
12
12
12
12
12
13
13
13
14
14
14
15
15
1.200
1.334
1.334
1.334
1. 467
1.467
1.467
1.467
1.467
1. 467
1. 467
1. 600
1. 600
1. 600
1.600
1.600
1.600
1.600
1. 600
1.734
1.734
1. 867
1.867
1.317
1.693
1. 693
1. 693
1. 882
1. 882
1.882
1.882
2.070
2.070
2. 070
2. 070
2. 070
2. 070
2.258
2.258
2.258
2.258
2.258
2.258
2.446
2.446
2.446
2. 634
2. 634
2. 634
2. 822
2. 822
0
4
0
0
0
5
2
1
1
3
3
0
0
4
3
2
1
0
1
5
2
4
0
1
0
0
3
1
2
0
3
4
0
2
2
3
3
2
2
4
1
1
1
2
0
0
0
8
6
3
4
-0.
1.
-0 .
-0 .
-0 .
2 .
0.
-0.
-0.
0 .
0 .
-0 .
-0.
1.
0.
0.
-0.
-0.
-0 .
1.
0 .
1.
-0.
-0.
-1.
-1.
0 .
-0 .
0.
-1.
0.
1.
-1.
-0 .
-0 .
0 .
0 .
-0.
-0.
0.
-0.
-0.
-0 .
-0 .
-1.
-1.
-1.
2 .
1.
0 .
0 .
.8623
.7669
. 8837
. 8837
. 9025
.1736
.3279
.2873
.2873
. 9432
. 9432
. 9191
. 9191
.3782
.8039
. 2295
.3448
. 9191
.3448
.7594
.1435
. 0820
. 9471
.2493
.1128
.1128
. 8586
.5361
. 0720
.1443
. 6802
.0932
.1722
. 0395
. 0395
.5269
.5269
.1367
.1367
. 9236
. 6669
.6669
.6669
.2223
.2193
.2193
.2394
.5248
.5837
. 0792
.5248
Combine litters with adjacent levels of the litter-specific covariate within dose groups until
the expected count exceeds 3.0, to help improve the fit of the X^2 statistic to chi-squared.
B-17
-------
8.0000
1
3
6
14
11
2
14
DF = 22
B-18
-------
NCTR Model with 0.95 Confidence Level
0.22
0.2
„ 0.18
OJi
I 0.16
I 0-14
"o
E 0.12
0.1
0.08
NCTR
0
10:3603/262003
BMDS MODEL RUN
BMDL
BMD
500
1000
1500
dose
2000
2500
3000 3500
The probability function is:
Prob. = 1 - exp[-(alpha + thl*Rij) - (beta + th2*Rij)*Dose^rho],
where Rij is the centralized litter specific covariate.
Restrict Power rho >= 1.
Total number of observations = 103
Total number of records with missing values = 0
Total number of parameters in model = 9
Total number of specified parameters = 2
^*** We are sorry but Relative Function and Parameter Convergence
**** are currently unavailable in this model. Please keep checking
t*** the web sight for model updates which will eventually
**** incorporate these convergence criterion. Default values used.
User specifies the following paramters:
thetal = 0
theta2 = 0
Default Initial Parameter Values
alpha = 0.109279
beta = 3.27008e-005
rho = 1
phil = 0.0233469
phi2 = 0.0519936
phi3 = 0.10788
phi4 = 0.0854085
Variable
alpha
beta
rho
Parameter Estimates
Estimate Std. Err.
0.109279 0.0217683
3.27008e-005 0.000239898
1 0.915599
B-19
-------
phil
phi2
phi3
phi4
Model
Full model
Fitted model
Reduced model
Analysis of Deviance Table
Log(likelihood) Deviance Test DF
-377.311
-462.583 170.542 98
-478.095 201.568 102
\IC:
Goodness of Fit
Litter_Size Est._Prob. Expected
10
11
12
13
14
15
16
10
11
12
13
14
15
9
10
11
12
13
14
10.
4
4
4
15
11
9
10
11
12
13
14
15
14
11
14
Chi-square =
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.950000
BMD = 3221.96
BMDL = 1788.93
B-20
-------
0.22
0.2
0.18
0.16
.1 0.14
0.12
0.1
0.08
RaiVR
RaiVR Model with 0.95 Confidence Level
BMDL
BMD
500
1000
1500
dose
2000
2500
3000
3500
10:3703/262003
The probability function is:
Prob. = [l-exp(-Alpha-Beta*Dose^Rho)]*exp(-(Thl + Th2*Dose)*Rij ) ,
where Rij is the litter specific covariate.
Restrict Power rho >= 1.
Total number of observations = 103
Total number of records with missing values = 0
Total number of parameters in model = 9
Total number of specified parameters = 2
Variable
alpha
beta
rho
phil
Default Initial Parameter Values
alpha = 0.109279
beta = 3.27008e-005
rho = 1
phil = 0.0233469
phi2 = 0.0519936
phi3 = 0.10788
phi4 = 0.0854085
Parameter Estimates
Estimate Std. Err.
0.109279 0.0217683
3.27008e-005 0.000239898
1 0.915599
0.0233469 0.0343278
B-21
-------
phi2
phi3
phi4
10
11
12
13
14
15
16
0
0
1
12
5
1
9
10
11
12
13
14
15
9
10
11
12
13
14
4
4
4
15
11
Chi-square =
9
10
11
12
13
14
15
14
11
14
P-value = 0.4877
To calculate the BMD and BMDL, the litter specific covariate is fixed at the mean litter specific
covariate of control group: 11.923077
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.950000
BMD = 3221.96
BMDL = 1788.93
B-22
------- |