r/EPA
EPA 635/R-03/008
www.epa.gov/iris
TOXICOLOGICAL REVIEW
OF
CYCLOHEXANE
(CAS No. 110-82-7)
In Support of Summary Information on the
Integrated Risk Information System (IRIS)
August 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.
11
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CONTENTS — TOXICOLOGICAL REVIEW OF
CYCLOHEXANE (CAS No. 110-82-7)
LIST OF TABLES v
LIST OF FIGURES vi
FOREWORD vii
AUTHORS, CONTRIBUTORS, AND REVIEWERS viii
1. INTRODUCTION 1
2. CHEMICAL AND PHYSICAL INFORMATION
RELEVANT TO ASSESSMENTS 3
3. TOXICOKINETICS RELEVANT TO ASSESSMENTS 5
3.1. ABSORPTION 5
3.2. DISTRIBUTION 7
3.3. METABOLISM 11
3.4. EXCRETION 13
4. HAZARD IDENTIFICATION 18
4.1. STUDIES IN HUMANS—EPIDEMIOLOGY, CASE REPORTS, CLINICAL
CONTROLS 18
4.2. PRECHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN
ANIMALS—ORAL AND INHALATION 20
4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND
INHALATION 24
4.4. OTHER STUDIES 29
4.4.1. Acute Toxicity 29
4.4.2. Neurological Studies 30
4.4.3. Genotoxicity 32
4.5. SYNTHESIS AND EVALUATION OF MAJORNONCANCER EFFECTS AND
MODE OF ACTION 33
4.6. WEIGHT-OF-EVIDENCE EVALUATION AND CANCER
CHARACTERIZATION 35
4.7. SUSCEPTIBLE POPULATIONS 36
4.7.1. Possible Childhood Susceptibility 36
4.7.2. Possible Gender Differences 36
5. DOSE-RESPONSE ASSESSMENTS 38
5.1. ORAL REFERENCE DOSE (RfD) 38
5.2. INHALATION REFERENCE CONCENTRATION (RfC) 38
5.2.1. Choice of Principal Study and Critical Effect—with Rationale and
Justification 38
iii
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5.2.2. Methods of Analysis—Benchmark Concentration Analysis 41
5.2.3. RfC Derivation—Including Application of Uncertainty Factors (UFs) . . 44
5.3. CANCER ASSESSMENT 45
6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION
OF HAZARD AND DOSE RESPONSE 46
6.1. HUMAN HAZARD POTENTIAL 46
6.2. DOSE-RESPONSE 47
7. REFERENCES 49
APPENDIX A: SUMMARY OF EXTERNAL PEER REVIEW COMMENTS AND
DISPOSITION
APPENDIX B: BENCHMARK DOSE
IV
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LIST OF TABLES
Table 3-1. Uptake and disappearance of cyclohexane
in mouse blood after inhalation exposure 8
Table 3-2. Tissue, saline, blood, and oil/air partition coefficients for cyclohexane 9
Table 3-3. Cyclohexane burdens in rat cerebrum and
perirenal fat after repeated exposure 10
Table 3-4. Concentration of cyclohexane in rat tissues after repeated exposure 10
Table 5-1. Mean pup weights (G)—Fl generation 39
Table 5-2. Mean pup weights (G)-F2 generation 39
Table B-l. Summary of model outcomes, reproductive
toxicity/multigenerational study (DuPontHLR, 1997a) B-3
Table B-2. Summary of model fits, reproductive toxicity/multigenerational
study (DuPontHLR, 1997a) B-4
Tables of Standardized Residuals (Tables B-3a - B-3f)
Table B-3a. Fl: linear, constant variance model for average d25 pup weight B-6
Table B-3b. Fl: linear, constant variance model for average pup weight gain d25-d7 B-6
Table B-3c. Fl: linear, heterogeneous variance model for average pup weight gain d25-d7 B-6
Table B-3d. Fl: quadratic, heterogeneous variance model for average pup
weight gain d25-d7 B-6
Table B-3e. F2: linear, constant variance model for average pup weight gain d25-d7 B-7
Table B-3f F2: quadratic, constant variance model for average pup weight gain d25-d7 B-7
Tables B-4 - B-6 and Figure 2 refer to the results from the rat developmental study
(DuPontHLR, 1997b)
Table B-4. Summary of model outcomes, developmental study B-l 1
Table B-5. Summary of model fits, developmental study B-12
Tables B-6a - B-6c. Standardized Residuals (DuPont HLR, 1997b)
Table B-6a. Linear, heterogeneous variance model for average dam weight
gain, gd!7-gd7 B-13
Table B-6b. Linear, constant variance model for average dam weight gain, gd!7-gd7 B-13
Table B-6c. Linear, constant variance model for average dam body weight, gd!7 B-13
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LIST OF FIGURES
Figure B-la. Fl: linear, constant variance model for average d25 pup weight B-8
Figure B-lb. Fl: linear, constant variance model for average pup weight gain d25-d7 B-8
Figure B-lc. Fl: linear, heterogeneous variance model for average pup weight
gain d25-d7 B-9
Figure B-ld. Fl: quadratic, heterogeneous variance model for average pup weight
gain d25-d7 B-9
Figure B-le. F2: linear, constant variance model for average pup weight gain d25-d7 B-10
Figure B-lf. F2: quadratic, constant variance model for average pup weight gain d25-d7 . . B-10
Figure B-2a. Linear, heterogeneous variance model for average dam weight gain, gd!7-gd7
(DuPont HLR, 1997b) B-14
Figure B-2b. Linear, constant variance model for average dam weight gain, gd!7-gd7
(DuPont HLR, 1997b) B-14
Figure B-2c. Linear, constant variance model for average dam body weight, gd!7 B-15
VI
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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
cyclohexane. It is not intended to be a comprehensive treatise on the chemical or toxicological
nature of cyclohexane.
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 301-345-2870.
vn
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AUTHORS, CONTRIBUTORS, AND REVIEWERS
CHEMICAL MANAGER
Colette S. Hodes, Ph.D.
Office of Pollution Prevention and Toxics
U.S. Environmental Protection Agency
Washington, DC
AUTHORS
Colette S. Hodes, Ph.D.
Office of Pollution Prevention and Toxics
U.S. Environmental Protection Agency
Washington, DC
Elizabeth H. Margosches, Ph.D.
Office of Pollution Prevention and Toxics
U.S. Environmental Protection Agency
Washington, DC
Andrea Pfahles-Hutchens
Office of Pollution Prevention and Toxics
U.S. Environmental Protection Agency
Washington, DC
Martha G. Price, Ph.D.
Office of Pollution Prevention and Toxics
U.S. Environmental Protection Agency
Washington, DC
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
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.
Vlll
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INTERNAL EPA REVIEWERS
Angela Auletta, Ph.D.
Office of Pollution Prevention and Toxics
U.S. Environmental Protection Agency
Washington, DC
Susan Rieth, M.P.H.
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
Karen Hogan
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
Meena Sonawane
Office of Pollution Prevention and Toxics
U.S. Environmental Protection Agency
Washington, DC
EXTERNAL PEER REVIEWERS
Raymond B. Baggs, Ph.D., D.V.M.
University of Rochester Medical Center
Division of Laboratory Animal Medicine
Rochester, NY
Ronald David Hood, Ph.D.
Department of Biological Sciences
The University of Alabama
Tuscaloosa, AL
AnnaM. Fan, Ph.D.
Office of Environmental Health Hazard
Assessment
California Environmental Protection Agency
Oakland, CA
Summaries of the external peer reviewers' comments and the disposition of their
recommendations are in Appendix A.
IX
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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
cyclohexane 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
1
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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, 1991), Guidelines for Reproductive
Toxicity Risk Assessment (U.S. EPA, 1996), Guidelines for Neurotoxicity Risk Assessment (U.S.
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 databases were searched: RTECS,
HSDB, TSCATS, CCRIS, GENE-TOX, DART/ETIC, EMIC, TOXLINE, CANCERLIT, and
MEDLINE. Any pertinent scientific information submitted by the public to the IRIS Submission
Desk was also considered in the development of this document. The relevant literature was
reviewed through March 2003.
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2. CHEMICAL AND PHYSICAL INFORMATION
RELEVANT TO ASSESSMENTS
Common synonyms for cyclohexane include hexahydrobenzene, hexamethylene, and
hexanaphthene (Merck Index, 1996). Some relevant physical and chemical properties of
cyclohexane are listed below.
CASRN
Empirical formula
Molecular weight
Physical state
Color
Odor
Boiling point (°C)
Melting point (°C)
Log Kow
Vapor pressure (at 25°C)
Water solubility (at 25°C)
Explosive Limits
Conversion factors
110-82-7
C6H12
84.2
Liquid (20°C)
Colorless
Solvent odor
80.7
6.47
3.44
97 mm Hg
55mg/L
LEL= 1.3%
UEL = 8%
1 ppm = 3.44 mg/m3
1 mg/m3 = 0.291 ppm
NIOSH, 1997
Merck Index, 1996
Merck Index, 1996
Merck Index, 1996
NIOSH, 1997
Merck Index, 1996
Merck Index, 1996
Merck Index, 1996
Hanschetal., 1995
Chaoetal., 1983
Verschueren, 1996
NIOSH, 1997
NIOSH, 1997
The primary use of cyclohexane is in the production of nylon. A total of 55% is used to
produce adipic acid and 26% is used to formulate caprolactam, both of which are then used to
generate nylon. Another 13% is exported and the remaining 6% is used in solvents, insecticides,
and plasticizers (Kavaler, 1998). The United States accounts for about one-third of the world's
consumption of cyclohexane, or about 1 billion gallons per year (Eastman and Mears, 1995). In
1991, the total U.S. production of cyclohexane was 3.55 x 108 gallons. Cyclohexane is present
in all crude oils in concentrations ranging from 0.1 to 1.0%, and it is also found in gasoline
formulations (Eastman and Mears, 1995). It has also been detected in volcanic emissions and in
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plant volatiles (Graedel, 1978). The general population is primarily exposed to cyclohexane
through the inhalation of ambient air due to its presence in gasoline vapors. The average
concentration of cyclohexane in the exhaust of six cars was 82 ppb (Blake et al., 1993).
In the ambient atmosphere, cyclohexane is expected to exist solely in the vapor phase
(Bidleman, 1988), based on a measured vapor pressure of 97 mm Hg at 25°C (Chao et al., 1983).
Vapor-phase cyclohexane is degraded by reacting with photochemically-produced hydroxyl
radicals (Atkinson, 1989), with an estimated half-life of 45 hours. Cyclohexane is expected to
have moderate mobility in soils, based on an estimated Koc value of 160, determined by a
structure estimation method that uses molecular connectivity indices (Meylan et al., 1992).
Volatilization from water surfaces is expected to be an important fate process, based on its
Henry's Law constant of 0.15 atm m3/mol (Bocek, 1976). Estimated volatilization half-lives for
a model river and model lake are 1 hour and 3.6 days, respectively (Meylan and Howard, 1991).
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3. TOXICOKINETICS RELEVANT TO ASSESSMENTS
3.1. ABSORPTION
Cyclohexane is rapidly absorbed into the blood via the lungs, gastrointestinal tract, and
skin. At higher doses, some cyclohexane is expired unchanged due to preferential partitioning to
alveoli rather than the blood, where it has low solubility (see Section 3.4.).
Several occupational monitoring studies of workers exposed via inhalation demonstrate
rapid uptake by the human body (Brugnone et al., 1980; Mutti et al., 1981; Perbellini and
Brugnone, 1980; Yasugi et al., 1994). Perbellini and Brugnone (1980) found strong correlations
between cyclohexane levels in factory air and alveolar air, and between alveolar air and blood
concentrations in 22 shoe factory workers during hours 4-8 of their work shift. Breathing space
air concentrations ranged from 17 to 2,484 mg/m . The exposed factory workers exhibited mean
levels of cyclohexane in alveolar air that corresponded to 78% of the workplace concentration.
Blood levels ranged from 29 to 367 [ig/L (mean 158 |-ig/L), corresponding to 53 to 78% of
alveolar concentration.
Yasugi et al. (1994) assessed a cohort of 33 female workers exposed 8 hours per day for
at least 1 year to glue solvent containing by volume up to 83% cyclohexane, 16% toluene, and
0.9% hexane. Measured by personal monitors, the geometric mean and maximum
concentrations of cyclohexane in breathing zone air were 27 and 274 ppm, respectively. Toluene
and hexane levels were 2.8 ppm (maximum 11 ppm) and 1.4 ppm (maximum 12 ppm),
respectively. Cyclohexane concentrations in blood samples collected at the end of the
workweek's last shift correlated significantly (p < 0.01) with exposure. Serum analyses of
various liver and kidney functions gave normal results. The analyses included total protein,
blood urea nitrogen, creatinine, uric acid, total cholesterol, triglycerides, HDL-aminotransferase,
aspartate aminotransferase, alanine aminotransferase, y-glutamyl transpeptidase, alkaline
phosphatase, leucine aminopeptidase, and lactate dehydrogenase.
Mutti et al. (1981) measured the total lung uptake of eight shoe factory workers during a
4-hour exposure period. Workplace air contained from 52.7 ± 7.1 to 266.5 ±11.2 mg/m
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cyclohexane. Results indicate that 34% of the alveolar cyclohexane and 23% of the total
respiratory intake was absorbed into the pulmonary blood. The total mean intake and uptake
were calculated at 354 ± 12 mg and 81.2 mg cyclohexane, respectively.
Brugnone et al. (1980) studied industrial exposure to cyclohexane in different factories.
Alveolar air and breathing zone air samples (n = 108) were collected simultaneously during the
work shift. The ratio of alveolar air (Ca) to workplace air (Ci) was high and gave a linear
correlation (r = 0.98). Alveolar retention (defined as Ci - Ca) rose constantly during hours 4 to 8
of the work shift. A linear correlation (r = 0.82) was found between workplace air concentration
and alveolar retention.
Uptake has also been demonstrated in occupational settings by examining urinary
metabolites of cyclohexane (see Section 3.4.1.) (Governa et al., 1987; Mraz et al., 1994;
Perbellini et al., 1980, 1987; Yasugi et al., 1994; Yuasa et al., 1996).
Absorption of cyclohexane via the lungs, gastrointestinal tract, and skin has been
described in several animal species. [ C]Cyclohexane administered to rats either by oral gavage
or by intravenous injection was rapidly absorbed and distributed to the tissues (RTI, 1984).
Savolainen and Pfaffli (1980) and Zahlsen et al. (1992) repeatedly exposed rats via inhalation
and found cyclohexane in blood, brain, fat, and other tissues. Rabbits exposed orally or via
inhalation excreted cyclohexane metabolites in their urine (Treon et al., 1943a, b; Elliott et al.,
1959).
Naruse (1984) exposed mice to varying amounts of an adhesive containing approximately
7.5, 13, or 19 g cyclohexane in a closed chamber. Air levels of cyclohexane stabilized after 1
hour, producing concentrations of 8,000, 14,000, or 17,500 ppm, respectively. Corresponding
blood concentrations were determined to be 27, 69, and 122 |_ig/mL, respectively, and were
correlated (r = 0.99) with the total dose introduced into the exposure chamber (Table 3-1).
Cyclohexane is absorbed through the skin surface. lyadomi et al. (1998) measured the
time course of dermal absorption in male WBN/ILA-Ht hairless rats. Solvent chambers glued to
<-\
abdominal skin allowed contact over a 6.28 cm area. Cyclohexane (1 mL) was injected into the
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chamber, and carotid blood samples (n = 8) were drawn from 5 minutes to 4 hours after initiation
of exposure. Arterial cyclohexane concentration increased rapidly up to 30 min, peaking in
about 1 hour at approximately 0.24 |amol/L. Thereafter, blood levels decreased in a linear
fashion. A rough blood clearance time was calculated as 400 minutes.
3.2. DISTRIBUTION
Perbellini et al. (1985) and Gargas et al. (1989) used different experimental methods to
determine physiologically based partition coefficients for cyclohexane in human cadaver and rat
tissues (Table 3-2). For volatile compounds such as cyclohexane, the air-to-blood and the blood-
to-tissue concentration ratios factor significantly in blood uptake and subsequent distribution to
the tissues. Not surprisingly for a nonpolar organic compound, cyclohexane partitions
preferentially to lipid-rich tissues such as fat, liver, and brain. The values obtained for muscle
tissue differ between the two studies.
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Table 3-1. Uptake and disappearance of cyclohexane
in mouse blood after inhalation exposure
Cage
dosea
7.5 g
13 g
19 g
Measured
levels"
AlRppm
(g/m3)
BLOOD
|j,g/mL
AlRppm
(g/m3)
BLOOD
Hg/mL
AlRppm
(g/m3)
BLOOD
Hg/mL
Exposure time
30 minutes
7
41
71
1 hour
8,000
(27.5)
27
14,000
(48.2)
69
17,500
(60.2)
122
Post-exposure time
30 minutes
7 (25%)
9 (13%)
16 (13%)
1 hour
4(15%)
5 (7%)
8 (7%)
2 hours
1 (4%)
2 (3%)
3 (2%)
3 hours
1
a Cage dose, cyclohexane grams in the chamber, calculated from Table I and Figure 8 of Naruse (1984).
b Measured levels in AIR calculated from Figure 8 of Naruse (1984).
Sikkema et al. (1994) found that cyclohexane preferentially accumulated in a microbial
phospholipid bilayer membrane (liposome) and calculated the partition coefficient for the bilayer
membrane to the potassium phosphate buffer system to be 498.
Cyclohexane has been detected in human milk in 5 out of 12 samples collected in urban
areas of the United States. The fat content of mothers' milk, approximately 3.8%, would
promote partition (Pellizzari et al., 1982). No other studies addressing the distribution of
cyclohexane in humans were found.
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Table 3-2. Tissue, saline, blood, and oil/air partition coefficients for cyclohexane
Tissue/substance
0.9% saline
blood
lung
heart
kidney
liver
brain
muscle
fat
olive oil
Perbellini et al. (1985)a
-
1.3±0.1
2.7 ± 0.1
5.8 ± 1.0
7.2 ± 1.0
10.8 ±0.9
10.7 ± 1.4
10.5 ±0.7
260 ± 11.0
293 ± 11.0
Gargas et al. (1989)
< 0.01 (approximated)
1.41 ±0.14 (human)
1.39 ±0.09 (rat)
-
-
-
7. 88 ±0.59 (rat)
-
1.03 ±0.17 (rat)
235 ±4 (rat)
293 ±2
a Two male cadavers, ages 30 and 40, cause of death, heart attack.
Existing studies provide information on the distribution of cyclohexane in rats. Rats
exposed repeatedly for 1 week to 300, 1,000, or 2,000 ppm cyclohexane exhibited body burdens
in fat and brain that reflect the ratios of the partition coefficients for these tissues. After 2 weeks
of exposure, however, the cyclohexane concentrations in fat increased disproportionately (Table
3-3). The body burden in perirenal fat was proportional (r > 0.95) to exposure levels and an
order of magnitude higher than that found in brain (Savolainen and Pfaffli, 1980).
Zahlsen et al. (1992) studied the distribution of cyclohexane in male Sprague-Dawley
rats up to 3 days of exposure (12 hrs/day) to a mean level of 100.4 ppm (range: 96-107 ppm).
The mean concentration in tissues was relatively steady during the 3 days of exposure except in
fat, where mean levels increased daily (Table 3-4). The high concentration of cyclohexane in
kidney cannot be explained by its solubility in kidney tissues alone. The authors concluded that
cyclohexane toxicokinetics are complex and cannot be extrapolated from solubility properties
alone.
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Table 3-3. Cyclohexane burdens in rat cerebrum and
perirenal fat after repeated exposure
Dose
300 (ppm)
1,000 (ppm)
2,000 (ppm)
Fat (nmol/g tissue)
1 week 2 weeks
538 ±43 483 ±106
2340 ± 592 1706 ± 199
3542 ± 658 2748 ± 806
Cerebrum (nmol/g)
1 week 2 weeks
14 ±8 8±2
101 ±28 21±10a
150±31 55±7a
Fat:Brain Ratio
1 week 2 weeks
38 60
23 81
24 50
a differs from the first week value at p < 0.001.
Source: Adapted from Savolainen and Pfaffli, 1980.
rl4,
Seventy-two hours after a single intravenous dose of 10 mg/kg [ C]cyclohexane or a
single oral dose of 200 mg/kg to adult male Fischer 344 rats, the concentration of radioactivity in
adipose was 16 times greater than that in blood. At higher oral doses (1,000 or 2,000 mg/kg), the
adipose tissue-to-blood ratio of radioactivity approximately 45. Although radioactivity in
adipose tissues was primarily cyclohexane (79-84%), in muscle, liver and skin, only 2-18% of
the C was identified as cyclohexane. Cyclohexane, cyclohexanol, and cyclohexanone were
present in all tissues (RTI, 1984).
Table 3-4. Concentration of cyclohexane in rat tissues after repeated exposure
Tissue
Blood
Liver
Brain
Kidney
Fat
Exposure Levels
(jimol/kg) (% change from previous day)
Dayl
4.0 ±0.3
22.6 ±3.0
31.7 ±2.2
86.5 ±2.0
417 ±66
Day 2
4.4 ±0.1
22.3 ±2.9
33.6 ±3.2
100.1 ±10.3
475 ± 27 (14%)
Day 3
4.1 ±0.9
26.4 ± 1.7
34.7 ± 1.1
99.4 ±13.0
482 ±17 (1.5%)
Mean
4.2
23.8
33.3
95.3
-
Post Exposure
(junol/kg)
Day 3 + 12 hours
0.1±0.1
0.5 ±0.4
2.0 ±2.5
1.3 ±0.1
169.17
- = not calculated by study authors, values still increasing.
Source: Adapted from Zahlsen et al. (1992).
10
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3.3. METABOLISM
Metabolic studies of the microsomal mixed-function oxidase (monooxidase) system in
liver confirm hydroxylation of cyclohexane to cyclohexanol. Cyclohexanol is the primary
metabolite of cyclohexane; however, lesser amounts of cyclohexanone and 1,2-cyclohexane-diol
have been identified. Cyclohexyl metabolites are conjugated to glucuronides for excretion, but
at high doses sulfate conjugation may occur. Information on the metabolic pathways of
cyclohexane is insufficient. No human studies of cyclohexane metabolism were found in the
literature.
A metabolic adaptation in mice to repeated exposures of cyclohexane was observed by
Naruse (1984). Immediately following a 1-hour exposure period to 14,000 ppm cyclohexane,
blood levels averaged 69 |_ig/mL and cleared within 3 hours. However, after 120 days of
repeated exposure (1 hr/day, 6 days/wk), blood levels measured immediately after exposure
dropped to 30 |_ig/mL, clearing within 2 hours.
Evidence suggestive of mixed-function oxidase activation in rats exposed to cyclohexane
was observed by Savolainen and Pfaffli (1980). Male Wistar rats were exposed repeatedly (6
hrs/day, 5 days/wk for 1-2 weeks) to cyclohexane vapor (300, 1,000, or 2,000 ppm) and the
body burden was examined (Table 3-3). After the first week of exposure, a linear correlation (r
= 0.99) was found between brain and fat cyclohexane levels. By the end of the second week,
cyclohexane levels had decreased in both tissues, indicating metabolic adaptation. Elimination
of cyclohexane from the brain was particularly enhanced when compared to body fat. This
changed the relationship between tissue concentrations to one described best by an exponential
function.
In these experiments, measures of brain metabolism, RNA, glutathione, and glutathione
peroxidase activity were not affected by cyclohexane exposure. However, the cyclohexane dose
increased, brain azoreductase activity decreased significantly and was still well below control
levels after a recovery period. Activation of the mixed-function oxidase system has been found
to inhibit azo reduction (Klaassen et al., 1986). The authors suggested that although increased
11
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blood circulation in the brain compared to fatty tissue enhances cyclohexane elimination from
the brain, the activation of a liver mixed-function oxidase system is the primary vehicle for
decreasing the cyclohexane concentration in the body as a whole.
Two studies describe cyclohexane metabolism in vivo (Elliott et al., 1959; RTI, 1984).
Elliott et al. (1999) studied the quantitative metabolism of single gavage doses of
[ C]cyclohexane (0.3 to 400 mg/kg) administered to adult chinchilla doe rabbits. The
experiment apparatus allowed for the capture of expired air and urine from the rabbits. Recovery
of the radioactivity was about 95%. Only [14C]-labeled carbon dioxide and unchanged
cyclohexane were detected in expired air. The concentration of unchanged cyclohexane in
expired air increased with the dose. In urine, the metabolites detected were primarily
cyclohexanol with lesser frvms-l^-cyclohexane-diol formation. Cyclohexanol and the diol were
both excreted as glucuronides.
The authors did not explain the presence of [ C]carbon dioxide, because the
dicarboxylic (adipic, succinic, maleic, malonic, or oxalic) acids produced in the ring schism were
not detected. Carbon dioxide has not been detected in other more recent studies of cyclohexane
metabolism in mammals. Carbon dioxide is a well-established metabolite of w-hexane
(Battershill et al., 1987), but w-hexane contamination was not reported in the current study.
Oral administration of cyclohexanol resulted in the same two metabolites as did
cyclohexane: cyclohexanol and the frvms-l^-cyclohexane-diol. Further, when cyclohexanone
and cyclohex-1-enyl acetate were administered, both were converted to cyclohexanol. All
cyclohexane derivatives were conjugated to glucuronides (Elliott et al., 1959). Perbellini and
Brugnone (1980) observed cyclohexanol and cyclohexanone in rat urine following cyclohexane
exposure.
A definitive study of the absorption, distribution, metabolism, and excretion of
[ C]cyclohexane following a single intravenous dose (10 mg/kg) or single oral doses (100, 200,
1,000, or 2,000 mg/kg) to adult male Fischer 344 rats was carried out under the auspices of the
National Institute of Environmental Health Sciences (RTI, 1984). After an oral dose of 200
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mg/kg [ C]cyclohexane, five unidentified metabolites in blood accounted for more than half of
the radioactivity, regardless of sampling time. Another radioactive metabolite in blood was
cyclohexanol, present at levels two to three times higher than cyclohexanone. The highest blood
concentrations of cyclohexanol and cyclohexanone occurred 2 hours after dosing. During the
first 12 hours, cyclohexanone levels accounted for 10% of the blood C but decreased to 1-3%
within 24 hours. Cyclohexane was also detected as a minor blood constituent.
3.4. EXCRETION
Inhaled cyclohexane is excreted primarily via expiration from the lungs. A small portion
partitions to and is excreted in the urine. Metabolites of cyclohexane are conjugated, primarily
to glucuronides and possibly to sulfates, and excreted in the urine.
A number of occupational health and monitoring studies of leather (shoe and luggage)
factory workers exposed to cyclohexane provide evidence of alveolar and urinary excretion.
Perbellini and Brugnone (1980) detected evidence of cyclohexane expiration in some shoe
factory workers, but did not identify it as such (see Section 3.1.1.).
Mutti et al. (1981) calculated that approximately 10% of an absorbed dose of
cyclohexane was expired after exposure, while most of the dose was retained. Expiration was
initially rapid (11.2 minute half-life) for the first hour, followed by a slower component (115.3
minute half-life) thereafter. The total lung uptake during a 4-hour exposure period and the total
alveolar expiration during a 6-hour postexposure period for eight factory workers was calculated
as 81.2 mg (mean) and 9 mg (mean) cyclohexane, respectively. The authors suggested that
alveolar excretion of cyclohexane in most occupational settings is low and fluctuates rapidly in
response to environmental concentration changes. At higher inhalation doses, some
cyclohexane is expired unchanged due to preferential partitioning to the alveoli rather than to
blood, where it has low solubility (see Section 3.1.1.).
Ghittori et al. (1987) proposed that, according to Henry's Law, a small portion of
cyclohexane must be excreted in the urine in its unchanged form. Unreacted cyclohexane in the
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blood should reach a steadystate equilibrium with alveolar air and glomerular filtrate. Measuring
cyclohexane in the urine of 43 human subjects, the authors calculated the urine-to-air partition
coefficient for cyclohexane (0.9) and found a relationship between environmental cyclohexane
levels and urinary cyclohexane levels using the following regression equation:
y = ax+b.
where: y = urinary cyclohexane concentration (nmol/mL)
x = time-weighted breathing zone cyclohexane concentration (ppm)
The result with correlation coefficient is:
y = 0.05 x +8.26 (r = 0.89)
Perbellini et al. (1980) detected urinary cyclohexanol at a mean level of 1.4 ± 1.6 mg/L in
shoe factory workers exposed to a leather adhesive containing nine solvents, including
cyclohexane (456 ± 464 mg/m ). Yuasa et al. (1996) found that the urinary concentration of
cyclohexanol in 18 female workers at a luggage factory that used a cyclohexane-based (76%)
solvent ranged from 0.12 to 1.51 mg/L over an 18-month period. Breathing space cyclohexane
levels ranged from 5 to 211 ppm. There was a strong correlation between the cyclohexane
exposure in personal air and urinary cyclohexanol.
Mutti et al. (1981) detected cyclohexanol and cyclohexanone in workers' urine,
accounting for no more than 0.5 to 1% of the calculated absorbed dose. The level of excretion
was poorly related to the level of exposure and showed wide scatter at higher occupational
concentrations. Similarly, Perbellini and Brugnone (1980) determined that urinary cyclohexanol
levels accounted for only 0.1 to 0.2% of the absorbed cyclohexane. In 22 workers, the mean
level of urinary cyclohexanol was 2.24 \igfL (range: 0.27-7.18), and the mean excretion rate was
0.92 |ag/minute (range: 0.05-3.23).
Two studies of shoe factory workers who were continually exposed to leather adhesive
and cleaning solvent vapors containing cyclohexane suggest that urinary excretion of
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cyclohexanol is rapid and declines to low levels by the morning following the exposure event.
Governa et al. (1987) detected cyclohexanol (0.23 ± 0.68 mg/L; range: 0.1-3.80 mg/L) in about
20% of urine samples collected at 9 a.m. from workers (n = 40). Perbellini et al. (1987) analyzed
urine samples collected from three workers before the start and at the end of the work shift for an
entire workweek. All morning samples were < 0.4 mg/L cyclohexanol, whereas levels in
afternoon samples (approximately 1-4 mg/L) increased proportionally to the mean occupational
exposure levels.
Perbellini and Brugnone (1980) determined that urinary cyclohexanol levels correlated
with environmental and alveolar cyclohexane concentrations and that urinary cyclohexanol
excretion rates correlated with alveolar cyclohexane. The correlation between blood
cyclohexane levels and urinary cyclohexanol levels and excretion rate were weaker.
Measurements were taken during the last 4 hours of an 8-hour work shift.
Occupational health monitoring by Yasugi et al. (1994) of 33 female workers exposed to
cyclohexane vapors evaluated urinary metabolites of cyclohexane. Urine was collected on the
fourth or fifth day of the workweek at shift's end and again the next morning. Analysis showed
the presence of cyclohexanol and cyclohexanone. The mean level of cyclohexanol in urine was
875.7 ± 2.86 [ig/g. More than 90% was conjugated as glucuronide, the remainder was unbound.
Some unconjugated cyclohexanone was also detected. No sulfate conjugates were detected.
Both urinary cyclohexanol and cyclohexanone concentrations correlated with exposure levels.
Quantitative estimation indicated that only 1% of the absorbed cyclohexane was excreted in the
urine as cyclohexanol by the end of the work shift. Cyclohexanol was still present in urine 16
hours after exposure (95.8 ± 2.86 |-ig/g). A rough biological half-life of 5 hours was calculated
assuming a single compartment, suggesting that clearance from the body is relatively rapid.
Treon et al. (1943a, b) demonstrated that oral gavage of young white rabbits at one-half
the lethal cyclohexane dose (2.88 g/kg) was followed by an increase in organic conjugated
compounds in urine that lasted 48 hours. Other rabbits, repeatedly exposed to cyclohexane
vapors (6-8 hours/day, 5 days/week for 2-26 weeks) at levels ranging from 434 to 18,565 ppm
(1.46 to 62.5 mg/L), exhibited an increase in glucuronic acid and conjugated organic sulfate in
the urine. At an exposure level of 434 ppm for 26 weeks, the excretion of glucuronic acid was
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elevated (225 mg/day), but it returned to normal (59 mg/day) when dosing was discontinued.
Approximately 50% of the cyclohexane dose was recovered as glucuronides. In the lowest-dose
group (434 ppm), excretion of organic sulfate was similar to controls but increased at higher
doses until a maximum response was reached. Sulfate excretion returned to normal levels on
termination of exposure.
When the dose of cyclohexane administered to rabbits was an order of magnitude lower,
there was no evidence of sulfate conjugation. The radioactivity from an oral gavage dose of
[ C]cyclohexane was almost completely recovered in urine and expired air within 2 days after
dosing. At low doses (0.3 mg/kg), around 85% of the radioactivity was recovered in urine as
glucuronide conjugates of cyclohexanol and frvms'-l,2-cyclohexanol. At high doses, the amount
in urine decreased to about 50% since the entire cyclohexane dose was not metabolized. Up to
38% was expired unchanged in air. While most of the metabolized cyclohexane (80-90%) was
excreted in the urine, radioactivity in expired carbon dioxide increased to about 15% as the dose
increased. At all doses, a small amount of radioactivity (< 5%) was detected in the feces and
tissues (Elliott et al., 1959).
Studies in mice (Naruse, 1984) and rats confirm that when exposure to cyclohexane is
discontinued, cyclohexane levels in the body drop rapidly. In rats, cyclohexane levels in
perirenal fat are high, but no cyclohexane was detected in other tissues 2 weeks after a 14-day
exposure period (Savolainen and Pfaffli, 1980). In rats dosed for 3 days, perirenal fat levels
were reduced by two-thirds only 12 hours after exposure ceased. Other tissues contained 6% or
less of their respective exposure values (Zahlsen et al., 1992).
In the RTI (1984) study (see Sections 3.2. and 3.3.), a comparison of the intravenous and
oral route of dosing in adult male Fischer 344 rats found that after intravenous dosing with 10
mg/kg [14C]cyclohexane, approximately 80% of the radioactivity was expired the first day - 54%
within the first hour. Slightly more than 1% of the dose was expired on the second and third
days. Urinary metabolites constituted 14% of the dose; more than 80% of the metabolites were
excreted during the first day. The ratio of the radioactivity expired to the radioactivity excreted
in urine was approximately 6.2 to 1. After oral dosing, the amount expired was proportional to
the dose. Twelve to 29% of the radioactivity was excreted in the urine, where levels were
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inversely proportional to the dose. With doses of 100 or 200 mg/kg [ C]cyclohexane, the ratio
of the radioactivity expired to that excreted in urine was 2 to 1, and at doses of 1,000 or 2,000
mg/kg the ratio was 5 to 1 and 6.5 to 1, respectively.
Independent of dose vector, cyclohexane accounted for 93-99% of the expired C. Less
than 1% was expired as cyclohexanol and cyclohexanone. In the urine, cyclohexane,
cyclohexanol, and cyclohexanone separately represented less than 0.1% of the excreted 14C. The
authors suggest that the remainder of the urinary C was conjugated. The level of Cinfeces
was not significant. No significant amounts of [ C]carbon dioxide were detected, and there did
not appear to be any substantial retention of radioactivity in any tissue. The body half-life for
14C was estimated at 10 to 15 hours (RTI, 1984).
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4. HAZARD IDENTIFICATION
4.1. STUDIES IN HUMANS—EPIDEMIOLOGY, CASE REPORTS, CLINICAL
CONTROLS
There are very few epidemiologic studies on cyclohexane alone. Most of the
occupational studies reported in the literature indicate that cyclohexane is only one of several
solvents used in the workplace. In most of these studies, it is suspected that other solvents such
as w-hexane or toluene comprise the majority of the workplace exposures and are responsible for
adverse health effects, including neurotoxicity and spontaneous abortion (Agnesi et al., 1997;
Lee et al., 1998; Mutti et al., 1981). Notwithstanding, two small studies (Yasugi et al., 1994;
Yuasa et al., 1996) have been conducted with workers who were exposed primarily to
cyclohexane. Although very few adverse health effects associated with cyclohexane exposure
were reported, definitive conclusions cannot be drawn due to certain limitations within both
studies.
Thirty-three women were exposed to cyclohexane in a Japanese factory where glue is
applied to surfaces by automated sprayers (Yasugi et al., 1994). The two glues used in the
factory contained at least 75% cyclohexane. The workers included in the study were exposed to
the glue solvents for at least 1 year. They provided a urine sample after 5-6 hours of exposure
and blood and urine samples after the 8-hour work shift. They were also administered a
questionnaire on subjective symptoms experienced within the last 3 months both at home and at
work. Personal monitors indicated that 274 ppm (27 ppm geometric mean) was the highest level
of cyclohexane measured.
The exposed subjects were divided into low (< 5-13 ppm) and high (15-274 ppm)
exposure groups with 17 and 16 workers, respectively. They were compared to 10 controls
using chi-squared tests. Hematology and serum biochemistry parameters were analyzed. At the
p < 0.05 significance level, no differences were found in the hematology or serum biochemistry
parameters in liver and kidney function. In the questionnaire, there were no significant
differences (p > 0.05) in the subjective symptoms experienced at work either individually or in
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combination. There was also no difference in sister chromatid exchange rate between exposed
and nonexposed workers.
Given the small number of women in the survey, the even smaller number of control
subjects, and the relatively low exposure levels, it is not surprising that effects were not reported
in this study. In addition, the report does not contain any definitions of the subjective symptoms
or the results of these tests or any specific results of the hematology and biochemistry tests. The
results are therefore difficult to interpret.
In another Japanese study, neurophysiological effects were analyzed in female luggage
factory workers exposed to glue containing 75% cyclohexane, 12% toluene, and 0.9% w-hexane
(Yuasa et al., 1996). Prior to the start of the study, w-hexane was the primary solvent used at the
plant, but it was gradually discontinued and replaced with cyclohexane in 1992. Therefore,
several of the participants had past exposures to w-hexane; 12 of 18 women in the first study year
(at a length of n-hexane exposure time of 0.3-20 years) and 8 of 9 in the second study year. The
workers were exposed for approximately 8 hours/day. In the first study year (April through July
1993), the women had been employed for 0.4-2.6 years; the shortest time between exposure of
any of the workers to w-hexane and the first study was 0.7 years. Due to changes in job
assignment, only 9 women in the first study year participated in the second observation period
(July 1994). The 18 control subjects consisted of medical students and clerical workers, and
were significantly different from the study subjects.
Air monitoring indicated that cyclohexane levels ranged from 5 to 211 ppm (geometric
mean, 28 ppm; median, 46 ppm), although it is unclear whether these levels were observed in the
first or second study year. Concentrations prior to the first study period were much higher
(6-720 ppm; geometric mean, 77 ppm). Based on the exposure data, high (> 100 ppm) and low
(< 100 ppm) exposure groups were formed for the first study period.
In the first study year, workers participated in biological monitoring (urinary
cyclohexanol), a neurophysiological study (nerve conduction velocity), and a subjective
symptom survey (fatigue, headache, etc.). Urinary cyclohexanol measurements ranged from
0.12 to 8.23 mg/L and were highly correlated to ambient cyclohexane levels in the wokplace.
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The symptom survey did not indicate any significant difference in subjective symptoms between
exposed workers and controls except for general fatigue (p < 0.1). The symptom survey was not
administered in the second year.
In the neurophysiological examination, the nerve conduction velocity measurements of
the exposed workers when compared to controls were not significantly different (p < 0.05), nor
were they significant when high- (n = 7) and low- (n = 11) exposure groups were compared to
controls. In the second year, only 9 of the original 18 exposed workers were included in the
study. A significant improvement in several of the nerve conduction velocity tests was noted in
these employees between the first and second years. Therefore, the authors concluded that n-
hexane affected the measurements in the first study and that the workers recovered by the second
period.
As previously mentioned, there are many limitations to this study and the results should
be interpreted carefully. The small number of participants, especially in the second year, limits
the reliability of the data. Also, the controls were not chosen from unexposed workers at the
same plant and were not matched well to the subjects. Past exposures to other solvents and
length of past exposure was not taken into consideration, even though it varied widely among
participating workers.
4.2. PRECHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN
ANIMALS—ORAL AND INHALATION
No adequate oral prechronic studies and no chronic studies of any exposure route were
located for cyclohexane. However, two unpublished, 90-day inhalation toxicity studies were
conducted with cyclohexane in mice and rats (DuPont HLR, 1996a, b). These studies were later
summarized and published as parts of Malley et al. (2000).
Exposure concentrations for the 90-day study of mice were selected based on the results
of a 2-week range-finding study and knowledge of the explosive properties of cyclohexane. As
summarized in DuPont HLR (1996a), the range-finding study (concentrations of 3,000, 6,000,
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and 9,000 ppm [10,329, 20,658, and 30,987 mg/m3]) showed that the two higher concentrations
decreased the response to an alerting stimulus during exposure. The response to an auditory
stimulus was evaluated prior to the initiation of exposure to cyclohexane, during exposure to
cyclohexane and during the time required to clear the chambers. Groups of animals were
observed for normal, diminished, or hyperresponsive altering behavior in response to an auditory
stimulus. In addition, mice exposed to 30,987 mg/m3 displayed sporadic incidences of jumping
and/or slow circling behavior. It was reported that male mice had compound-related increases in
relative lung weights in the 20,658 and 30,987 mg/m3 groups. Female mice in the same groups
had significantly increased absolute and relative liver weights (DuPont HLR, 1996a).
In DuPont HLR (1996a), male and female Crl:CD-l BR mice were exposed by whole-
body inhalation to cyclohexane vapor at concentrations of 0, 500, 2,000, or 7,000 ppm (0, 1,721,
6,886, or 24,101 mg/m3) for 6 hours/day, 5 days/week for 90 days. Initially, there were
20/sex/concentration for the control and high-concentration groups and 10/sex/concentration for
the low- and intermediate-concentration groups. Ten mice per sex from the control and high-
concentration groups were allowed a 1-month recovery period prior to sacrifice.
There were no treatment-related deaths. A few mice in each group died due to blood
sampling errors. There were no treatment-related effects on body weight, body weight gain, or
food consumption. Clinical signs and response to an alerting stimulus during exposure exhibited
a dose-response relationship. Inhalation exposure of mice to 24,101 mg/m3 produced clinical
signs of hyperactivity and marked central nervous system stimulation, which persisted for a short
period after the end of each daily exposure. The clinical signs of toxicity included:
hyperactivity, circling, jumping/hopping, excessive grooming, kicking of rear legs, standing on
hind legs, and occasional flipping behavior. These signs were evident by the fourth exposure
and persisted throughout the remaining exposures. The clinical observations of response to an
alerting stimulus varied as it was diminished in some instances and it could not be assessed due
to hyperactivity at other periods. During the recovery period, no central nervous system
stimulation was observed.
The relative liver weights increased in male and female mice in the 24,101 mg/m3 group,
mean relative liver weights (percent of body weight [standard deviation], males: 4.822 [0.313]
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vs. controls 4.151 [0.411]; females: 4.726 [0.307] vs. controls 4.272 [0.388]). At the end of the
recovery period, no significant changes in liver weights were observed. The erythrocyte mass
was increased, but the cause of this change is unknown although it may have been due to the
bleeding schedule. Mice in the 6,886 mg/m3 group showed hyperactivity late in the exposure
period and sedative effects that were apparent through most of the exposure period.
Exposure concentrations were selected for a similarly designed 90-day inhalation toxicity
study of cyclohexane in male and female Sprague-Dawley rats based on the results of a 2-week
range-finding study and knowledge of the explosive properties of cyclohexane. In this range-
finding study, as reported in DuPont HLR (1996b), concentrations of 3,000, 6,000, and 9,000
ppm (10,329, 20,658, and 30,987 mg/m3) were tested. The two higher concentrations decreased
the response to an alerting stimulus during exposure. The body weights of the rats exposed to
30,987 mg/m3 were lower than the controls and there was a higher incidence of mitotic figures in
hepatocytes from the high-concentration group and 20,658 mg/m3 males (DuPont HLR, 1996b).
In the 90-day inhalation toxicity study ((DuPont HLR, 1996b; Malley et al., 2000) in
male and female Sprague-Dawley rats (20/sex/concentration for control and high-concentration
groups and 10/sex/concentration for low- and intermediate-concentration groups) were exposed
by whole-body inhalation to cyclohexane vapor at concentrations of 0, 500, 2,000, or 7,000 ppm
(0, 1,721, 6,886, or 24,101 mg/m3) for 6 hours/day, 5 days/week for 90 days. At the end of 90
days, 10 rats/sex/concentration were sacrificed. After a 1-month recovery period, 10 remaining
rats/sex from the control and high-dose groups were sacrificed and their tissues were examined
for histologic changes.
There were no treatment-related deaths and there were no significant differences in body
weight among the control and the treatment groups. The most common clinical observation was
diminished alerting responses in the chamber during exposure at 6,886 and 24,101 mg/m3. This
effect was characterized as transient and was not observed immediately after removing the
animals from the chamber. Rats in the 1,721 mg/m3 group did not show this central nervous
system effect.
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The only other treatment-related effect was increased relative liver weights in high-
concentration males at 90 days and after the 1-month recovery period. Both males and females
in the 24,101 mg/m3 group had hepatocellular hypertrophy, which was considered the cause of
the enlarged livers. Decreases in the activity of some serum enzymes related to hepatic function
were statistically significant. Significant decreases in aspartate aminotransferase, sorbitol
dehydrogenase, lactate dehydrogenase and creatine phosphokinase were detected, but the mean
values generally did not exhibit a dose-response relationship. Although increases in such
enzyme activities can indicate tissue damage, decreases in these enzyme activities generally are
not considered biologically significant effects.
The mean relative liver weights in high-concentration males were significantly higher
than those of controls at the 90-day and 1-month recovery terminations (percent of body weight
[standard deviation], day 95: 4.001 [0.265] vs. control 3.649 [0.214]; day 123: 4.009 [0.313] vs.
control 3.767 [0.240]). Gross observations showed large livers in 10/10 and 4/10 males at
24,101 mg/m3 at 90 days and after 1-month recovery period, respectively. Centrilobular
hepatocellular hypertrophy was observed microscopically in 9/10 males and 5/10 females in the
24,101 mg/m3 group at the 90-day sacrifice but not after the 1-month recovery period. The study
authors considered the hepatic enlargement to be an adaptive response and not an adverse effect.
On the other hand, in the absence of long-term exposure data, the hepatic enlargement and the
incomplete reversibility of the effect during the recovery period may indicate that there could be
a progression of liver effects to frank toxicity with longer exposure to cyclohexane.
In summary, 90-day inhalation toxicity studies (Malley et al., 2000) were conducted with
cyclohexane in mice (DuPont HLR, 1996a) and rats (DuPont HLR, 1996b). In mice,
hyperactivity and diminished response to an alerting stimulus was observed at 6,886 and 24,101
mg/m3. In rats, diminished response to an alerting stimulus was also observed at 6,886 and
24,101 mg/m3. Relative liver weights were increased in both rats and mice treated with 24,101
mg/m3 cyclohexane. In the absence of pathological changes in the liver, it cannot be determined
whether these changes are the first signs of a potential liver toxicity that would only become
apparent with long-term exposure to cyclohexane.
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4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND INHALATION
No adequate studies of reproductive or developmental toxicity of oral exposure to
cyclohexane were located. Unpublished reports of two-generation reproduction toxicity in rats
and prenatal developmental toxicity in rats and rabbits exposed to cyclohexane by inhalation
were submitted by industry (DuPont HLR, 1997a, b, c) and later summarized and published
(Kreckmann et al., 2000).
A two-generation reproduction inhalation toxicity study of rats was conducted with
cyclohexane involving the production of one set of litters in each generation (DuPont HLR,
1997a; Kreckmann et al., 2000). Male and female Crl:CD BR rats (Sprague-Dawley strain,
30/sex/concentration) were exposed by whole body inhalation to cyclohexane vapor at
concentrations of 0, 500, 2,000 or 7,000 ppm (0, 1,721, 6,886, or 24,101 mg/m3). Following 10
weeks of exposure (generally 6 hours/day for 5 days/wk, excluding holidays), the animals were
bred within their respective treatment groups and allowed to deliver and rear their offspring until
weaning. Pregnant females were exposed daily, 6 hours/day, during days 0-20 of gestation. As
specified in the protocol, they were not exposed from day 21 of gestation until day 4 of lactation.
At day 5 of lactation, daily exposure of dams was resumed. Neonates were not directly exposed
to cyclohexane. At weaning, Fl rats were randomly selected to produce the next generation and
exposed to cyclohexane as described above. At least 11 weeks after weaning, the Fl rats were
bred to produce the F2 litters.
It was reported that the high concentration was based on a pilot developmental toxicity
study demonstrating that maternal body weight and food consumption were reduced at 6,000
ppm (20,658 mg/m3) and above. The explosive hazard of cyclohexane under pressure limited the
high concentration tested to 7,000 ppm (24,101 mg/m3), corresponding to approximately 60% of
the lower explosive limit. The remaining concentrations were selected by equal spacing on a log
scale. The purity of cyclohexane was reported to be greater than 99.9%.
Clinical observations during exposure showed a diminished response or no response to a
sound stimulus beginning at exposure 15 in animals exposed to 6,886 or 24,101 mg/m3.
Specifically, females showed a diminished or absent altering response during exposure to 8,886
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or 24,101 mg/m3 cyclohexane. Because the premating animals were exposed 5 days/week, this
would be approximately day 19 of the exposure period. The sedation was characterized as
transient and was no longer apparent shortly after the rats were removed from the chamber. The
animals in these two groups also showed salivation, stained perioral area, and wet chin. These
clinical signs may be related to the sedation.
For the PI generation, there were no compound-related reductions in body weight or food
consumption for the males. Mean body weight for the PI females in the 24,101 mg/m3 group
was significantly reduced by day 64 of the premating period (94% of the control). The authors
stated that although the mean body weight was significantly reduced as compared with controls
for the gestation and lactation periods, body weight gain was similar to that of controls during
these periods, indicating that the differences were probably due to the preexisting weight deficits
established during the premating period. For the Fl generation, body weight was significantly
reduced throughout the study for Fl males in the 24,101 mg/m3 group and throughout the
premating, gestation, and lactation periods for Fl females in the 24,101 mg/m3 group (92% of
control at the end of premating). The authors stated that, as with the PI generation females, the
lower body weights during gestation and lactation of the Fl females were probably a
continuation of the body weight deficits established during the premating period as reflected in
generally similar body weight gains between treated and control females in the later two periods.
There were no significant differences in mating, fertility or gestation indices,
implantation efficiency, or gestation length in either the PI or the Fl generation. There were no
dose-related trends in the mean number of implantation sites, mean number of pups/litter, or any
survival indices for both Fl and F2 litters (sex ratio, percent born alive, 0-4 day viability,
lactation index, and litter survival). The mean pup weight was significantly reduced from
postpartum day 7 through the remainder of the 25-day lactation period for Fl and F2 litters (see
Tables 5-1 and 5-2 in Section 5.2.1.). The method of statistical analysis in the report was an
analysis of covariance with litter size and sex ratio as covariates, followed when significant with
linear contrast of the least square means. No compound-related effects on organ weights, gross
observations, or microscopic findings were found.
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In summary, the reproductive toxicity study concluded that the inhalation exposure of
rats to 24,101 mg/m3 cyclohexane vapor produced significant reductions in body weights in PI
and Fl females and Fl males, significant reduction in pup weights from lactation days 7 to 25
for Fl and F2 litters, and the clinical observation of diminished or absent response to a sound
stimulus while in the exposure chamber. At the 6,886 mg/m3 level, the only cyclohexane effect
was the subjective clinical observation that rats as a group had a diminished response to a sound
stimulus while in the exposure chamber. Based on the reduced rat pup weights during lactation
in the two generations tested, the NOAEL for developmental effects in this reproductive toxicity
study was 6,886 mg/m3.
Inhalation developmental toxicity studies of cyclohexane (Kreckmann et al., 2000) were
conducted using rats (DuPont HLR, 1997b) and rabbits (DuPont HLR, 1997c).
Pregnant rats (Crl:CD:BR strain, 25/concentration) were exposed by whole-body
inhalation to test material vapor at concentrations of 0, 500, 2,000, or 7,000 ppm (0, 1,721,
6,886, or 24,101 mg/m3) for 6 hours/day on days 6-15 of gestation. In addition to the standard
control group, a pair-fed control group was included; this group received an amount of food
equal to the cumulative average amount of food consumed by the high-concentration group on
the corresponding gestation day. Maternal body weights were recorded on days 0, 6-15, and 21;
food consumption was recorded daily, and clinical signs were recorded daily on days 0-6 and
16-21 and twice daily on days 6-15. Dams were sacrificed on day 21, and the fetuses were
weighed, sexed, and examined for external and skeletal abnormalities; one-half of the fetuses
were examined for visceral and head abnormalities (DuPont HLR, 1997b; Kreckmann et al.,
2000).
Maternal toxicity was evident in the mid- and high-concentration groups. Mean maternal
body weight gain was significantly reduced during the treatment period in both groups
(approximately 69% of control for the 24,101 mg/m3 group), and food consumption was
significantly reduced during the treatment period for the 24,101 mg/m3 group. Mean body
weight gain was also significantly reduced in the pair-fed control group. At 24,101 mg/m3 there
was a significant increase in the clinical sign "stain chin." While the incidence of salivation was
also increased, the difference was not statistically significant. The source of the stained fur was
26
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presumed to be salivation. These signs lasted 10-15 minutes after exposure but were not
observed prior to or during exposure. Animals in the 6,886 and 24,101 mg/m3 groups exhibited
a diminished response or no response to a sound stimulus while in the chambers during
exposure, indicating a transient sedative effect. Necropsy revealed no gross lesions. On the
basis of these results, the LOAEL for maternal toxicity in rats was 6,886 mg/m3, and the NOAEL
was 1,721 mg/m3.
No evidence of statistically significant developmental toxicity was presented for rats at
any dose level. There were no significant differences between control and treatment groups in
the number of resorptions, the number of live fetuses, average fetal weight, or sex ratios of pups.
There were also no statistically significant differences in the number of external, skeletal,
visceral, or head abnormalities in pups. The total incidence of fetal malformations was four
fetuses from four litters in the 24,101 mg/m3 group, one fetus in one litter in the 6,886 mg/m3
group, none in the 1,721 mg/m3 group, two fetuses in two litters from pair-fed controls, and none
in the ad libitum-fed control group. In general, finding one defective fetus in all four litters is of
greater concern than an observation of four such fetuses in one litter (Hood, 1996). However, the
malformations in the four high-dose fetuses were of different types, malformations were
observed in fetuses from two litters from the pair-fed controls, and no other signs of
developmental toxicity were noted. The authors of the study did not comment on these data but
stated that statistical analyses were only conducted on individual endpoints, which would miss
this increase in the total incidence of fetal malformations. They concluded, "No compound-
related effect on the incidence of fetal malformations was observed" (DuPont HLR, 1997b). The
study authors stated that the NOAEL for developmental toxicity was the highest dose tested,
24,101 mg/m3.
In a study designed similarly to the rat study, pregnant New Zealand white rabbits
(20/concentration) were exposed by whole-body inhalation to cyclohexane vapor at
concentrations of 0, 500, 2,000, or 7,000 ppm (0, 1,721, 6,886, or 24,101 mg/m3) for 6 hours/day
on gestation days 6-18 (DuPont HLR, 1997c; Kreckmann et al., 2000).
In the control, 1,721, and 24,101 mg/m3 groups, all 20 females had litters. There were no
treatment-related deaths. Measurements of body weight and weight gain, food consumption,
27
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clinical observations, and postmortem findings did not show any treatment-related effects. In the
6,886 mg/m3 group, 17 females were pregnant and one aborted the litter before the end of
gestation. No statistically significant differences were found between control and treatment
groups in pregnancy rate, abortion rate, total resorption rate, the mean number of implantations
per litter, or the mean number of live fetuses per litter.
There was a statistically significant decrease in the mean number of corpora lutea for
females in the 6,886 and 24,101 mg/m3 groups, which had a mean number of 8.9 corpora lutea in
17 litters and 8.8 in 20 litters, respectively. While the 1,721 mg/m3 group had a mean number of
8.9 corpora lutea in 20 litters, it was not significantly different from the control value (10.2 in 20
litters). However, the study authors did not judge the decrease in the mean number of corpora
lutea to be biologically significant. First, the decreased means were within the range of control
data (historical control data in the report, Appendix M), which showed that 17 studies conducted
between 1990 and 1995 had a mean value of 8.9 +1.1; the maximum was 10.9 and the minimum
was 7.0. Secondly, the number of corpora lutea for the concurrent control group was near the
high end of that range. Most importantly, because ovulations and implantation occurred prior to
exposure to the test substance, the decrease in the mean number of corpora lutea for treated
groups was not considered to be compound related.
There was a significant trend in sex ratio (number of males/total number of pups), with
the ratios being higher for the 6,886 and 24,101 mg/m3 groups (0.59 and 0.54, respectively,
compared with the control ratio of 0.48). However, the authors concluded that this trend
appeared to be of unknown significance because of the disparity between the ratios for the 1,721
mg/m3 (0.42) and the 6,886 mg/m3 (0.59) groups. No true dose-response for the differences in
the sex ratios was found, and the values generally fell within the historical control values
(0.40-0.56).
Among the fetuses, there were no significant differences between control and treatment
groups in early, late, or total resorptions, fetal weight, malformations, or variations. No
significant differences were observed between control and treatment groups in any measures of
developmental toxicity. Therefore, the NOAEL for maternal and fetal effects in rabbits was
24,101 mg/m3, the highest concentration tested.
28
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In summary, inhalation developmental toxicity studies of cyclohexane were conducted
using rats (DuPont HLR, 1997b) and rabbits (DuPont HLR, 1997c) and later summarized and
published as parts of Kreckmann et al., 2000. Inhalation exposure to concentrations of 6,886 and
24,101 mg/m3 cyclohexane resulted in maternal toxicity in CD rats, as demonstrated by a
significant reduction in body weight gain. There was no evidence of developmental toxicity in
rat pups at the highest concentration tested, 24,101 mg/m3. Therefore, the LOAEL for maternal
toxicity in rats was 6,886 mg/m3, and the NOAEL was 1,721 mg/m3. The NOAEL for
developmental toxicity in rat pups was 24,101 mg/m3 (DuPont HLR, 1997b; Kreckmann et al.,
2000). Inhalation exposure to the highest concentration of 24,101 mg/m3 cyclohexane resulted
in no evidence of maternal or developmental toxicity in rabbits (DuPont HLR, 1997c;
Kreckmann et al., 2000).
4.4. OTHER STUDIES
4.4.1. Acute Toxicity
Acute oral lethality of cyclohexane was apparently affected by age in rats; LD50 values
were 8, 39, and 16.5 mL/kg in 14-day-old, young adult, and older adult rats, respectively
(Kimura et al., 1971). The minimum lethal dose after a single gavage exposure in white rabbits
(strain not specified) was between 5,500 and 6,000 mg/kg; body weight loss, diarrhea, increased
respiratory rate, conjunctival congestion, and lethargy were observed in rabbits administered
cyclohexane at or above 1,000 mg/kg (specific dose levels not reported) (Treon et al., 1943a). In
mice, an oral LD50 of 813 mg/kg has been reported (NIOSH, 2000). No mortality, change in
body weight gain, or gross pathological changes were observed in a group of five male and five
female rats (unspecified strain) up to 2 weeks after a single gavage exposure to 5,000 mg/kg
cyclohexane, but clinical signs suggestive of central nervous system involvement included
transient depression, salivation, and soft feces (HLA, 1982a).
Similarly, no mortality, change in body weight gain, or gross pathological changes were
observed in a group of five male and five female rats (unspecified strain) up to 2 weeks after a
single inhalation exposure to 21,250 ppm cyclohexane for 4 hours, but clinical signs suggestive
29
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of central nervous system involvement included tremors, hyperactivity, rapid respiration, ataxia,
and prostration (HLA, 1982b). Cyclohexane did not produce significant upper airway irritation
in mice after a single inhalation exposure to 21,750 ppm, as indicated by no exposure-related
change in respiration rate (HLA, 1982c). Trembling, "disturbed" equilibrium, or complete
recumbency were reported in mice, guinea pigs, rabbits, and cats exposed by inhalation to
18,000 ppm cyclohexane for 30 or fewer minutes (Flury and Zernik, 1931).
4.4.2. Neurological Studies
An acute operant behavior study of cyclohexane by inhalation in rats and a 90-day
inhalation neurotoxicity study of cyclohexane in the rat were conducted by DuPont HLR (1996c,
d), Christoph et al. (2000), and Malley et al. (2000). Similar to observations in previously
discussed studies, there was no evidence of neurotoxicity or impaired response caused by
inhalation exposure to cyclohexane beyond the subjective clinical observation of a diminished
response to an alerting stimulus at the time of exposure in rats in the current study.
An acute operant behavior study of cyclohexane by inhalation in rats examined the
effects of 6-hour exposures on schedule-controlled operant behavior in Crl:CD:BR rats (DuPont
HLR, 1996c; Christoph et al., 2000). Rats were exposed to 0, 500, 2,000, or 7,000 ppm (0,
1,721, 6,886, or 24,101 mg/m3) in a chamber. Behavioral toxicity was detected at the high dose
of 24,101 mg/m3 only. No effects were seen in control animals or at exposure concentrations of
1,721 or 6,886 mg/m3. At 24,101 mg/m3, a transient decrease in the mean fixed-ratio rate of
responding was apparent from the analysis of variances over the day before exposure, the day of
exposure, and the day after exposure. None of the other recorded parameters showed statistically
significant effects. Fixed-ratio pause duration showed a slight, nonsignificant increase after
exposure, which is difficult to interpret since prior to exposure the mean fixed-ratio pause
duration for the two high-dose groups was considerably lower than for the other groups.
Although no effect was detected in the study, subtle effects may have been missed by the
protocol. Positive control data for amphetamine sulfate and chlorpromazine on schedule
controlled operant behavior were submitted with the study. Although this is a well-conducted
study, these data do not demonstrate evidence that the experimental paradigm could detect both
30
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increases and decreases in response rates in all four of the relevant parameters, demonstrating the
low sensitivity of this type of acute testing in practice.
In the 90-day inhalation neurotoxicity study in rats (DuPont HLR, 1996d; Malley et al.,
2000), the neurotoxicity screening battery included a functional observational battery, motor
activity, and neuropathology. Twelve male and female Crl:CD BR rats per group were exposed
to concentrations of 0, 500, 2,000, or 7,000 ppm (0, 1,721, 6,886, or 24,101 mg/m3) of
cyclohexane for 6 hours/day, 5 days/week for approximately 90 days. All rats were evaluated
using motor activity and functional observational battery assessments prior to exposure to
establish baseline measurements. The tests were conducted again on weeks 4, 8, and 13;
however, the precise timing of the tests relative to exposure was not stated. Observations of
responsiveness to an alerting stimulus were made during exposure. Rats were also observed for
postexposure clinical signs. Following testing, six rats/gender/group were sacrificed, perfused,
and examined grossly. Histopathology was conducted on the control and high-dose groups.
Positive control data indicate that the equipment and procedures were capable of
detecting effects that may be seen in this type of neurotoxicity study. No compound-related
effects were found on food consumption, body weight, or body weight gain at any exposure
concentration. Clinical observations included a dose-related effect on alerting response noted
while rats were in the exposure chamber. Rats in the group exposed to 6,886 mg/m3 exhibited a
normal alerting response during 4 exposure sessions, a diminished response during 32 sessions,
and no response during 35 sessions. Rats in the group exposed to 24,101 mg/m3 had a
diminished response in 3 exposure sessions and no response in 68 sessions. These effects were
interpreted as a compound-related sedative effect. There were no compound-related effects on
alerting during exposure to 1,721 mg/m3.
Immediately following exposure, female rats in the 6,886 and 24,101 mg/m3 exposure
groups and males in the 24,101 mg/m3 group had an increased incidence of stained chin.
Although this clinical observation was dose-related, the toxicological significance is not clear.
This observation was characterized as transient by the study authors.
-------
Functional observational battery assessments were conducted at some time prior to motor
activity testing. No significant treatment-related effects were found for either sex in any of the
34 functional observational battery parameters evaluated at any of the exposure concentrations.
There were no compound-related effects on forelimb or hind limb grip strength or hind limb foot
splay.
Motor activity was measured in 10-minute intervals for a session total of 60 minutes,
using automated Coulbourn activity monitors. Data were evaluated as the group mean total
motor activity counts and the mean duration of activity. The 6,886 mg/m3 male group showed
significant effects such as decreased duration of movement during the second 10-minute interval
at week 8. At week 13, there was a decrease in duration of movement and mean number of
movements over the third and fourth interval, and there was an overall decrease in mean duration
and mean number of movements. However, these effects were not dose-related.
Of the 12 rats used for functional observational battery and motor activity, 6 rats/sex of
the high-concentration (24,101 mg/m3) and control groups were randomly selected for perfusion
and neuropathology. Brain and spinal cord sections were processed by routine
neuropathological techniques for paraffin embedment. Sections of sciatic and tibial nerves,
dorsal and ventral roots, and dorsal root ganglia were processed, embedded, and stained using
standard procedures. Microscopic evaluation revealed no morphological differences from
control rats. There were no compound-related microscopic observations in this study.
In summary, no statistically significant compound related effects on functional
observational battery, motor activity, or neuropathology measures were found following
exposure to any concentration (1,721, 6,886, and 24,101 mg/m3) tested in the 90-day study of
rats (DuPont HLR, 1996d; Malley et al., 2000).
4.4.3. Genotoxicity
In comparison with controls, cyclohexane did not alter the number of revertant colonies
with or without exogenous metabolic activation in Salmonella/microsome tests using five
strains of Salmonella typhimurium (HLA, 1982d; McCann et al., 1975; Mortelmans et al., 1986).
32
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Cyclohexane also did not alter the number of benzo(a)pyrene-induced revertants in S.
typhimurium strain TA100 (Maron et al., 1981). The incidence of forward mutations was
increased in cultured L5178Y mouse lymphoma cells in the presence of exogenous metabolic
activation, but the increase was not dose-related at exposure levels up to that which inhibited cell
growth (HLA, 1982e). A second study showed no increase in mouse lymphoma cell forward
mutations either in the presence or the absence of exogenous metabolic activation (Litton
Bionetics, 1982). No increase in the number of sister chromatid exchanges either in the presence
or the absence of exogenous metabolic activation was observed in cultured Chinese hamster
ovary cells at exposure levels up to that which inhibit cell growth (HLA, 1982f). The rate of
DNA synthesis in cultured human lymphocytes, as measured by [3H]thymidine uptake, was not
affected by cyclohexane either in the presence or the absence of exogenous metabolic activation
(Perocco et al., 1983). Equivocal results were obtained in E. coli in the DNA cell binding assay
(an assay of the binding of DNA to cellular protein mediated by active carcinogenic chemicals)
(Kubinski et al., 1981). No significant increase in chromosome structural aberration frequency
was observed in bone marrow cells of male or female rats exposed by inhalation for 5
consecutive days to levels of cyclohexane up to 1,000 ppm (Litton Bionetics, Inc., 1981).
4.5. SYNTHESIS AND EVALUATION OF MAJOR NONCANCER EFFECTS AND
MODE OF ACTION
Existing human studies are inadequate to determine the toxicity of cyclohexane in
humans. There have been two small studies published recently in which workers were exposed
primarily to cyclohexane. Although very few adverse health effects associated with cyclohexane
exposure were reported, definitive conclusions cannot be drawn due to limitations of both
studies. No chronic studies of toxicity of cyclohexane in animals were located.
In a well-conducted two-generation study of rats (DuPont HLR, 1997a; Kreckmann et al.,
2000), cyclohexane exposure of dams was associated with low pup weights in both the Fl and
F2 litters. Pup weights were significantly reduced from lactation days 7 through 25. Body
weights in adult PI and Fl females and Fl males were also significantly reduced. Rats in both
the two-generation study (DuPont HLR, 1997a; Kreckmann et al., 2000) and a developmental
33
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toxicity study (DuPont HLR, 1997b) exhibited maternal toxicity in the form of decreased body
weight. On the other hand, rabbits in DuPont HLR (1997c) did not exhibit signs of maternal
toxicity at the same doses (Kreckmann et al., 2000).
In the 90-day inhalation toxicity studies of rats and mice (DuPont HLR, 1996a, b; Malley
et al., 2000), the animals exhibited liver changes, including increased relative liver weights,
hepatocellular hypertrophy, and changes in liver enzyme profiles. The study authors argued that
these liver changes were adaptive responses, associated only with the high dose and were not
indicative of cyclohexane toxicity. However, not all of the changes were reversible in male rats.
In the absence of longer-duration studies (6 months, 1 year, or lifetime), it cannot be concluded
that these liver changes were not the first indications of systemic toxicity.
Observations indicating central nervous system effects have been noted in many of the
studies. Acute studies in rats, mice, guinea pigs, rabbits, and cats exposed by inhalation showed
trembling, sedation, and other neurological effects. In 90-day inhalation studies (Malley et al.,
2000), diminished response to an alerting signal was observed in rats while in the exposure
chambers (DuPont HLR, 1996b) while hyperactivity and diminished response to an alerting
response were observed in mice (DuPont HLR, 1996a). However, in an acute study (DuPont
HLR, 1996c; Christoph et al., 2000) and a 90-day study of neurotoxicity (DuPont HLR, 1996d;
Malley et al., 2000) in adult rats, no statistically significant changes were detected. In the 90-
day neurotoxicity study, as in other studies of that duration, clinical observations of a diminished
response to an alerting stimulus were detected.
No proposed mechanisms of action were found to explain the observed toxic effects.
Extrapolating from in vitro studies of membrane disruption by cyclohexane, it has been
postulated that cyclohexane and other solvents may cause central nervous system effects via
disruption of ion balances or membrane proteins in neurons (Naskali et al., 1993, 1994).
Studies at the cellular and subcellular level report effects of cyclohexane on membrane
functions, enzyme kinetics, and metabolic regulation of the cell. In several studies, Naskali et al.
(1993, 1994) and Tahti and Naskali (1992) demonstrated in vitro that cyclohexane disturbs
ATPase-dependent astrocytic regulation of ion balance in the neuronal environment. Initial
34
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studies using whole-brain reaggregate or granule cell cultures established that cyclohexane
treatment produced concomitant changes in the neural membrane fluidity and integral enzyme
activity. Subsequently, the activity of astrocytic membrane-bound ATPase was measured using
isolated neural membranes from primary astrocyte cultures of neonatal Sprague-Dawley rats.
Incubation with 3, 6, or 9 mM cyclohexane caused significant enzyme inhibition in a dose-
dependent manner (up to 18% of control activity at 9 mM concentration). ATPase inhibition
was greater than previously found in whole-brain or granule cell cultures. The level of inhibition
caused by cyclohexane was similar to the inhibition caused by hexane but much smaller than that
caused by other industrial solvents such as benzene or toluene. The authors suggested that
astrocytes are very sensitive to cyclohexane and other solvents that affect the central nervous
system.
In vitro, cyclohexane impaired the activity of phospholipid liposomes and reconstituted
cytochrome c oxidase proteoliposomes from Escherichia coll (Sikkema et al., 1994).
Partitioning of cyclohexane from the membrane surface to the bilayer center caused swelling and
increased fluidity in a concentration-dependent manner, disrupting the proton motive force (the
electrical potential) and the pH gradient across the membrane and increasing the permeability to
ions and low molecular weight compounds. The cumulative result was inhibition of cytochrome
c oxidase activity and dissipation of the energy-transducing properties of the membrane. The
authors suggested that affecting the proton motive force across the membrane is a critical part of
the effects of cyclohexane on membranes and membrane-embedded proteins. The results
support the findings of Uribe et al. (1990) on intact yeast cells and isolated mitochondria
(Saccharomyces cerevisiae\ where cyclohexane disrupted the permeability barrier of the inner
mitochondrial membrane.
4.6. WEIGHT-OF-EVIDENCE EVALUATION AND CANCER CHARACTERIZATION
Cancer studies in humans and inhalation or oral carcinogenicity assays using animals
were not located. The genotoxicity studies that were performed using cyclohexane are generally
negative. Under EPA's Proposed Guidelines for Carcinogen Risk Assessment (U. S. EPA, 1999),
35
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the available information on cyclohexane would be evaluated as "Data are inadequate for an
assessment of human carcinogenic potential."
4.7. SUSCEPTIBLE POPULATIONS
4.7.1. Possible Childhood Susceptibility
No specific data were located that address the relative sensitivity of children and adults to
the toxic effects of cyclohexane. A two-generation inhalation reproductive study of rats (DuPont
HLR, 1997a; Kreckmann et al., 2000) that examined the effect of cyclohexane exposure on
animals exposed in utero through sexual maturity found decreased pup weights during the
lactation period in both first- and second-generation offspring. Two well-conducted inhalation
studies in rats (DuPont HLR, 1997b) and rabbits (DuPont HLR, 1997c) showed no evidence of
developmental abnormalities in utero at doses that caused maternal toxicity (evidenced by
reduced body weights) in rats (Kreckmann et al., 2000).
People of all ages are probably exposed to cyclohexane in the ambient air. In addition,
nursing children may be exposed to cyclohexane in mothers' milk. In one very limited study,
cyclohexane has been detected in 5 of 12 samples of human milk taken from urban areas across
the United States, but no information about potential exposures of the women donating the
samples is available (Pellizzari et al., 1982). There is also a lack of adequate data from animal
studies to evaluate the potential toxic effects of oral exposure to cyclohexane.
Although no neurotoxicity or impaired response caused by the inhalation of cyclohexane
was found in acute (DuPont HLR, 1996c; Christoph et al., 2000) and 90-day neurotoxicity tests
of adult rats (DuPont HLR, 1996b; Malley et al., 2000), clinical observations of changed
responses to an alerting signal were seen in adult rats and mice; therefore, developmental
neurotoxicity is an area of concern for childhood susceptibility where a data gap exists.
4.7.2. Possible Gender Differences
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The limited occupational studies available provide no data to suggest that gender
differences in toxicity might occur as a result of exposure to cyclohexane. In animal studies,
some gender differences were noted. In 90-day studies, both female and male mice exhibited
significantly increased relative liver weights (DuPont HLR, 1996a; Malley et al., 2000).
Although both male and female rats exhibited hepatocellular hypertrophy, only male rats
exhibited significantly increased relative liver weights (DuPont HLR, 1996d; Malley et al.,
2000).
Females in general or pregnant females in particular may be more susceptible than males
to decreased body weights after inhalation exposure to cyclohexane. A two-generation
inhalation reproductive study of rats (DuPont HLR, 1997a; Kreckmann et al., 2000) examined
the effect of cyclohexane exposure and found significantly decreased adult body weights in first-
and second-generation treated females (PI and Fl) but only in Fl adult males. The deficit in the
females' body weight occurred in the premating phase of the study, and body weight gain was
not significantly different than controls' during gestation or lactation. In prenatal developmental
toxicity studies, inhalation exposure also produced decreased body weight in rat dams but no
similar responses in rabbits (DuPont HLR, 1997b, c; Kreckmann et al., 2000).
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5. DOSE-RESPONSE ASSESSMENTS
5.1. ORAL REFERENCE DOSE (RfD)
No adequate studies for the derivation of an RfD were located. Available information
was inadequate for a route-to-route extrapolation from the inhalation pathway to the oral
pathway.
5.2. INHALATION REFERENCE CONCENTRATION (RfC)
5.2.1. Choice of Principal Study and Critical Effect—with Rationale and Justification
There are no adequate human studies and no chronic or lifetime animal studies available
to determine the RfC. The two-generation reproduction inhalation toxicity study of rats (DuPont
HLR, 1997'a; Kreckmann et al., 2000) was chosen as the principal study. In this study, one set of
litters was produced in each generation. Cyclohexane exposure of dams was associated with low
pup weights—the chosen critical effect—in the Fl and F2 litters. Pup weights were significantly
reduced from lactation days 7 through 25 (Tables 5-1 and 5-2), and this was chosen as the
critical effect. Significant decreases in body weight in young animals can be associated with
developmental delays and lifelong mental and physical deficiencies (U.S. EPA, 1991; Hood,
1996).
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Table 5-1. Mean pup weights (G)—Fl generation
Maternal Group Concentration
(mg/m3)
Day 0
Day 4 preculling
Day 4 postculling
Day 7
Day 14
Day 21
Day 25
0
6.7
11.0
11.0
16.2
30.0
48.5
67.5
1,721
6.7
11.0
11.0
16.2
29.9
48.5
67.8
6,886
6.7
11.2
11.3
16.3
29.7
48.3
68.3
24,101
6.6
10.6
10.6
15. la
26.5 a
43. la
62.2 a
a Statistically significant difference from control (p < 0.05).
Source: Adapted from DuPont HLR, 1997a.
Table 5-2. Mean pup weights (G)-F2 generation
Maternal Group Concentration
(mg/m3)
Day 0
Day 4 preculling
Day 4 postculling
Day 7
Day 14
Day 21
Day 25
0
6.4
10.8
10.9
16.3
31.0
50.0
69.3
1,721
6.6
10.8
10.8
16.0
30.2
48.3
67.1
6,886
6.3
10.1
10.1
15.3
28.9
46.4
65.6
24,101
6.3
10.2
10.1
14.3a
26.2a
42.8a
61. 3a
a Statistically significant difference from control (p < 0.05).
Source: Adapted from DuPont HLR, 1997a.
Adult PI and Fl females and Fl males also had significantly reduced body weights. In
adults, however, reduced body weights were less severe (generally less than 10% difference
from controls) and in some cases appeared reversible, as affected females increased body weight
gains over controls in later stages of the study (DuPont FtLR, 1997a; Kreckmann et al., 2000).
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The two-generation and developmental toxicity studies of rats exhibited maternal toxicity
in the form of decreased body weight, and decreased body weight gain in the developmental
study (DuPont HLR, 1997a, b; Kreckmann et al., 2000).
Other effects noted in the principal study (decreased body weights in adult males and
females) and in the developmental study of rats (decreased body weights and body weight gains
in dams) and the 90-day rodent inhalation exposure studies (increased relative liver weights,
hepatocellular hypertrophy, and changes in liver enzymes) were not associated with pathological
changes in the liver. The clinical observation of diminished response to a sound stimuli while in
the exposure chamber, noted in most 90-day animal studies as the most sensitive endpoint, added
information to the qualitative assessment of the toxicity of cyclohexane but does not provide data
of the quality necessary for the quantitative estimation of an RfC (U.S.EPA, 1994b). These are
subjective observations (the observers know which treatment group they are observing), and are
made on a per group basis rather than an individual test animal basis (only a few animals in the
exposure chamber are visible when the chamber is hit with the rod to produce an alerting
stimulus) (Malley et al., 2000).
Guidance for extra-respiratory effects of category 3 gases (U.S. EPA, 1994b) was used
for the cyclohexane assessment. Category 3 gases typically induce extra-respiratory effects
because they are considered nonreactive (i.e., they have a low propensity for dissociation or
metabolism to reactive forms) in the respiratory tract and have relatively low water solubility,
which would promote rapid partitioning into the bloodstream and transport away from
respiratory tissues. Cyclohexane is classified as a category 3 gas for this assessment on the basis
of toxicological data from repeated exposure studies. The studies indicate that cyclohexane is
not appreciably reactive in biological tissues since no histological effects were observed in either
upper or lower respiratory tissues in animals at exposure concentrations that caused systemic
effects (DuPont HLR, 1997a, b; Kreckmann et al., 2000; Malley et al., 2000). Cyclohexane is
considered to be insoluble in water, and it is absorbed readily into the bloodstream from
inhalation exposures.
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5.2.2. Methods of Analysis—Benchmark Concentration Analysis
Adjusted exposure concentrations were calculated from all experimental exposure
concentrations. Prior to adjustment, all concentrations were converted from ppm to mg/m3
assuming cyclohexane acted as an ideal gas at 25°C and 760 mm Hg pressure, using the
following the example equation (U.S. EPA, 1994b):
BMCL (mg/m3) = BMCL (ppm) x MW (g/mole)/ 24.45 (L/g/mole)
BMCL (mg/m3) = BMCL (ppm) x 84.2/24.45
BMCL (mg/m3) = BMCL (ppm) x 3.443
Exposure concentrations for developmental and reproductive toxicity repeated-exposure
studies were duration-adjusted to provide estimated equivalent continuous exposure levels (U.S.
EPA, 1994b). Dams were exposed to cyclohexane in chambers for 6 hours/day, 7 days/week
(DuPont HLR, 1997a, b; Kreckmann et al., 2000). Therefore, experiment values were multiplied
by a factor of 6/24 (or I/ 4) to estimate 24-hour equivalent continuous exposure levels.
For extra-respiratory effects of category 3 chemicals, the guidance indicates that human
equivalent concentration (HEC) values are obtained from each adjusted exposure concentration
following the example equation (U.S. EPA, 1994b):
(mg/m3) = BMCL[ADJ] (mg/m3) x (Hb/g)A/(Hb/g)H,
where
the BMCL (or other exposure concentration) expressed in mg/m3,
dosimetrically adjusted for differences between humans and
animals in absorptivity of cyclohexane into blood;
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BMCL[ADJ] = the BMCL (or other exposure concentration) expressed in mg/m3,
adjusted for exposure schedule to estimate equivalent continuous
exposure concentration if appropriate; and
(Hb/g)A/(Hb/g)H = the ratio of blood/gas partition coefficients of cyclohexane for the
animal value to human value.
U.S. EPA (1994b) guidance indicates that the default value of the (Hb/g)A/(Hb/g)H ratio should be
set equal to 1 if the blood:air partition coefficient data are not available for either humans or
animals or if the value is greater than 1. Only one animal blood:air partition coefficient was
located (rat heparinized blood 1.39 ± 0.09 [Gargas et al., 1989]). Two averaged human values
were located in the literature (human heparinized blood 1.41 ± 0.14, Gargas et al., 1989 and 1.3
±0.1, Perbellini et al., 1985). Although the ratio of the rat value to the averaged human values
would be marginally greater than 1, a calculated value is not included since the available animal
and human values cannot be distinguished statistically. Therefore, the default value of 1 was
used and HEC values for cyclohexane were set equal to the duration adjusted exposure
concentrations expressed in mg/m3.
Reduced F2 pup weight gain during lactation from days 7 to 25 was modeled as the
critical effect (DuPont HLR, 1997'a; Kreckmann et al., 2000). All of the endpoints examined for
the basis of RfC calculation were weight related. Weight-related endpoints are attractive for
benchmark concentration (BMC) analyses because observations are not typically omitted by
design on any animals (contrary to, say, histopathology; even pups are ordinarily weighed prior
to any examination). Consequently, a model can be established throughout the range of dose
response. Because such a model gives a point of departure that is not obliged to coincide with an
experimental dose and is sensitive to the numbers of animals in a study, BMC analysis was
chosen as the basis for RfC calculation. Presuming this model to be parallel to human response
is implicit in using a BMC.
Because weight-related endpoints are continuous, however, decisions must be made
regarding the magnitude of change that will be associated with the BMC. None of the pup
weight endpoints for the multigeneration study has a typical magnitude identified as a level of
change for benchmark consideration. Consequently, following the recommendations of U.S.
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EPA's Benchmark Dose Technical Guidance Document (U.S. EPA, 2000c), change was
measured in units of standard deviations from the control mean. If, for example, the bottom
1-2% of the weight gain distribution among controls is regarded as a plausible magnitude of
change (whatever that weight gain is), then a shift downward of the mean weight gain by an
interval of one standard deviation from the control mean weight gain would shift about 10% of
the population into that range. In all instances, the benchmark response (BMR) was taken as one
standard deviation change from the control value. The BMC limit (BMCL) was taken as the
lower limit of a one-sided 95% confidence interval on the BMC.
Pup weights were available at all three test doses and control. All data were modeled as
their HEC values; the BMC and BMCL values are also in terms of their HECs. U.S. EPA's
Benchmark Dose Software (BMDS) Version 1.30 was used to establish a model and estimate a
BMC. Although individual pup data were available to help describe litter parameters, modeling
was carried out on the basis of litter averages because the current version of BMDS only
accommodates nested quantal data and the mean responses were used for benchmark
comparisons. The empirical curve of the data was not monotonic increasing, and thus a limited
number of continuous models was examined.
The model finally chosen was quadratic with constant variance for the F2 generation for
pup weight gain from days 7 to 25. Details appear in the tables and figures of Appendix B
(Tables B-l through B-3, Figure B-l). This model yielded a BMC(lsd) of 5,250.05 mg/m3 and a
BMCL(lsd) of 1,822.48 mg/m3 (the lowest BMCL(lsd) of the models for the Fl and F2 pup weight
gain).
Maternal body weight gain was among the endpoints examined, and it was found to
decrease in a statistically significant manner in the developmental study (DuPont HLR, 1997b;
Kreckmann et al., 2000) as well as in the multigeneration reproduction study (DuPont HLR,
1997a; Kreckmann et al., 2000). Species, strain, and nominal administered doses were the same
in the two studies, although the dosing regimen was not comparable. As with the pup weight
endpoints for the multigeneration study, neither of the maternal weight endpoints considered for
the developmental study (absolute body weight was also examined) has a typical magnitude
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identified as a level of change for benchmark consideration. Consequently, the BMR was taken
as one standard deviation change from the control mean.
Maternal weight gain from gestation days 7 to 17 and maternal body weight on gestation
day 17 were modeled for the developmental study using benchmark dose analyses of HEC
values for all test concentrations. The BMC(lsd) for the former, based on an unrestricted linear
model with constant variance, is 3,153.34 mg/m3 and its BMCL(lsd) is 2,500.82 mg/m3. The latter
had a BMC(lsd) of 6,654.66 mg/m3 and a BMCL(lsd) of 4,437.79 mg/m3, also based on an
unrestricted constant variance linear model. These results are within the range of the values
from the pup data of the reproductive toxicity study.
5.2.3. RfC Derivation—Including Application of Uncertainty Factors (UFs)
A factor of 3 (equivalent to approximately 101/2) was applied to account for interspecies
differences between humans and laboratory test animals. The factor for interspecies differences
has two components: pharmacokinetics and pharmacodynamics. In this assessment the
pharmacokinetic component was addressed by the calculation of the human equivalent
concentration (HEC) according to the RfC methodology for a category 3 gas (U.S. EPA, 1994a,
1994b, 2002). Accordingly, only the pharmacodynamic area of uncertainty remains as a partial
factor for interspecies uncertainty.
A factor of 10 was used to account for intraspecies variation among humans. Although
the RfC is based on a sensitive lifestage (developing offspring), the uncertainty factor is
appropriate because of the lack of any information on the range of responses in humans exposed
to cyclohexane.
A factor of 10 was also applied to account for database deficiencies. There is a lack of
long-term or chronic studies of animals in the data base available for deriving the RfC (U.S.
EPA, 1994b). The subjective clinical observation of altered response to an alerting stimulus by
adult mice and rats increases concern for developmental neurotoxicity, although specific
neurotoxicity testing of adult rats did not reveal significant changes (DuPont HLR, 1996a, b, c,
d; Christoph et al., 2000; Malley et al., 2000). Similarly, the increased liver size detected in
44
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mice and rats in 90-day studies (Malley et al., 2000), although not accompanied by pathological
changes in necropsy, may be early indications of changes that might progress to frank liver
toxicity with long-term exposure.
Consistent with EPA practice (U.S. EPA, 1991, 1996), an additional uncertainty factor
was not used to account for the extrapolation from endpoints in less-than-chronic studies to
chronic effects since developmental toxicity (reduced pup body weight during lactation) was
used as the critical effect. The developmental period is recognized as a sensitive lifestage where
exposure during critical developmental time windows may induce effects not caused by lifetime
adult exposure.
The resulting RfC calculated with the HEC BMCL(lsd) of 1,822.48 mg/m3 is 6 mg/m3:
RfC = 1,822.48 mg/m3 / 300 = 6 mg/m3
5.3. CANCER ASSESSMENT
No adequate cancer or chronic studies were located. No data were located regarding the
existence of an association between cancer and cyclohexane exposure in humans. The are no
adequate animal studies of cancer or of chronic duration by any exposure route. The
genotoxicity studies that have been performed are generally negative. Under EPA's draft cancer
guidelines (U.S. EPA, 1999), data are inadequate for an assessment of human carcinogenic
potential.
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6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION
OF HAZARD AND DOSE RESPONSE
6.1. HUMAN HAZARD POTENTIAL
Humans are likely to be exposed to cyclohexane through inhalation due to its presence in
gasoline vapors and crude oils as well as the use of purified cyclohexane in solvents,
insecticides, and plasticizers.
Cyclohexane is rapidly absorbed into the bloodstream via the lungs, gastrointestinal tract,
and skin. At higher inhalation doses, some cyclohexane is expired unchanged due to preferential
partitioning to the alveoli rather than to blood, where it has low solubility. In human cadaver
and animal in vivo studies, cyclohexane partitioned to the lipid-rich areas in the body, including
fat, brain, and liver. Cyclohexane has been detected in 5 of 12 samples of human milk taken
from across the United States, but no information about potential exposures of the women
donating the samples is available. In studies of workers, cyclohexane was excreted primarily via
expiration from the lungs and secondarily in urine.
Existing human occupational studies are inadequate to determine the toxicity of
cyclohexane in humans. No chronic toxicity studies of cyclohexane in animals were located.
In a well-conducted two-generation study of rats, cyclohexane exposure of dams was
associated with low pup weights during lactation in the Fl and F2 generations (DuPont HLR,
1997a; Kreckmann et al., 2000). This effect was chosen as the critical effect for calculating the
RfC. There is concern that if there is increased susceptibility of young animals, developmental
neurotoxicity could occur, but no testing was located.
The two-generation (DuPont HLR, 1997a) and developmental (DuPont HLR, 1997b)
toxicity studies of rats both demonstrated maternal toxicity in the form of decreased body weight
(Kreckmann et al., 2000). Although only statistically significant at the highest dose tested, mid-
dose dams contributed to the downward trend in body weight used in the benchmark dose
analyses.
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In the 90-day inhalation toxicity studies (Malley et al., 2000) of rats (DuPont HLR,
1996b) and mice (DuPont HLR, 1996a), animals exhibited liver changes including increased
relative liver weights, hepatocellular hypertrophy, and changes in liver enzyme profiles. The
authors stated that the changes were reversible adaptive responses associated only with the high
dose and not indicative of cyclohexane toxicity (Malley et al., 2000). However, not all of the
changes were reversible in male rats. In the absence of longer-duration studies (6 months, 1
year, or lifetime), it cannot be concluded that these liver changes are not first indications of
potential liver toxicity that would become apparent with longer exposure periods.
Clinical observations of an altered response to an alerting stimulus were noted in many of
the subchronic studies. In the 90-day rat study, authors (DuPont HLR, 1996b; Malley et al.,
2000) report transient sedation with diminished response to an alerting signal was observed
while the rats were in the exposure chamber. In the 90-day study of mice, both hyperactivity and
transient sedation were observed. However, these are subjective observations (the observers
know which treatment group they are observing), and are made on a per group basis not an
individual test animal basis (only a few animals in the exposure chamber are visible when the
chamber is hit with the rod to produce an alerting stimulus). Therefore while helpful in the
qualitative characterization of the toxicity of cyclohexane, the observations do not yield data of
adequate quality for use in quantitative assessment. Furthermore, in both acute (DuPont HLR,
1996c; Christoph et al., 2000) and 90-day (DuPont HLR, 1996d; Malley et al., 2000)
neurotoxicity studies in adult rats, no effects beyond the clinical observations of diminished
response to an alerting stimulus were detected.
6.2. DOSE-RESPONSE
The RfC of 6 mg/m3 was derived by dividing the HEC BMCL(lsd) of 1822.48 mg/m3 by
the product of uncertainty factors of 300, as described in Section 5.2. The benchmark dose is the
preferred approach because it incorporates data for all exposure concentrations tested instead of
only the NOAEL. Other effects noted in the principal study (decreased body weights in adult
males and females), the developmental study of rats (decreased body weights and body weight
gains in dams), and the 90-day rodent inhalation exposure studies (increased liver weights,
47
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hepatocellular hypertrophy, and changes in liver enzymes) may reflect transient changes and
appear less severe than the reduced pup weight during lactation.
Confidence in the principal inhalation study (DuPont HLR, 1997a; Kreckmann et al.,
2000) is high because it uses an adequate number of study animals and exposure levels to
evaluate an adequate set of endpoints. Confidence in the remainder of the inhalation toxicity
data base is low to moderate because although it is comprised of a number of well-designed 90-
day toxicity, neurotoxicity, and developmental toxicity animal bioassays, no data were available
for long-term or lifetime exposures or for developmental neurotoxicity. The database included
some evidence suggestive of neurological effects in occupationally-exposed humans, but these
subjects were exposed to mixtures of chemicals, including those more clearly demonstrated to
have such effects (w-hexane and toluene). Adult rats and mice exhibited altered responses to an
alerting stimulus at the mid-level and high doses tested in subchronic studies, indicating the
possibility of neurotoxicity. However, the observations were subjective (the observers knew
what dose group they were watching), the observations were not on an individual animal basis,
and no significant effects were detected in the neurotoxicity test battery conducted on adult rats.
Therefore, confidence in the RfC is low to moderate, reflecting primarily the lack of chronic
duration exposure and a lack of developmental neurotoxicity testing.
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U.S. EPA. (1991) Guidelines for developmental toxicity risk assessment. Federal Register
56(234):63798-63826.
54
-------
U.S. EPA. (1994a) Interim policy for particle size and limit concentration issues in inhalation
toxicity: notice of availability. Federal Register 59(206): 53799.
U.S. EPA. (1994b) Methods for derivation of inhalation reference concentrations and application
of inhalation dosimetry. EPA/600/8-90/066F.
U.S. EPA. (1995) Use of the benchmark dose approach in health risk assessment. EPA/630/R-
94/007.
U.S. EPA. (1996) Guidelines for reproductive toxicity risk assessment. Federal Register
61(212):56274-56322.
U.S. EPA. (1998a) Guidelines for neurotoxicity risk assessment. Federal Register
63(93):26926-26954.
U.S. EPA. (1998b) Science policy council handbook: peer review. Prepared by the Office of
Science Policy, Office of Research and Development, Washington, DC. EPA 100-B-98-001.
U.S. EPA. (1999) Guidelines for carcinogen risk assessment [review draft]. NCEA-F-0644, July.
Risk Assessment Forum, Washington, DC.
U.S. EPA. (2000a) Science policy council handbook: peer review. Second edition. Prepared by
the Office of Science Policy, Office of Research and Development, Washington, DC. EPA 100-
B-00-001.
U.S. EPA. (2000b) Science policy council handbook: risk characterization. Prepared by the
Office of Science Policy, Office of Research and Development, Washington, DC. EPA 100-B-
00-002.
U.S. EPA. (2000c) Benchmark dose technical guidance document. External review draft,
EPA/630/R-00/001, October, Office of Research and Development, Risk Assessment Forum,
Washington, DC.
U.S. EPA. (2000d) Supplementary guidance for conducting health risk assessment of chemical
mixtures. Office of Research and Development, Risk Assessment Forum, Washington, DC.
EPA/630/R-00/002.
U.S. EPA. (2002) A review of the reference dose and reference concentration processes.
December 2002. U.S. Environmental Protection Agency, Risk Assessment Forum, Washington,
DC. EPA/630/P-02/002F
Verschueren, K, ed. (1996) Handbook of environmental data on organic chemicals. 3rd ed. New
York: Van Nostrand Reinhold; p. 565-567.
55
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Yasugi, T; Kawai, T; Mizunuma, K; et al. (1994) Exposure monitoring and health effect studies
of workers occupationally exposed to cyclohexane vapor. Int Arch Occup Environ Health
65(5):343-350.
Yuasa, J; Kishi, R; Eguchi, T; et al. (1996) Concentrations of trichloroethylene and its
metabolites in blood and urine after acute poisoning by ingestion. Occup Environ Med
53(3): 174-179.
Zahlsen, K; Eide, I; Nilsen, AM; Nilsen, OG. (1992) Inhalation kinetics of C6 to CIO aliphatic,
aromatic and naphthenic hydrocarbons in rat after repeated exposures. Pharmacol Toxicol
71(2): 144-149.
56
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APPENDIX A: SUMMARY OF EXTERNAL PEER REVIEW
COMMENTS AND DISPOSITION
-------
APPENDIX A: SUMMARY OF EXTERNAL PEER REVIEW COMMENTS AND
DISPOSITION
The support document and IRIS summary for cyclohexane 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.
(1) General Comments
The three external reviewers offered editorial comments, all of which have been incorporated
into the text when feasible. Substantive scientific comments are addressed below.
A. Comment: Are additional data/studies recommended for inclusion?
One reviewer recommended an additional study, based on the abstract by Gupta, KP; Mehrotra,
NK. (1990) Mouse skin ornithine decarboxylase induction and tumor promotion by cyclohexane.
Cancer Lett 51:227-23 3.
Response: The study was retrieved and reviewed. Although the reviewer thought the study
detailed evidence that cyclohexane might have tumor promoting capabilities the study
concentrates on the induction of an enzyme associated with skin irritation (ornithine
decarboxylase) and only secondarily measures skin tumor promotion activity of cyclohexane
when applied after known chemical initiators and promoters. The cancer-promotion activity of
cyclohexane on mouse skin is not clearly demonstrated by this study. It appears that the
continual irritation caused by cyclohexane could cause the observed increase in tumors when
applied three times a week for months to shaved albino mouse skin, but this is not discussed by
the study authors. The mouse skin could have been dry and cracked, and skin wounding has
long been known to promote skin cancer, but no mention of the skin condition is made.
Furthermore, it appears there was no statistical analysis of the tumor data. The methodology
also appears deficient because there is no indication of randomization of the study animals or
A-l
-------
coding of skin samples ("blind" reading). The tumor classification, if conducted, was not
presented. There is a general statement that most tumors were benign. Although some control
groups were used, the most appropriate control of parallel treatment (initiator, promoter, and
acetone treatment in place of cyclohexane) was not used. Because of the shortcomings outlined,
the study does not add to the understanding of potential systemic carcinogenicity of cyclohexane
and was not added to the IRIS summary or supporting documentation.
B. Comment: For the RfD and RfC, has the most appropriate critical effect been chosen (i.e.,
that adverse effect appearing in a dose-response continuum)?
All reviewers agreed with EPA's selection of reduced pup weight as the critical effect for the
RfC. One reviewer was concerned about the concept of liver changes as adaptive. The reviewer
also stated concern that the central nervous system effects - as the most sensitive, although
transient - might indicate a potential risk to humans. The reviewer believes changes occur but
the lack of sensitivity of the currently available neurotoxicological testing methods makes these
changes undetectable.
Response: The language in the toxicological review and IRIS summary has been clarified to
convey that insufficient data are available to negate the potential liver toxicity and neurotoxicity
associated with chronic exposures to cyclohexane.
C. Comment: Has the noncancer assessment been based on the most appropriate study? This
study should present the critical effect in the clearest dose-response relationship. If not, what
other study (or studies) should be chosen and why?
All reviewers agreed with the conclusions reached by EPA on the selection of the most
appropriate study for the critical effect from the existing data base for cyclohexane. One
reviewer listed all the studies in the toxicological review and all NOAELs and LOAELs from the
studies and then stated that EPA's conclusion was "reasonable."
Response: EPA considers reduced pup weight during lactation an adverse developmental effect.
When evaluating the critical effect for cyclohexane, EPA considered effects from all available
subchronic and systemic toxicological studies of adequate quality. The reduced rat pup weight
in the reproductive toxicity study and the dam weight seemed to be the most relevant toxic
A-2
-------
endpoints in humans and illustrate some dose-response relationship. Reduced pup weight could
lead to lifelong developmental delays and deficiencies, whereas in some cases, adult weight loss
showed reversibility. As a result, no substantial changes are proposed to the current IRIS file as
a consequence of these comments. Wording was changed in the IRIS documents to make it
clearer that the potential for chronic liver effects or neurotoxicity in humans could not be ruled
out by the available animal studies.
D. Comment: Are there other data that should be considered in developing the uncertainty
factors (UFs) or the modifying factor? Do you consider that the data supports the use of
different (default) values other than those proposed?
The reviewers agreed with the UFs applied by EPA. One reviewer stated that the UFs were
reasonable to provide a "health-protective" RfC.
E. Comment: Do the confidence statements and the weight-of-evidence statements present a
clear rationale and accurately reflect the utility of the studies chosen, the relevancy of the effects
to humans, and the comprehensiveness of the data? Do these statements make sufficiently
apparent all the underlying assumptions and limitations of these assessments? If not, what needs
to be added?
Two reviewers agreed with the confidence statements and weight-of-evidence statements. One
reviewer did not directly comment.
F. Comment: Do you agree with the methods of analysis and the benchmark dose (BMD)
methodology/calculations that were used to evaluate dose-response data for the chosen critical
effects?
All reviewers agreed with the BMD methodology applied in order to derive an RfC.
(2) Chemical-specific Comments
A number of effects suggestive of adverse changes to liver and central nervous system were
observed in animals following subchronic repeated inhalation exposures at comparable or lower
exposure levels than the pup weight effects (identified as the critical effects in the data base of
A-3
-------
inhalation studies), but these effects did not show a clear, lexicologically relevant continuum of
severity and/or marked progression of response with increasing dose nor were there any
treatment-related corroborative gross pathologies or histopathological lesions. In the absence of
data from chronic exposure studies, these effects may be considered either adaptive, minimal, or
of uncertain relevance to effects in humans from chronic exposures.
A. Comment: Do you agree with that conclusion? Is sufficient rational given to support that
conclusion?
One reviewer agreed with the conclusions made by EPA that effects such as increases in enzyme
levels and hepatomegaly were possibly adaptive responses and not considered as critical events
clearly leading to potential adverse effects. Two other reviewers were in agreement with EPA
that the observed effects on the liver and central nervous system should not be used as the
critical effects in the data base of subchronic inhalation studies but thought wording should be
added that made it clear that data gaps such as the lack of chronic and developmental
neurotoxicity studies left concern for these endpoints. In particular, one reviewer objected to the
word "adaptive" that had been used by study authors and referred to as one possible explanation
by EPA.
Response: All three reviewers essentially agreed with EPA's decision not to use data on the
liver, adult body weights, or central nervous system as the basis for the critical effect because
these effects, taken either individually or in combination, did not show a clear, toxicological
continuum of severity and/or marked progression of response with increasing dose or any
treatment-related corroborative gross pathologies or histopathological lesions; as a result, they
were considered to be either "adaptive," minimal, or of uncertain relevance to effects in humans
from chronic exposures. For example, central nervous system-related effects were reported in
animal studies, but they occurred only while the animals were in the exposure chamber. In the
absence of histopathologies in nervous system tissues and the absence of other neurological
effects in the acute and 90-day neurotoxicity studies, the observed effects were considered to be
primarily transient responses with uncertain relevance to chronic effects in humans from long-
term exposure to cyclohexane. Similarly, effects that may be associated with changes in the
liver were also observed in animals but were lacking in any evidence of histopathology (for
further
discussion, see Section 4.5.2). In the Toxicological Review of Cyclohexane, realizing the lack
A-4
-------
of sensitivity of available toxicological test methods, data gaps were outlined (chronic studies
and developmental neurotoxicity), and a UF of 10 was applied for data base deficiencies. As a
result, no major changes are proposed to the current IRIS file as a consequence of these
comments. Some language was changed in the IRIS documents to clarify EPA's conclusions
and distinguish them from the study authors' conclusions.
A reviewer noticed a possible effect in the rat developmental toxicity study that had not been
noted in the study or the Toxicological Review. A paragraph detailing the reviewer's
observation was added.
OVERALL RECOMMENDATION
All reviewers stated that the document is acceptable with minor revisions.
A-5
-------
APPENDIX B: BENCHMARK DOSE CALCULATIONS
-------
APPENDIX B: BENCHMARK DOSE CALCULATIONS
U.S. EPA's Benchmark Dose Software (BMDS) Version 1.30 was used to establish a
model and estimate a benchmark concentration. Reduced F2 pup weight gain from days 7 to 25
was modeled as the critical effect (DuPont HLR, 1997c). Pup weights were available at all three
test doses and control. Although individual pup data were available to help describe litter
parameters, modeling was carried out on the basis of litter averages (HEC) because the current
version of BMDS only accommodates nested quantal data and the mean responses were used for
benchmark comparisons. The empirical curve of the data was not monotonic increasing, and
thus a limited number of continuous models was examined. Although, for instance, higher-order
polynomial models or a Hill model could have been tried, their shapes would not have given
plausible fits or been readily interpretable. All models were evaluated using likelihood ratio
goodness-of-fit tests. Selection among models was assisted by examining their Akaike
Information Criteria (AIC).
The models examined using the Fl generation pup weight gain data included the linear
with constant and heterogeneous variance, the quadratic with heterogeneous variance, and the
power model with heterogeneous variance. None had restrictions on the coefficients. All
instances except the power model were considered to adequately describe the data (p>0.05 on
"model fit" test). The AIC of the quadratic model was 427.54. Although the least AIC was
shown by the power model, its BMC was about 1,000 points greater than that of either linear
model, owing to its unusual shape (while the power was 9.9, the slope was essentially zero), and
the BMC limit (BMCL) computation failed. Between the two linear models, one with
heterogeneous variance was favored by the goodness-of-fit comparisons, but the constant had the
lower (426.71 versus 428.44) AIC. Nonetheless, BMDS currently is limited in the choices of
variance models it provides, and the model of the variance it could incorporate into the
benchmark model was inadequate to the form of heterogeneity in the data. Additionally,
computation of the BMCL curve failed to use the heterogeneous variance model, so the constant
variance model was selected for this data set. This model yielded a BMC(lsd) of 5755.67 mg/m3
and a BMCL(lsd) of 4,117.51 mg/m3. Results are shown in Tables B-l and B-2.
Modeling also was carried out using the pup weight gain from days 7 to 25 in the F2
generation. These data appeared strictly decreasing. Linear, quadratic, and cubic constant
variance models were examined; the constant variance assumption was not rejected in model
B-l
-------
fitting, but the cubic model was an overfit model. The BMC(lsd) from the linear model was
6,042.91 mg/m3 and the BMCL(lsd) was 4,165.97 mg/m3. The BMC(lsd) and BMCL(lsd) values
appear similar to those of the Fl generation. The BMC(lsd) from the quadratic model was
5,250.05 mg/m3 and the BMCL(lsd) was 1,822.48 mg/m3. Results are shown in Tables B-l and
B-2.
Dam weights of the PI and Fl and absolute pup weights of Fl and F2 generations were
also modeled. The results appear in Tables B-l and B-2.
B-2
-------
Table B-l. Summary of model outcomes, reproductive
toxicity/multigenerational study (DuPont HLR, 1997a)a
Form of model
PI - day71bwe
linear
Fl - avg d25pupwf
linear
linear
quadratic
Fl - pw d25-d7g
linear
linear
quadratic
power
F2 - avg d25pupw
linear
quadratic
F2 - pw d25-d7
linear
quadratic
cubic
Assumptions on
variance3
const, var.
const, var.
het. var.
het. var.
const, var.
het. var.
het. var.
het. var.
const, var.
const, var.
const, var.
const, var.
const, var.
AICh
866.97
488.55
492.30
489.53
426.71
428.44
427.54
425.50
437.14
438.67
391.09
392.70
394.04
Model fit?"
yes
het. var.
yes
yes
het. var.
yes
yes
poor
yes
yes
yes
yes
overfit
BMCcn^
7,556.88
6,113.36
6,078.26
6,166.96
5,755.67
5,682.17
6,058.50
6,021.64
5,815.77
4,687.20
6,042.91
5,250.05
5,943.87
BMCLC
5,079.5
4,303.23d
failed
5,176.92
4,117.51d
4,274.39
5,053.81
5,016.80
4,053.03d
1,751.17
4,165.97
l,822.48d
1,943.68
a The first column is the form of the model, the second is the assumptions on variance, const. = constant variance, het
= heterogeneous; i.e., different variances for different groups. Details of models, including functional form and
parameter estimates are shown in Table B-2.
b "Model fit?" designates a summary of several plausible ratio tests. See Table B-2 for additional considerations.
"Het. var." indicates a recommendation was made to fit a model with heterogeneous variances.
"Log(likelihood) for fit" is shown in Table B-2.
0 BMC designates Benchmark Concentration; BMCL designates lower limit on Benchmark Concentration based
on a 95% confidence limit obtained by profile likelihood methods. Parenthetically, the basis for the BMR
(Benchmark Response) is identified as 1.0 standard deviation from the mean (1 sd). In a number of instances,
the BMCL was provided at 1 sd, but no curve could be plotted because computation of the BMCL could not be
completed at some other values. In particular, BMCL computation failed at all values for the linear,
heterogeneous variance model for Fl average day 25 pup weight.
d Designates the recommended BMCL for this endpoint.
e day71bw = dam body weight on day 71 of study.
f avg d25pupw = average pup weight on day 25 after birth.
g pw d25-d7 = average pup weight gain from day 7 to day 25 after birth.
h AIC = Akaike Information Criterion.
B-3
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Tables of Standardized Residuals (Tables B-3a - B-3f)
Table B-3a. Fl: linear, constant variance model for average d25 pup weight
Dose (mg/m3)
0
430.3
1,722
6,025
N
28
26
27
27
Obs Mean
67.3
67.8
68.3
62.2
Obs Std Dev
7.26
4.60
5.91
4.66
Est Mean
68.3
67.9
66.7
62.7
Est Std Dev
5.72
5.72
5.72
5.72
ChiA2 Res.
-4.96
-0.52
7.63
-2.14
Table B-3b. Fl: linear, constant variance model for average pup weight gain d25-d7
Dose (mg/m3)
0
430.3
1,722
6,025
N
28
26
27
27
Obs Mean
51.1
51.6
52.0
47.1
Obs Std Dev
5.58
3.47
4.11
3.61
Est Mean
52.0
51.7
50.7
47.5
Est Std Dev
4.29
4.29
4.29
4.29
ChiA2 Res.
-5.7.0
-0.20
8.24
-2.34
Table B-3c. Fl: linear, heterogeneous variance model for average pup weight gain d25-d7
Dose (mg/m3)
0
430.3
1,722
6,025
N
28
26
27
27
Obs Mean
51.1
51.6
52.0
47.1
Obs Std Dev
5.58
3.47
4.11
3.61
Est Mean
52.2
51.9
50.8
47.4
Est Std Dev
4.49
4.44
4.29
3.80
ChiA2 Res.
-6.07
-0.67
8.10
-1.43
Table B-3d. Fl: quadratic, heterogeneous variance model
for average pup weight gain d25-d7
Dose (mg/m3)
0
430.3
1,722
6,025
N
28
26
27
27
Obs Mean
51.1
51.6
52.0
47.1
Obs Std Dev
5.58
3.47
4.11
3.61
Est Mean
51.3
51.6
51.9
47.1
Est Std Dev
4.33
4.37
4.42
3.69
ChiA2 Res.
-1.35
0.244
0.868
0.428
B-6
-------
Table B-3e. F2: linear, constant variance model for average pup weight gain d25-d7
Dose (mg/m3)
0
430.3
1,722
6,025
N
21
22
22
24
Obs Mean
53.1
51.1
50.3
47.0
Obs Std Dev
4.73
5.14
5.76
5.92
Est Mean
52.2
51.8
50.6
46.8
Est Std Dev
5.34
5.34
5.34
5.34
ChiA2 Res.
3.56
-2.72
-1.45
0.61
Table B-3f. F2: quadratic, constant variance model for average pup weight gain d25-d7
Dose (mg/m3)
0
430.3
1,722
6,025
N
21
22
22
24
Obs Mean
53.1
51.1
50.3
47.0
Obs Std Dev
4.73
5.14
5.76
5.92
Est Mean
52.5
51.8
50.0
47.0
Est Std Dev
5.33
5.33
5.33
5.33
ChiA2 Res.
2.09
-3.0
0.975
-0.0643
B-7
-------
Linear Model with 0.95 Confidence Level
70
68
64
62
60
Linear —
BMD Lower Bound —
1000
2000
3000
dose
4000
5000
6000
Figure B-la. Fl: linear, constant variance model for average d25 pup weight.
Linear Model with 0.95 Confidence Level
54
53
52
1 51
c
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(D
cc
i 49
O)
S 48
47
46
45
Linear
BMD Lower Bound --- =
BMDLj
0 1000 2000 3000 4000 5000
dose
6000
Figure B-lb. Fl: linear, constant variance model for average pup weight gain d25-d7.
B-8
-------
Linear Model with U.yb Confidence Level
0 1000 2000 3000 4000 5000
dose
6000
a BMCL computation failed for one or more point on the BMCL curve. The BMCL curve was not plotted.
Figure B-lc. Fl: linear, heterogeneous variance model for average pup weight gain d25-d7a
Polynomial Model with 0.95 Confidence Level
1000 2000 3000 4000 5000 6000
dose
' BMCL computation failed for one or more point on the BMCL curve. The BMCL curve was not plotted.
Figure B-ld. Fl: quadratic, heterogeneous variance model for average pup weight gain
d25-d7.
B-9
-------
Linear Model with 0.95 Confidence Level
56
54
CD 52
co
c
o
I 50
cc
cc
0)
48
46
44
Linear
BMfr Lower Bound
BMDLJ.
3MD '
0 1000 2000 3000 4000 5000 6000
dose
Figure B-le. F2: linear, constant variance model for average pup weight gain d25-d7.
Polynomial Model with 0.95 Confidence Level
56
54
CD 52
if)
C
O
I 50
ce
| 48
46
44
Polynomial
Btvtfr Lower Bound
MD
0 1000 2000 3000 4000 5000 6000
dose
Figure B-lf. F2: quadratic, constant variance model for average pup weight gain d25-d7.
B-10
-------
Tables B-4 - B-6 and Figure 2 refer to the results from the rat developmental study
(DuPont HLR, 1997b).
Table B-4. Summary of model outcomes, developmental study
Form of
model
dams - bw gd
17-gd 7e
linear
linear
dams - gd 17
body wtf
linear
Assumptions on
variance"
het. var.
const, var.
const, var.
AIO
503.64
500.065
610.19
Model
fit?"
const, var.
yes
yes
BMCcn^
3,284.19
3,153.34
6,654.66
BMCLcn ^
2,541.63
2,500.82d
4,437.79
a The first column is the form of the model, the second is the assumptions on variance, const. = constant variance.
het = heterogeneous; i.e., different variances for different groups. Details of models, including functional form
and parameter estimates are shown in Table B-5.
b "Model fit?" designates a summary of several plausible ratio tests. See Table B-5 for additional
considerations. "Const, var." indicates a recommendation was made to fita model with a constant variance.
Log (likelihood) for fit is shown in Table B-5.
c BMC designates Benchmark Concentration; BMCL designates lower limit on Benchmark Concentration based
on a95% confidence limit obtainedby profile likelihood methods. Parenthetically, the basis for the BMR
(Benchmark Response) is identified as 1.0 standard deviation from the mean (1 sd).
d Designates the recommended BMCL for this endpoint.
e bw gd 17-gd 7 = average dam body weight gain from gestation day 7 to gestation day 17.
f gd 17 body wt = average dam weight gestation day 17.
8 AIC = Akaike Information Criterion.
B-ll
-------
Table B-5. Summary of model fits, developmental study
Form of model
dams - bw gd 17-gd 7b
linear:
Y = beta_0 + beta_l*dose + e(ij)
linear (same as above)
dams - gd 17 body wtc
linear (same as above)
Assumptions on variance"
het. var.
const, var.
const, var.
log(likeliho od) for fit
0.748
0.807
0.374
df
2
2
2
p value
0.69
0.67
0.83
Parameters (standard error)
alpha = 6.47858(26.4514)
rho = 0.671852(1.01509)
beta_0 = 62.7782(1.42678)
beta_l = -0.00311345(4. 16781e-4)
alpha = 96.9084(14.5272)
beta_0 = 62.7953(1.38954)
beta_l = -0.00312184(4. 35222e-4)
alpha = 334.004(50.0693)
beta_0 = 333.049(2.57969)
beta_l = -0.00274631(8. 0799e-4)
a het.var. means the variance is modeled as Var(i) = alpha*mean(i)Arho. A means raised to the power of.
bbw gd 17-gd 7 = average dam body weight gain from gestation day 7 to gestation day 17.
c gd 17 body wt = average dam weight gestation day 17.
B-12
-------
Tables B-6a - B-6c. Standardized Residuals (DuPont HLR, 1997b)
Table B-6a. Linear, heterogeneous variance model for average dam weight gain, gd!7-gd7
Dose
(mg/m3)
0
430.3
1,722
6,025
N
21
22
23
23
Obs
Mean
64.2
60.1
57.2
44.2
Obs Std Dev
10.8
11.2
8.51
9.57
Est Mean
62.8
61.5
57.4
44.0
Est Std Dev
10.2
10.2
9.92
9.08
ChiA2 Res.
2.9
-2.72
-0.466
0.283
Table B-6b. Linear, constant variance model for average dam weight gain, gd!7-gd7
Dose
(mg/m3)
0
430.3
1,722
6,025
N
21
22
23
23
Obs Mean
64.2
60.1
57.2
44.2
Obs Std Dev
10.8
11.2
8.51
9.57
Est Mean
62.8
61.5
57.4
44.0
Est Std Dev
9.84
9.84
9.84
9.84
ChiA2 Res.
2.98
-2.84
-0.476
0.339
Table B-6c. Linear, constant variance model for average dam body weight, gd!7
Dose
(mg/m3)
0
430.3
1,722
6,025
N
21
22
23
23
Obs Mean
332
331
330
316
Obs Std Dev
23.4
17.4
14.8
18.5
Est Mean
333
332
328
317
Est Std Dev
18.3
18.3
18.3
18.3
ChiA2 Res.
-0.712
-1.15
2.5
-0.613
B-13
-------
Linear Model with 0.95 Confidence Level
70
65
S 60
c
cr
55
50
45
40
Linear
BMuT Lower Bound
BMDLJ iylD
2000 3000 4000 5000
dose
6000
Figure B-2a. Linear, heterogeneous variance model for average dam weight gain, gd!7-gd7.
(DuPont HLR, 1997V)
Linear Model with 0.95 Confidence Level
70
65
60
o
Q.
(fl
0)
O.
| 50
45
40
Linear
Blvtfr Lower Bound
BMDL|
MD
1000 2000 3000 4000 5000 6000
dose
Figure B-2b. Linear, constant variance model for average dam weight gain, gd!7-gd7.
(DuPont HLR, 1997b)
B-14
-------
Linear Model with 0.95 Confidence Level
345
340
335
0)
I 330
Q_
0)
0) Q.-,r-
ce 325
c
CT3
| 320
315
310
305
Linear
BMDLJ
0 1000 2000 3000 4000 5000 6000 7000
dose
15:1012/272001
Figure B-2c. Linear, constant variance model for average dam body weight, gd!7.
B-15
------- |